Errata Sheet for 

Performance Assessment for the Tank Farm Facility 

at the Idaho National Engineering and Environmental Laboratory 

DOE/ID-10966

Section 2.1.6.2, “Grout Kd Values,” page 2-55 

This section states that “Concrete Kd values for reducing conditions have been chosen.” This is true for 

the tank contents; however, the concrete vault was assigned Kd values for oxidizing conditions. The 

section further states that “The concrete Kd values were chosen from Region II of Bradbury and Sarott 

(1995).” This is also true; however, in Region II, the reducing and oxidizing Kd values are the same for all 

elements except technicium and uranium. The technicium Kd values for reducing and oxidizing conditions 

are 1 and 0.001 m3

/kg, respectively. The uranium Kd values for reducing and oxidizing conditions are 5 

and 2 m3

/kg, respectively. 

TFF PA Results for 14C 

The projected doses for 14C are extremely low in the final performance assessment (PA). 14C was 

originally modeled in error with a low Kd value for grout in the draft PA report. Therefore, 14C was 

maintained in the final PA analyses to show continuity with the draft reports; however, the Kd value was 

corrected for grout and the 14C doses were significantly reduced in the draft final PA report. 

Revised Table 3-6, page 3-34 

Table 3-6 now includes data for 129I and the Kd concrete vault values. 

Table 3-6. Non-decay chain DUST-MS input parameters. 

Nuclide 

Half-life 

(yr) 

Kd

Sand Pad 

(m3

/kg) 

Kd

Grout 

(m3

/kg) 

Kd

Concrete Vault 

(m3

/kg) 

14C 5.73E+03 5.00E−03 5.00E+00 5.00E+00 

129I 1.57E+07 1.00E−03 8.00E−03 8.00E−03 

90Sr 2.86E+01 1.50E−02 3.00E−03 3.00E−03 

99Tc 2.13E+05 1.00E−04 2.50E+00 1.00E−03 
Revised Table 7-3, page 7-5 

Table 7-3 now includes data for the sand pad sorption and concrete vault sorption coefficients. 

Table 7-3. Overview of the parameter values for sensitivity and uncertainty analysis. 

Measured 

Item Best Scenario 

Realistic 

Scenario 

Conservative 

Scenario Worst-Case Scenario 

Solid 

Radionuclide 

Inventory 

50% reduction 

from worst 

case 

25% reduction 

from worst 

case 

10% reduction 

from worst case 

Depicts sodium-bearing 

waste (undiluted tankheel residual) 

Liquid 

Radionuclide 

Inventory 

95% reduction 

from worst 

case 

80% reduction 

from worst 

case 

50% reduction 

from worst case 

Depicts sodium-bearing 

waste (undiluted tankheel residual) 

Infiltration 1.1 cm/yr 1.1 cm/yr 4.1 cm/yr 12.4 cm/yr 

C 10 10 5.0 1.0 

I 0.03 0.03 0.008 0.002 

Sr 0.006 0.006 0.003 0.001 

Grout Sorption 

Coefficients 

(m3/kg) 

Tc 5 5 2.5 1 

C 0.005 0.005 0.005 0.005 

I 0.001 0.001 0.001 0.001 

Sr 0.015 0.015 0.015 0.015 

Sand Pad 

Sorption 

Coefficients 

(m3/kg) 

Tc 0.0001 0.0001 0.0001 0.0001 

C 10 10 5 1 

I 0.03 0.03 0.008 0.002 

Sr 0.006 0.006 0.003 0.001 

Concrete Vault 

Sorption 

Coefficients 

(m3/kg) 

Tc 0.001 0.001 0.001 0.001 

Unsaturated Sediment 0.52 0.52 0.29 0.052 

Zone 

Longitudinal 

Dispersivities 

(m) 

Basalt 3.36 3.36 1.85 0.34 

Unsaturated Sediment 0.26 0.26 0.14 0.026 

Zone 

Transverse 

Dispersivities 

(m) 

Basalt 1.7 1.7 0.94 0.17 

C 20 20 10 2 

I 5 5 0.1 0.01 

Sr 24 24 18 12 

Interbed 

Sediment 

Sorption 

Coefficients 

(mL/g) Tc 0.1 0.1 0.01 0 

C 7.1 7.1 5.0 1.7 

I 1 1 0.1 0 

Sr 13 13 6 1 

Basalt Sorption 

Coefficients 

(mL/g) 

Tc 0.24 0.24 0.01 0 
Inventory Error and Effect on Results 

An error in the inventory for liquids was found for 129I and 99Tc. This error consisted of a 34% reduction 

in the 99Tc liquid inventory of the realistic and best cases, and a 46% reduction in the 129I liquid inventory 

of the best case. Because the error reduces the total inventory, the projected doses for this inventory 

would also be proportionally reduced. The corresponding inventories for 129I and 99Tc have been corrected 

in Table A-2 (beginning on page A-12). The correct concentrations are shown in the table below. 

Table A-2. Liquid radionuclide inventory. 

Tank Source Inventory in 2016 

 

Liquids - Based on WM188 

Analytical Data of 4,989 L 

Conservative 

Scenario 

50% Reduction of 

Liquid 

Concentration 

(Through 

Flushing) 

Realistic Scenario 

80% Reduction of 

Liquid Concentration 

(Through Flushing) 

Best Scenario 

95% Reduction of Liquid 

Concentration 

(Through Flushing) 

Nuclide 

ORIGEN2 

or 

Analytical 

Ratio 

Ratio 

Factor 

Activity 

(Ci/L) 

Total

Ci (Ci/L) 

Total

Ci Ci/L 

Total 

Ci Ci/L 

Total 

Ci 

Tc-99 

 1.19E−05 5.96E−02 2.99E−06 1.49E−02 

I-129 

 7.44E−09 3.71E−05 
DOE/ID-10966 

Revision 1 

April 2003 

Performance Assessment for the 

Tank Farm Facility at the Idaho 

National Engineering and 

Environmental Laboratory 

Volume 1 

Sections 1 through 9
DOE/ID-10966

Revision 1

Performance Assessment for the Tank Farm Facility at 

the Idaho National Engineering and Environmental 

Laboratory 

Volume 1 

Sections 1 through 9 

April 2003 

Prepared for the 

U.S. Department of Energy 

Idaho Operations Office 
iii 

ABSTRACT 

This report documents the projected radiological dose impacts associated 

with the closure of the Tank Farm Facility at the Idaho Nuclear Technology and 

Engineering Center at the Idaho National Engineering and Environmental 

Laboratory. The performance assessment analysis is based on conservative 

assumptions and parameters. The calculations involved modeling (a) the release 

of radionuclides from the tanks, sand pads, and piping, and subsequent transport 

through the environment and (b) exposures to members of the public via air, 

groundwater, and food chain pathways. This performance assessment analysis 

includes responses to comments received during reviews by the Nuclear 

Regulatory Commission and the Department of Energy. Additional analyses have 

been performed on the Tank Farm Facility to further evaluate the expected 

performance. In addition, a complete sensitivity/uncertainty analysis has been 

conducted to ensure that the reported doses for compliance purposes are 

conservative. The results of the analyses indicate compliance with the established 

performance objectives and provide reasonable assurance that public health and 

safety will be protected. 
iv 
v 

EXECUTIVE SUMMARY 

Performance assessments (PAs) involve estimating radiological exposure 

to future members of the public. These analyses include comparing exposure 

estimates to specific performance measures to determine whether there is a 

reasonable expectation that the performance measures will be met in the future. 

This report documents the projected radiological dose impacts associated with 

the closure of the Tank Farm Facility (TFF) at the Idaho Nuclear Technology and 

Engineering Center (INTEC). INTEC is one of the facilities at the Idaho National 

Engineering and Environmental Laboratory (INEEL). The performance 

objectives and other performance-related factors required to be considered in this 

performance assessment are for deactivated high-level waste facility closure 

actions that contain residual low-level waste. This PA includes responses to 

comments received during reviews by the Nuclear Regulatory Commission and 

the Department of Energy. 

The potential risks posed by closure of the TFF were assessed by 

conservatively estimating the release and transport of radionuclides from (a) the 

release of radionuclides from the tanks, sand pads, and piping and subsequent 

transport through the environment and (b) exposures to members of the public 

via air, groundwater, and food chain pathways. Contaminated soils associated 

with the TFF are not addressed in the TFF closure or this PA. Instead, past 

releases of contaminants are being addressed under the Comprehensive 

Environmental Response, Compensation, and Liability Act (CERCLA). 

Contaminated soils are also addressed in the TFF Composite Analysis. 

Two receptor types were assessed. The first was a member of the public. 

During the operational and institutional control periods, it is assumed that this 

individual resides at the INEEL Site boundary. During the post-institutional 

control period, the member of the public resides at the INTEC facility. This 

receptor was used to evaluate the drinking water pathway, all pathways 

(including food chains), and air pathway doses for comparison to the 

performance objectives. 

The second type of receptor evaluated was an intruder. This hypothetical 

receptor was assumed to inadvertently intrude onto the TFF during the postinstitutional control period. Four intruder scenarios were evaluated in this PA. 

The first was an acute scenario, which consisted of exposure while developing a 

water well. In the well-drilling scenario, the receptor was exposed to 

contaminated drill cuttings spread over the ground. The second scenario involved 

an acute scenario where the intruder was assumed to excavate a basement. The 

third scenario was chronic exposure in an agricultural scenario based on drill 

cuttings from the development of the water well. The well cuttings were assumed 

to be spread over the ground, and the intruder then began farming activities. The 

fourth scenario was also a chronic exposure, which consisted of an agricultural 

scenario based on the excavation of contaminated soil for the development of a 

basement. 

The performance assessment process consists of conceptual models that 

link radionuclide inventory, release (or source term), environmental transfer, and 
vi 

impact assessment. The waste inventory used in the TFF PA was based upon 

bounding calculations, process knowledge, and sample analysis. The final step is 

determining radiological doses to receptors. 

The exposure pathways evaluated included the ingestion of contaminated 

water, ingestion of contaminated food, inhalation of contaminated airborne 

particulates, and external exposure to radionuclides in the air and on the ground 

(or soil) surface. Contaminated agricultural products consumed by members of 

the public also were considered. The analysis of the potential transport pathways 

from the TFF indicates that the groundwater pathway was the most significant in 

terms of radionuclide transport to the receptors. 

Several computer models were used in the TFF PA including DUST-MS, 

PORFLOW, GWSCREEN, and RadDecay. Each model was used for a specific 

analysis. DUST-MS was used to evaluate the release of radionuclides from the 

tanks, sand pads, and piping. PORFLOW was used to evaluate the transport of 

radionuclides through the groundwater pathway for the 1,000-yr compliance 

period. GWSCREEN was used to evaluate the groundwater transport of 

radionuclides over very long periods of time. RadDecay was used to evaluate the 

ingrowth of radionuclide progeny in the waste. 

In addition to computer codes, several analytical models were developed 

to evaluate the degradation of concrete vaults, stainless steel tanks, and piping. 

Multiplicative models also were developed to evaluate the doses to intruders and 

public receptors. 

As shown in the table below, the conservative analysis results indicate 

compliance with all performance objectives. 

Performance Objective PA Result 

All-pathways 

25 mrem/yr 

1.35 mrem/yr (groundwater pathway) 

0.51 mrem/yr (atmospheric pathway) 

Airborne emissions excluding radon 

10 mrem/yr 0.51 mrem/yr 

Average annual radon flux 

20 pCi/m2/s 0.39 pCi/m2

/s 

Protection of groundwater 

4 mrem/yr 0.77 mrem/yr 

Acute drilling scenario 

500 mrem 232 mrem 

Acute construction scenario 

500 mrem 0.80 mrem 

Chronic post-drilling scenario 

100 mrem/yr 91.1 mrem/yr 

Chronic post-construction scenario 

100 mrem/yr 26.1 mrem/yr 
vii 

To ensure that the sensitivities and uncertainties of the groundwater 

pathway modeling were thoroughly understood, a sensitivity/uncertainty analysis 

was conducted. A total of four scenarios were considered for the groundwater 

pathway analyses. These scenarios were termed (1) worst case, (2) conservative 

case, (3) realistic case, and (4) best case. The conservative case was used in the 

PA analyses to demonstrate compliance with the performance objectives. The 

worst-case scenario was used to evaluate the potential worst-case doses from the 

use of worst-case parameter values including the inventories and transport 

characteristics. The realistic scenario presents the analyses in relation to the 

expected behavior of the groundwater transport and the contaminant inventories 

expected to remain after cleaning of the tanks and piping. The best scenario uses 

the same transport parameter values as the realistic case; however, the inventory 

was further reduced to the level expected from the best available analysis of the 

cleaning technology. Such an evaluation was intended to demonstrate that the 

conservative case was, in fact, conservative and appropriate to use as the 

compliance case. 

The following table provides the results of the sensitivity/uncertainty 

analysis for the four groundwater pathway scenarios. 

This sensitivity/uncertainty analysis indicates that the performance objectives for 

the groundwater pathway are met for three cases (i.e., conservative, realistic, and 

best). However, the worst-case scenario doses are above the performance 

objectives. To assess the degree of uncertainty in the groundwater doses, the four 

groundwater scenarios were run for the matrix of possibilities: (1) inventory, 

(2) infiltration rate, and (3) transport parameters (i.e., distribution coefficients). 

This matrix resulted in 36 groundwater scenario combinations. 

The matrix analysis of parameter combinations resulted in a worst-case 

scenario considered to be a realistic “upper-bound” of the dose analysis. This 

combination was the worst-case transport parameters, conservative inventory, 

and a conservative infiltration rate. The conservative inventory is considered to 

be the upper bound of the final closure inventory based on preliminary data 

obtained from cleaning Tank WM-182. The cleaning process has been effective 

and the final inventories are expected to be less than those predicted in the 

conservative inventory calculations. The infiltration rate of 4.1 cm/yr is also 

considered to be the upper bound of the infiltration rate at the TFF. The TFF 

currently has a soil cover; the 4.1 cm/yr infiltration rate is based on 

measurements at the INEEL Central Facilities Area soil-covered landfill. Again, 

the infiltration rate will likely be reduced since an engineered cover is proposed 

Groundwater Scenario 

Drinking Water Dose 

(mrem/yr) 

All-Pathway Dose 

(mrem/yr) 

Worst Case 23.1 85.8 

Conservative Case 0.77 1.35 

Realistic Case 0.04 0.07 

Best Case 0.03 0.04 
viii 

for the TFF. However, the PA analyses do not take credit for this reduction for 

the compliance case. The realistic and best cases do include a reduction in the 

infiltration rate. The largest uncertainty in the analyses was considered to be the 

distribution coefficients for the TFF grout and the unsaturated zone. 

The combination of the worst-case distribution coefficients, conservative 

inventory, and conservative infiltration rate resulted in a realistic, “upper-bound,” 

worst-case dose of 3.4 mrem/yr for the drinking water pathway (i.e., groundwater 

protection) and a dose of 15 mrem/yr for the all-pathways dose. These doses are 

less than the performance objectives. In summary, the worst-case groundwater 

doses were found to exceed the performance objectives; however, a combination 

of parameters in the matrix analysis presents a realistic, “upper-bound” dose 

estimate for the worst-case that shows compliance with the performance 

objectives. 

Uncertainty in the atmospheric pathway, radon flux, and intruder analyses 

doses were also investigated based on the four inventories for the TFF (i.e., 

worst-case, conservative, realistic, and best). The following table summarizes 

the results of the uncertainty analysis. 

Scenario and Performance 

Objective 

Worst-Case 

Inventory 

Conservative 

Inventory 

Realistic 

Inventory 

Best 

Inventory 

Acute Drilling Intruder 

(500 mrem) 

276 232 193 144 

Chronic Drilling Intruder 

(100 mrem/yr) 

106 91.1 75.1 52.8 

Acute Construction Intruder 

(500 mrem) 

0.89 0.80 0.67 0.45 

Chronic Construction 

Intruder (100 mrem/yr) 

29.1 26.1 21.8 14.5 

Rn Flux 

(20 pCi/m2/s) 

0.48 0.39 0.29 0.19 

Atmospheric 

(10 mrem/yr) 

1.1 0.51 0.22 0.05 

The uncertainty analysis for the intruder, atmospheric releases, and radon 

fluxes indicate that the performance objectives are attained for all inventory 

scenarios except for the chronic intruder post-drilling scenario. The worst-case 

inventory is based on the assumption that the tank internals have been washed 

and the remaining heel is unaltered at a depth of 1 in. However, cleaning of Tank 

WM-182 has proven the effectiveness of the cleaning process, and the heel is 

much less is than 1 in. The residual radionuclide inventory in Tank WM-182 is 

less than any of the inventories presented in the analysis. Also, the intruder doses 

are the result of short-lived radionuclides remaining after the end of institutional 

control. The analysis assumed that the intruder would drill through the reinforced 

concrete vault at 100 years. Credit for the reinforced concrete vault deterring the 
ix 

intruder from drilling through the vault was not taken in the analysis. The doses 

are reduced dramatically as the time of intrusion is extended a few years. 

The sensitivity/uncertainty analysis indicates that the conservative case 

analyses provide a “bounding analysis” of the potential doses to the public from 

the closure of the TFF. The sensitivity/uncertainty analysis indicates that the 

potential for the doses exceeding the performance objectives are low, and that a 

combination of worst-case assumptions would be required to result in the doses 

exceeding the performance objectives. 
x 
xi 

CONTENTS 

ABSTRACT.................................................................................................................................................iii

EXECUTIVE SUMMARY .......................................................................................................................... v

ACRONYMS...........................................................................................................................................xxiii

NOMENCLATURE ...............................................................................................................................xxvii 

1. INTRODUCTION...........................................................................................................................1-1

1.1 General Approach .................................................................................................................1-4

1.1.1 Steps in Developing the Performance Assessment..................................................1-5

1.1.2 Conservative Parameters Used in Performance Assessment ...................................1-6

1.2 General Facility Description .................................................................................................1-7

1.2.1 Tank Farm Facility ..................................................................................................1-7

1.2.2 Closure Concept.......................................................................................................1-9

1.2.3 Waste Characterization and Certification Program ...............................................1-14

1.2.4 General Land Use Patterns ....................................................................................1-15

1.3 Related Documents .............................................................................................................1-15

1.4 Performance Criteria ...........................................................................................................1-17

1.4.1 Public Protection Performance Objectives ............................................................1-17

1.4.2 Water Resource Impact Assessment......................................................................1-18

1.4.3 Intruder Analysis ...................................................................................................1-23

1.4.4 ALARA Analysis...................................................................................................1-24

1.4.5 Summary of Performance Criteria.........................................................................1-24

1.5 Summary of Key Assumptions ...........................................................................................1-25

1.5.1 Radionuclide Inventory..........................................................................................1-25

1.5.2 Degradation Analysis.............................................................................................1-25

1.5.3 Groundwater Model...............................................................................................1-26

1.6 Summary .............................................................................................................................1-27

2. CLOSURE FACILITY AND SITE CHARACTERISTICS ...........................................................2-1

2.1 Site Characteristics................................................................................................................2-1

2.1.1 Geography and Demography...................................................................................2-1

2.1.2 Meteorology and Climatology...............................................................................2-12
xii 

2.1.3 Ecology..................................................................................................................2-14

2.1.4 Geology, Seismology, and Volcanology ...............................................................2-21

2.1.5 Hydrology..............................................................................................................2-28

2.1.6 Geochemistry.........................................................................................................2-49

2.1.7 Natural Resources..................................................................................................2-57

2.1.8 Natural Background Radiation ..............................................................................2-58

2.2 Principal Facility Design Features ......................................................................................2-60

2.2.1 Water Infiltration ...................................................................................................2-60

2.2.2 Disposal Unit Cover Integrity................................................................................2-60

2.2.3 Structural Stability .................................................................................................2-60

2.2.4 Inadvertent Intruder Barrier...................................................................................2-61

2.3 Waste Characteristics..........................................................................................................2-61

2.3.1 Tank Inventory.......................................................................................................2-61

2.3.2 Sand Pad Inventory................................................................................................2-67

2.3.3 Piping Inventory ....................................................................................................2-72

2.3.4 Residual Radionuclide Inventory in Piping ...........................................................2-74

3. ANALYSIS OF PERFORMANCE.................................................................................................3-1

3.1 Source Term..........................................................................................................................3-1

3.1.1 Radionuclide Inventory............................................................................................3-1

3.1.2 Engineered Barriers Degradation and Failure..........................................................3-4

3.2 Pathways and Scenarios ......................................................................................................3-14

3.2.1 Time Periods of Concern .......................................................................................3-14

3.2.2 Transport Pathways ...............................................................................................3-15

3.2.3 Exposure Scenarios................................................................................................3-19

3.2.4 Groundwater Transport Radionuclide Screening...................................................3-19

3.3 Models and Assumptions ....................................................................................................3-28

3.3.1 Source Term Model ...............................................................................................3-28

3.3.2 Groundwater Flow and Transport Modeling .........................................................3-35

3.3.3 Radon Transport Model.........................................................................................3-55

3.3.4 Volatile Radionuclide Model.................................................................................3-59

3.3.5 Dose Model............................................................................................................3-61

3.3.6 The All-Pathways Dose Assessment Model..........................................................3-62

4. RESULTS OF ANALYSIS.............................................................................................................4-1
xiii 

4.1 Groundwater Transport Source Term ...................................................................................4-1

4.1.1 Radionuclide Releases from the Piping, Sand Pad, and Tanks................................4-1

4.2 Groundwater Transport Results ............................................................................................4-2

4.2.1 General Transport Characteristics ...........................................................................4-3

4.2.2 Results for Individual Radionuclides.......................................................................4-5

4.3 Radon Analysis Results ........................................................................................................4-5

4.4 Volatile Radionuclide Analysis Results................................................................................4-6

4.5 Dose Analysis Results...........................................................................................................4-6

4.5.1 Groundwater Protection...........................................................................................4-6

4.5.2 All-Pathways Dose ..................................................................................................4-7

5. INADVERTENT INTRUDER ANALYSES..................................................................................5-1

5.1 Intruder Scenario Screening..................................................................................................5-1

5.1.1 Intruder Inventory....................................................................................................5-1

5.1.2 Acute Intruder Scenarios .........................................................................................5-2

5.1.3 Chronic Exposure Scenarios....................................................................................5-3

5.2 Acute Intruder-Drilling Scenario Definition.........................................................................5-4

5.3 Acute Intruder-Construction Scenario ..................................................................................5-9

5.4 Chronic Intruder Post-Drilling Scenario Definition............................................................5-19

5.5 Chronic Intruder Post-Construction Scenario Definition....................................................5-27

5.6 Intruder Analysis Results ....................................................................................................5-27

5.6.1 Acute Intruder-Drilling Scenario ...........................................................................5-27

5.6.2 Acute Intruder-Construction Scenario ...................................................................5-35

5.6.3 Chronic Intruder Post-Drilling Scenario................................................................5-35

5.6.4 Chronic Intruder Post-Construction Scenario........................................................5-35

5.7 Intruder Sensitivity/ Uncertainty Analysis..........................................................................5-35

6. INTERPRETATION OF RESULTS...............................................................................................6-1

6.1 All-Pathways Dose................................................................................................................6-1

6.2 Radon Flux Results ...............................................................................................................6-1

6.3 Volatile Radionuclide Releases.............................................................................................6-1
xiv 

6.4 Groundwater Protection ........................................................................................................6-1

6.5 Intruder Analysis Results ......................................................................................................6-2

7. PERFORMANCE EVALUATION.................................................................................................7-1

7.1 Comparison of Results to Performance Objectives...............................................................7-1

7.1.1 Analysis of Airborne Emissions ..............................................................................7-1

7.1.2 Analysis of Radon Flux ...........................................................................................7-2

7.1.3 Analysis of Groundwater Pathway Doses................................................................7-2

7.1.4 Analysis of Chronic Intruder Post-Drilling Scenario...............................................7-3

7.2 Use of Performance Assessment Results ..............................................................................7-4

7.3 Groundwater Pathway Sensitivity/Uncertainty Analysis......................................................7-4

7.3.1 Radionuclide Inventories .........................................................................................7-4

7.3.2 Sorption Coefficients ...............................................................................................7-6

7.3.3 Hydraulic Dispersivity Variations .........................................................................7-12

7.3.4 Infiltration Variations ............................................................................................7-13

7.3.5 Time of Degradation..............................................................................................7-14

7.3.6 Groundwater Pathway-Sensitivity and Uncertainty Analysis Results...................7-16

7.4 Sensitivity/Uncertainty in Hydraulic Parameters................................................................7-26

7.4.1 Vertical Conductivity Sensitivity Analysis............................................................7-26

7.4.2 Transport Resulting From Dam Failure Flood.......................................................7-26

7.4.3 Transmissivity in the Regional Aquifer.................................................................7-29

7.4.4 Big Lost River Seepage .........................................................................................7-29

7.4.5 Basalt Porosity.......................................................................................................7-29

7.4.6 Unsaturated Hydraulic Characteristic....................................................................7-29

7.5 Piping Inventory Uncertainty..............................................................................................7-32

7.6 Sand Pad Uncertainty..........................................................................................................7-32

7.7 DUST-MS Release Model Uncertainty ..............................................................................7-33

7.7.1 Heterogeneity in Tank Grout .................................................................................7-33

7.7.2 Comparison to Analytical Calculations .................................................................7-36

7.8 Radon Analysis ...................................................................................................................7-38

7.9 Volatile Radionuclide Analysis...........................................................................................7-38

7.10 Intruder Analyses ................................................................................................................7-39
xv 

7.11 Dose Analyses.....................................................................................................................7-40

7.12 Future Work ........................................................................................................................7-41

8. PREPARERS...................................................................................................................................8-1

9. REFERENCES................................................................................................................................9-1

Appendix A—INTEC Tank Farm Facility Closure Radionuclide Inventory ...........................................A-1 

Appendix B—Quality Assurance.............................................................................................................. B-1 

Appendix C—Detailed INTEC Facility Description ................................................................................ C-1 

Appendix D—ALARA Analysis ..............................................................................................................D-1 

Appendix E—Degradation Analysis of the Grouted Tank/Vault and Piping System at the Idaho 

Nuclear Technology and Engineering Center Tank Farm Facility and Preliminary Results 

for the Detailed Analysis of Releases from the Grouted Pipe and Encasement System ................ E-1 

Appendix F—Results of Previous Modeling Analyses of TFF Performance............................................F-1 

Appendix G—PORFLOW Input Files......................................................................................................G-1 

Appendix H—Derivation of the Radon Flux Mathematical Solutions.....................................................H-1 

Appendix I—CAP88-PC Output Files for Volatile Radionuclides ............................................................I-1 

Appendix J—All-Pathways FORTRAN Source Code .............................................................................. J-1 

Appendix K—Shielding Factor for 0.662 MeV Photons..........................................................................K-1 

Appendix L—Intruder Scenario Results................................................................................................... L-1 

Appendix M—Sensitivity/Uncertainty Graphs........................................................................................ M-1 

FIGURES 

1-1. Location of the INEEL, INTEC, and the TFF in relation to the surrounding region. ........................1-2

1-2. A map of the INTEC Tank Farm Facility. The numbers shown are building or structure 

numbers. ..........................................................................................................................................1-8

1-3. Vault and dome of Tank WM-185 showing the concrete beams and concrete risers on top............1-10

1-4. Monolithic octagonal vault for Tank WM-180.................................................................................1-10

1-5. Pillar and panel octagonal vault for Tank WM-183. ........................................................................1-11

1-6. Monolithic square vault for Tank WM-190......................................................................................1-11
xvi 

1-7. Simplified tank closure sequence. ....................................................................................................1-12

1-8. Typical cover layering. .....................................................................................................................1-14

2-1. Idaho map showing the location of the INEEL on the Eastern Snake River Plain.............................2-2

2-2. INEEL vicinity map............................................................................................................................2-3

2-3. Major facilities at INTEC. ..................................................................................................................2-4

2-4. Location of the TFF tanks and vaults at INTEC.................................................................................2-4

2-5. 2000 population distribution for the area surrounding INTEC (INEEL 1999b). ...............................2-7

2-6. 2010 projected population distribution for the area surrounding INTEC (INEEL 1999b).................2-8

2-7. 2020 projected population distribution for the area surrounding INTEC (INEEL 1999b).................2-8

2-8. Selected land use of the INEEL and surrounding vicinity................................................................2-10

2-9. Annual average wind direction and speed at INEEL meteorological monitoring stations...............2-13

2-10. Approximate distribution of vegetation at the INEEL (INEEL 1999b). ........................................2-17

2-11. General geologic map of the INEEL (Kuntz et al. 1994). ..............................................................2-22

2-12. North-south geological cross section..............................................................................................2-23

2-13. Seismicity map showing the ESRP in relation to the Intermountain seismic belt and the 

Centennial Tectonic belt (WCFS 1996). .......................................................................................2-26

2-14. Earthquake epicenters in the INEEL region, 1850–1995 (INEEL 1999b) .....................................2-27 

2-15. Volcanic rift zones at the INEEL....................................................................................................2-29

2-16. Surface water features of the INEEL and surrounding vicinity......................................................2-30

2-17. Probable maximum flood hydrograph (INEEL 1999b)..................................................................2-35

2-18. Probable maximum flood INEEL inundation map (INEEL 1999b)...............................................2-36

2-19. Flood hydrograph at INTEC...........................................................................................................2-38

2-20. Basalt moisture curve. ....................................................................................................................2-40

2-21. Upper perched water at INTEC. .....................................................................................................2-42

2-22. Lower perched water zone at INTEC. ............................................................................................2-45

2-23. Natural gamma and neutron log for Well USGS-51.......................................................................2-46

3-1. Flow chart for the PA analyses...........................................................................................................3-2
xvii 

3-2. 252Cf radionuclide decay chain............................................................................................................3-5

3-3. 249Cf radionuclide decay chain............................................................................................................3-6

3-4. 251Cf radionuclide decay chain............................................................................................................3-7

3-5. 250Cf radionuclide decay chain............................................................................................................3-8

3-6. Conceptualization of the degradation sequence for the representative tank, grout, vault, and 

piping system at the TFF. ..............................................................................................................3-10

3-7. Representative tank, vault, and grout showing the dimensions for the degradation analysis...........3-11

3-8. Base-case results for the percentage of concrete and grout degraded from sulfate and 

magnesium attack, carbonation, and calcium hydroxide leaching for Zones a (tank and 

grout), b (grout), and c (vault). ......................................................................................................3-13

3-9. Base-case results for percent grout loss for the maximum degradation from sulfate and 

magnesium attack, carbonation, and calcium hydroxide leaching for Zone d (piping).................3-13

3-10. Base-case results for the maximum percent of concrete and grout degraded from 

reinforcement corrosion, sulfate and magnesium attack, carbonation, and calcium 

hydroxide leaching for Zones a (tank and grout), b (grout), and c (vault). ...................................3-13

3-11. Base-case results of the hydraulic conductivity of Zone a (tank and grout), Zone b (grout 

inside vault and outside tank), and Zone c (vault).........................................................................3-13

3-12. Base-case results of the hydraulic conductivity of Zone d (piping). ..............................................3-14

3-13. Potential transport and exposure pathways from a near-surface facility (pathways selected 

for analysis are identified by italics)..............................................................................................3-16

3-14. DUST-MS conceptual model for release from the TFF vaults and tanks.......................................3-33

3-15. Illustration of the two-dimensional modeling slice used in PORFLOW........................................3-35

3-16. Model grid superimposed on geologic cross section for groundwater model. ...............................3-45

3-17. Model prediction of the hydrologic conditions illustrating the perched water zones resulting 

from the Big Lost River seepage. ..................................................................................................3-46

3-18. Model prediction of the hydrologic conditions illustrating the perched water zones resulting 

from the Big Lost River and percolation ponds seepage...............................................................3-48

3-19. The extent of upper perched water at the INTEC facility based on perched water well data 

in Table 3-7....................................................................................................................................3-51

3-20. Distribution of tritium resulting from discharge into percolation ponds (units in pCi/mL). ..........3-55

3-21. A diagram of the one-dimensional slab model...............................................................................3-57

3-22. A diagram of the 1-dimensional slab model with overburden........................................................3-59
xviii 

3-23. A block diagram of the all-pathways dose assessment and a description of where the 

parameter values enter the calculation...........................................................................................3-66

4-1. Releases of 90Sr, 99Tc, 129I, and 14C from the sand pad and tanks and piping. ....................................4-2

4-2. Groundwater modeling domain showing 99Tc concentrations and location of maximum 

concentrations (all concentrations based on a unit source inventory). ............................................4-4

4-3. Groundwater concentrations for the conservative analysis. ...............................................................4-4

4-4. Drinking water doses for the conservative analysis............................................................................4-7

4-5. All-pathway doses for the conservative analysis................................................................................4-8

5-1. Graphical representation of the acute intruder-drilling scenario. .......................................................5-6

5-2. Chronic intruder post-drilling scenario.............................................................................................5-19

5-3. Results of the acute intruder-drilling scenario..................................................................................5-33

5-4. Results of the acute intruder-drilling scenario for individual radionuclides.....................................5-34

5-5. Results of the acute intruder-drilling scenario by exposure pathway...............................................5-34

5-6. Results of the acute intruder-construction scenario for individual radionuclides.............................5-36

5-7. Results of the acute intruder-construction scenario by exposure pathway.......................................5-36

5-8. Results of the chronic intruder post-drilling scenario.......................................................................5-37

5-9. Results of the chronic intruder post-drilling scenario for individual radionuclides. ........................5-37

5-10. Results of the chronic intruder post-drilling scenario by exposure pathway..................................5-38

5-11. Results of the chronic intruder post-construction scenario for individual radionuclides. ..............5-38

5-12. Results of the chronic intruder post-construction scenario by exposure pathway..........................5-39

7-1. Longitudinal dispersivity graph (Mills et al. 1985). .........................................................................7-13

7-2. 129I drinking water doses for the four groundwater scenarios...........................................................7-19

7-3. 99Tc drinking water doses for the four groundwater scenarios. ........................................................7-19

7-4. 90Sr/90Y drinking water doses for the four groundwater scenarios. ..................................................7-20

7-5. 14C drinking water doses for the four groundwater scenarios...........................................................7-20

7-6. Sensitivity/uncertainty of dispersion coefficients on drinking water doses......................................7-22

7-7. Sensitivity/uncertainty of unsaturated zone Kd values on 129I and 90Sr/90Y drinking water 

doses ..............................................................................................................................................7-23 
xix 

7-8. Sensitivity/uncertainty of infiltration on 129I drinking water doses. .................................................7-24

7-9. Sensitivity/uncertainty of infiltration on 90Sr/90Y drinking water doses...........................................7-24

7-10. Sensitivity/uncertainty of grout Kd values on 129I releases. ............................................................7-25

7-11. Sensitivity/uncertainty of grout Kd values on 90Sr/90Y releases......................................................7-25

7-12. Sensitivity/uncertainty of inventory on 129I drinking water doses. .................................................7-27

7-13. Sensitivity/uncertainty of inventory on 90Sr/90Y drinking water doses...........................................7-27

7-14. Sensitivity of the moisture content distribution in the unsaturated zones with variations in 

vertical hydraulic conductivity values in the D-DE−2 interbed (Zone 9) of 0.025, 0.0025, 

and 0.00025 m/d. ...........................................................................................................................7-28

7-15. Comparison of radionuclide concentrations at the water table for non-flooding and flooding 

conditions. .....................................................................................................................................7-30

7-16. Extent of perched water due to non-flooding conditions and flooding conditions.........................7-31

7-17. Comparison of DUST-MS tank release rates and analytical calculations for the 129I Tank 

source using worst-case scenario parameters with release rates estimated assuming 

diffusion into concrete cracks with fracture flow..........................................................................7-35

7-18. Comparison of DUST-MS release rates with analytical calculations for the 129I tank source 

using worst-case scenario parameters............................................................................................7-37

TABLES 

1-1. Preliminary performance assessment components, objectives, and points of compliance...............1-18

1-2. Comparison of MCLs for radionuclides. ..........................................................................................1-20

1-3. Summary of groundwater protection performance objective options...............................................1-22

1-4. Summary of adopted performance objectives for the period of active institutional control.............1-24

1-5. Summary of adopted performance objectives for the post-institutional control period. ..................1-24

1-6. Summary of adopted performance objectives for inadvertent intruders...........................................1-25

2-1. Regional population of the INEEL; selected years 1990–2025..........................................................2-7

2-2. Greatest precipitation measured at the INEEL (regardless of location). ..........................................2-15

2-3. Flora at the INEEL Site. ...................................................................................................................2-16

2-4. Fauna at the INEEL. .........................................................................................................................2-19
xx 

2-5. Monthly discharge of the Big Lost River at Lincoln Boulevard near INTEC (ft3). .........................2-32

2-6. Characteristics of Mackay Dam and the INEEL Diversion Dam and reservoir. ..............................2-33

2-7. Results of probable maximum flood induced overtopping failure of Mackay Dam. .......................2-37

2-8. Estimated volume of water recharging the perched water bodies at INTEC....................................2-50

2-9. Summary of sorption coefficients measured at the INEEL. .............................................................2-51

2-10. Basalt Kd values from the NEA database. ......................................................................................2-51

2-11. Sorption coefficient ranges identified in WAG 3. ..........................................................................2-52

2-12. Overview of the radionuclide sorption coefficients (mL/g) used in transport modeling................2-52

2-13. Selection rationale for the range of Kd values (mL/g) used in the sensitivity analysis...................2-53

2-14. Solubility limits for the TFF grout..................................................................................................2-56

2-15. INEEL production wells and annual volume pumped....................................................................2-59

2-16. Conservative single tank inventory. ...............................................................................................2-65

2-17. Inventory for the sand pads at the time of closure (2016). .............................................................2-73

2-18. Radionuclide inventory for piping at 2016.....................................................................................2-76

3-1. Distribution coefficients for grout and sand. ....................................................................................3-22

3-2. Tank activity groundwater screening for the drinking water pathway. ............................................3-24

3-3. Sand pad activity groundwater screening for the drinking water pathway.......................................3-29

3-4. Radionuclides remaining from the initial screening based on waste pore-water 

concentrations................................................................................................................................3-32

3-5. Parameter values for source term analysis using the DUST-MS model...........................................3-34

3-6. Non-decay chain DUST-MS input parameters.................................................................................3-34

3-7. Perched water wells at INTEC. ........................................................................................................3-50

3-8. Perched layer, depth to water, and water table elevations for the wells encountering the lower 

perched zone (Rodriguez et al. 1997)............................................................................................3-54

3-9. Values of the gas transport parameters for the grouted heel along with the source of the 

information. ...................................................................................................................................3-58

3-10. The input parameters and their distributions used in these analyses. .............................................3-63

3-11. Parameter values used in the all pathway dose calculation in Maheras (1997)..............................3-65
xxi 

4-1. Peak releases rates, groundwater concentration, and doses for the groundwater pathway 

analyses. ..........................................................................................................................................4-3

5-1. Dose conversion factors used in the acute intruder-drilling scenario. ..............................................5-10

5-2. Parameter values used in the acute intruder-drilling scenario. .........................................................5-14

5-3. Unit concentration dose factors for acute intruder-drilling scenario. ...............................................5-15

5-4. Plant, beef, and milk uptake factors used in the chronic post-drilling scenario. ..............................5-23

5-5. Parameter values for the chronic post-drilling scenario. ..................................................................5-28

5-6. Unit concentration dose factors for the chronic post-drilling scenario.............................................5-29

7-1. Comparison of conservative results to the all-pathways, air emissions, and radon 

performance objectives....................................................................................................................7-1

7-2. Comparison of intruder results to the performance measures. ...........................................................7-2

7-3. Overview of the parameter values for the sensitivity/uncertainty analysis. .......................................7-5

7-4. Summary of sorption coefficients measured at the INEEL. ...............................................................7-8

7-5. Basalt Kd values from the NEA database. ..........................................................................................7-8

7-6. Sorption coefficient ranges identified in WAG 3. ..............................................................................7-8

7-7. Selection rationale for the range of Kd values (mL/g) used in the sensitivity analysis.......................7-9

7-8. Dispersivity values measured or calculated and reported in the literature (in meters).....................7-14

7-9. Summary of the degradation analysis, showing the time of failure for each component. ................7-15

7-10. Drinking water doses for sensitivity/uncertainty groundwater analyses. .......................................7-17

7-11. All-pathway doses for sensitivity/uncertainty groundwater analyses.............................................7-18

7-12. Comparison of the sand pad conceptual models.............................................................................7-33

7-13. Comparison of the sand pad drinking water doses for each conceptual model. .............................7-33

7-14. Doses for each inventory option for radon flux, atmospheric releases, and intruder 

scenarios. .......................................................................................................................................7-39
xxii 
xxiii 

ACRONYMS 

ALARA as low as reasonably achievable 

amsl above mean sea level 

bgs below ground surface 

BLM Bureau of Land Management 

BRA baseline risk assessment 

CEDE committed effective dose equivalent 

CERCLA Comprehensive Environmental Response, Compensation, and Liability Act 

CFA Central Facilities Area 

CFR Code of Federal Regulations 

CPP Chemical Processing Plant (now INTEC) 

DCF dose conversion factor 

DEQ (Idaho) Department of Environmental Quality 

DOE Department of Energy 

DOE-ID Department of Energy Idaho Operations Office 

DUST-MS Disposal Unit Source Term–Multiple Species 

EBR-I Experimental Breeder Reactor-I 

EDE effective dose equivalent 

EIS environmental impact statement 

EPA Environmental Protection Agency 

ESRP Eastern Snake River Plain 

FDF Flood-Diversion Facility (INEEL) 

FFA/CO Federal Facility Agreement and Consent Order 

FORTRAN Formula Translation 

HLLW high-level radioactive liquid waste 

HLW high-level waste 

HWMA Idaho Hazardous Waste Management Act 
xxiv 

ICRP International Commission on Radiological Protection 

IDAPA Idaho Administrative Procedures Act 

IDHW Idaho Department of Health and Welfare 

INEEL Idaho National Engineering and Environmental Laboratory 

INTEC Idaho Nuclear Technology and Engineering Center 

Ka kilo annum 

LLW low-level waste 

M magnitude 

Ma million annum 

MCL maximum contaminant level 

mpy mils per year 

MTR Materials Test Reactor 

MW monitoring well 

NEA Nuclear Energy Agencies 

NOAA National Oceanic and Atmospheric Administration 

NRC Nuclear Regulatory Commission 

OU operable unit 

PA performance assessment 

PMF probable maximum flood 

PMP probable maximum precipitation 

PW perched water 

RCRA Resource Conservation and Recovery Act 

RESL Radiological and Environmental Sciences Laboratory 

RI/FS remedial investigation/feasibility study 

RI/BRA remedial investigation/baseline risk assessment 

ROD record of decision 

RWMC Radioactive Waste Management Complex 
xxv 

SDA Subsurface Disposal Area 

TAN Test Area North 

TEDE total effective dose equivalent 

TFF Tank Farm Facility 

TRA Test Reactor Area 

TRU transuranic waste 

USC United States Code 

USGS United States Geological Survey 

WAG Waste Area Group 

WCC Woodward-Clyde Consultants 

WCFS Woodward-Clyde Federal Services 

WIR Waste-Incidental-to-Reprocessing 
xxvi 
xxvii 

NOMENCLATURE 

Ac actinium 

Ag silver 

Am americium 

At astatine 

Ba barium 

Be beryllium 

Bi bismuth 

C carbon 

Cd cadmium 

Ce cerium 

Cf californium 

Cl chlorine 

Cm curium 

Co cobalt 

Cr chromium 

Cs cesium 

Eu europium 

Fe iron 

Fr francium 

Gd gadolinium 

H hydrogen 

Hg mercury 

Ho holmium 

HTO tritiated water, occurs when a water molecule exchanges a hydrogen atom for a tritium atom 

I iodine 

In indium 
xxviii 

Kr krypton 

La lanthanum 

Nb niobium 

Nd neodymium 

Ni nickel 

Np neptunium 

Pa protactinium 

Pb lead 

Pd palladium 

Pm promethium 

Po polonium 

Pr praseodymium 

Pu plutonium 

Ra radium 

Rb rubidium 

Rh rhodium 

Rn radon 

Ru ruthenium 

Sb antimony 

Se selenium 

Sm samarium 

Sn tin 

Sr strontium 

Tc technetium 

Te tellurium 

Th thorium 

Tl thallium 
xxix 

Tm thulium 

U uranium 

Y yttrium 

Zr zirconium 
1-1

Performance Assessment for the Tank Farm Facility 

at the Idaho National Engineering 

and Environmental Laboratory 

1. INTRODUCTION 

In 1949, the Atomic Energy Commission, 

predecessor to the Department of Energy (DOE), 

established the National Reactor Testing Station, 

now known as the Idaho National Engineering and 

Environmental Laboratory (INEEL). Here, the 

Atomic Energy Commission built, tested, and 

operated various types of nuclear reactors, support 

facilities, and equipment. The INEEL became the 

setting for the world’s largest concentration of 

nuclear reactors. Fifty-two test reactors were built 

and tested. Of these reactors, only three are still 

operating. 

In the 1970s, laboratory work expanded into 

the fields of systems engineering, environmental 

remediation, waste management, technology 

development, and energy research. In 1974, the 

laboratory’s name was changed to the Idaho 

National Engineering Laboratory to reflect this 

new mission. The following year, the INEEL 

became the nation’s second largest national 

environmental research park. 

In 1997, DOE changed the laboratory’s name 

to the Idaho National Engineering and 

Environmental Laboratory to underscore projects 

in developing waste cleanup and other 

environment-related technologies. Today, the 

INEEL is a federally-funded research and 

development center emphasizing applied 

engineering. It is also DOE’s lead laboratory for 

environmental management. 

The INEEL occupies approximately 890 mi2

(2300 km2

)

a

 of desert in a rural, sparsely populated 

area of southeastern Idaho on the Eastern Snake 

 

a. The measurements for this report were originally made in 

U.S. customary units. Where conventional, we have converted 

these to SI units and preserved the correct number of 

significant digits. In some cases, numbers end with a period; 

this is to preserve the correct number of significant digits. 

Flow rates and specifications have not been converted. 

River Plain. The eastern boundary of the INEEL 

Site is 22 mi (35 km) west of Idaho Falls, Idaho. 

The Idaho Nuclear Technology and 

Engineering Center (INTEC) is one of the 

facilities at the INEEL. Prior to May 1998, INTEC 

was named the Idaho Chemical Processing Plant 

(CPP). INTEC covers approximately 200 acres 

(80 ha) and encompasses more than 150 buildings, 

including storage, treatment, and laboratory 

facilities for spent nuclear fuel, mixed high-level 

waste (HLW), and mixed transuranic waste. The 

employees at INTEC are focused primarily on four 

areas: 

• Safely storing spent nuclear fuel 

• Preparing spent nuclear fuel for shipment to an 

offsite repository 

• Developing technologies to safely treat highlevel and liquid radioactive waste 

• Remediating past environmental releases. 

With the exception of that at Argonne 

National Laboratory-West, all HLW at the INEEL 

exists at INTEC (Palmer et al. 1999). The HLW 

was created from 1953 until 1988, from 

reprocessing spent nuclear fuel at INTEC. The 

resulting radioactive waste exists in both liquid 

and solid forms. All of the liquid waste is stored in 

the INTEC Tank Farm Facility (TFF). The TFF is 

a collection of 15 belowground stainless-steel 

tanks. The eleven 300,000-gal tanks are enclosed 

in belowground concrete vaults, while the four 

30,000-gal tanks are directly buried in the soil. 

Figure 1-1 shows the relative locations of the 

INEEL, INTEC, and the TFF. The INEEL HLW 

Program manages all of the liquid waste stored at 

the TFF. 
1-2

Figure 1-1. Location of the INEEL, INTEC, and the TFF in relation to the surrounding region. 
1-3

Two types of liquid waste have been stored at 

the TFF: high-level liquid waste (HLLW), 

sometimes referred to as non-sodium-bearing 

waste, and sodium-bearing waste. The HLLW was 

generated as a direct result of reprocessing spent 

nuclear fuel; the sodium-bearing waste was 

generated from incidental activities, such as 

decontamination, associated with INTEC 

operations. 

In April 1992, then Secretary of Energy James 

Watkins announced that spent nuclear fuel 

reprocessing would halt and he called for the 

shutdown of reprocessing facilities at INTEC. In 

1988, the dissolution of nuclear fuel and creation 

of HLW ceased at INTEC. Since that time, no 

HLLW has been generated. The production of 

sodium-bearing waste is dependent on the type of 

work done at INTEC in the future. Sodiumbearing waste is generated primarily from 

decontamination and from operations associated 

with laboratories, fuel basins, and closure 

activities. 

Under the terms of a 1992 Consent Order (and 

subsequent modifications) between the Idaho 

Department of Health and Welfare (IDHW)b

 and 

DOE (IDHW 1992), DOE must permanently cease 

use of tanks in the TFF at the INEEL or bring the 

tanks into compliance with secondary containment 

requirements as set forth by the Idaho 

Administrative Procedures Act (IDAPA) 

58.01.05.009 (2002) (40 Code of Federal 

Regulations [CFR] 265.193, 2002). The Consent 

Order further specifies that this compliance cannot 

be achieved through an equivalency demonstration 

or by obtaining a variance as provided by IDAPA 

58.01.05.009 [40 CFR 265.193(d)(4) and (h)]. 

DOE plans to close the TFF tanks because high 

radiation fields would make compliance with 

secondary containment requirements difficult, and 

a need for such storage is not evident after 2012. 

 

b. On July 1, 2000, the Idaho legislature elevated the Division 

of Environmental Quality, a part of the Idaho Department of 

Health and Welfare, to the Idaho Department of 

Environmental Quality (DEQ). DEQ now oversees the 

implementation of the Consent Order. As such, citations and 

references to IDAPA 16 Title 01 (under IDHW), are now 

IDAPA 58 Title 01 (under DEQ). 

The TFF includes 11 belowground 300,000-

gal and 318,000-gal tanks (hereinafter referred to 

as 300,000-gal tanks) and four belowground 

30,000-gal tanks. The 300,000-gal tanks are 

numbered WM-180 through WM-190, and the 

30,000-gal tanks are designated as WM-103 

through WM-106. The second modification to the 

Consent Order specifies that DOE must cease use 

of Tanks WM-182, WM-183, WM-184, WM-185,c

and WM-186 by June 30, 2003, and the remaining 

tanks by December 31, 2012. Ceasing use of the 

tanks, as defined in the Consent Order, means that 

“DOE must empty the tanks down to their heels, 

i.e., the liquid level remaining in each tank will be 

lowered to the greatest extent possible by the use 

of existing transfer equipment” (IDHW 1998). 

According to the Idaho Hazardous Waste 

Management Act of 1983 (HWMA) and the 

Resource Conservation and Recovery Act 

(RCRA), the TFF is an interim status hazardous 

waste management unit (State of Idaho 1983; 42 

United States Code [USC] 6901, 1976). Because 

of this status, the requirements of 40 CFR 265 

(2002) apply to the TFF closure (rather than 

40 CFR 264 [2002]). 

The TFF will continue to operate until 2012 

while various parts of the facility are being closed. 

The final closure of any component of the TFF 

will not be complete until all the tanks have been 

closed and the remedial investigation/feasibility 

study (RI/FS) for operable unit (OU) 3-14 (Tank 

Farm Soils) is completed. The final closure plan 

will address closure and any required post-closure 

care of the TFF. A decision to close the unit as a 

landfill or as clean closure will be determined 

during final closure. Clean closure is anticipated, 

as discussed in the HWMA/RCRA Closure Plan 

(DOE-ID 2001a). 

Two significant releases from TFF ancillary 

equipment to surrounding soils occurred in 1955 

and 1972. No releases have occurred from the 

tanks to environmental media. These releases are 

subject to investigation and remediation as 

necessary under the INEEL Comprehensive 

 

c. The Consent Order allows Tank WM-185 to be used as an 

emergency spare tank. 
1-4

Environmental Response, Compensation, and 

Liability Act (CERCLA) (42 USC 9601–9675, 

1980) program as described in the Federal Facility 

Agreement and Consent Order (FFA/CO) (IDHW, 

Environmental Protection Agency [EPA], and 

DOE Idaho Operations Office [DOE-ID] 1991). 

Performance assessments (PAs) are used by 

DOE to ensure a reasonable expectation that lowlevel waste (LLW) disposal will meet the 

radiological performance objectives in DOE 

Manual 435.1-1 (DOE 2001a). This PA addresses 

closure of the TFF, including the 11 belowground 

300,000-gal stainless-steel tanks and 

corresponding vaults, four belowground 30,000-

gal stainless-steel tanks, and associated ancillary 

equipment (e.g., piping and jets). This PA 

evaluates only the radiological waste constituents 

of the TFF; the hazardous constituents are 

addressed in the HWMA/RCRA Closure Plan for 

Tanks WM-182 and WM-183 (DOE-ID 2001a). 

1.1 General Approach 

As required by Sections I.2.F, “Field Element 

Managers,” and II.U(3), “Closure,” of DOE 

Manual 435.1-1 (2001a), the responsible Field 

Element Manager is required to prepare two 

facility/site closure plans (Tier 1 and Tier 2) for 

each deactivated HLW facility or site or a 

combination of facilities (named a site) that are to 

be analyzed and closed together due to similarities 

or proximity. The closure plans will 

Define the approach that will be taken for 

ensuring the long-term protection of the 

public and the environment from the 

closure of deactivated HLW facilities 

containing residual LLW or [transuranic]

TRU wastes. This two-phase closure plan 

process is considered appropriate and 

necessary because much of the sitespecific and detailed facility and waste 

characterization data needed to conduct 

projected performance analysis of the 

closure action may not be available at the 

time the closure action is being defined 

and resources are being solicited (DOE 

2000). 

The Tier 1 Closure Plan defines and bounds 

the parameters of the TFF closure action and must 

be approved by the Deputy Assistant Secretaries 

for Waste Management and Environmental 

Restoration before physical closure activities 

begin. This review and approval is intended to 

reduce the risk of committing a significant amount 

of resources to the closure before the appropriate 

levels of management have approved the closure 

plan. The Tier 2 Closure Plan, which will be 

approved by the Field Element Manager, will 

provide the detailed information related to the TFF 

closure using the bounding conditions contained in 

the Tier 1 Closure Plan. The Tier 2 Plan is 

considered an extension of the Tier 1 Plan and will 

give facility/site-specific analyses of each facility 

defined and bounded by the Tier 1 Closure Plan. 

Thus, the Tier 2 Plan documentation should 

demonstrate that the performance objectives and 

other commitments identified in the Tier 1 Plan 

can be met and maintained. 

This PA has been developed to support the 

DOE-ID Tier 1 Closure Plan for the TFF 

(DOE-ID 2003a). The methodology employed for 

the preparation of the TFF PA has been based on a 

logical sequence of steps as described by 

• Guidelines for Radiological Performance 

Assessment of DOE Low-Level Radioactive 

Waste Disposal Sites (Case and Otis 1988) 

• Guidance from DOE Order 435.1 (2001b), 

Manual 435.1-1 (2001a), and Guidance 

435.1-1 (1999a) 

• Tier 1 Closure Plan Content Guide (DOE 

2000) 

• Format and Content Guide for U.S. 

Department of Energy Low-Level Waste 

Disposal Facility Performance Assessments 

and Composite Analyses (DOE 1999b). 

In addition, an approved LLW PA for the 

Radioactive Waste Management Complex 

(RWMC) at the INEEL was used to support the 

development of the TFF PA (Maheras et al. 1994 

and 1997). 
1-5

1.1.1 Steps in Developing the 

Performance Assessment 

The initial step in a PA is to set the scope of 

the analysis and identify the applicable 

performance objectives. The result of this process 

is documented in the remainder of Section 1. 

The next step is to document relevant site 

characterization data, relevant features of the 

environment and facility design features that are 

important to the development of conceptual 

models, and the definition of analysis parameters. 

The relevant site characterization data and facility 

design features for the TFF PA are provided in 

Section 2. 

Since the TFF PA supports the Tier 1 Closure 

Plan, the analysis must be bounding and is based 

on limited waste characterization data. An 

important parameter used in a PA is the estimated 

residual radionuclide inventory. Limited 

characterization data are available for determining 

the radionuclide inventory for the TFF. In 

addition, the projected final result of 

decontamination for the tanks and associated 

piping is currently unknown. Closure of the TFF 

tanks will be performed in phases; Tanks WM-182 

and WM-183 will be closed in the first phase. The 

closure of these two tanks will serve as a proof-ofprocess demonstration of the waste removal, 

decontamination, and sampling techniques for the 

closure of the remaining TFF tanks. However, 

until the closure of the first two tanks is complete, 

professional judgment and bounding calculations 

are required to determine a radionuclide inventory 

for the TFF. 

The radionuclide inventory for the TFF is 

based on computer modeling, professional 

judgment, process knowledge, and limited 

sampling data. This inventory was then assumed to 

be contained in each of the eleven 300,000-gal 

tanks. Similar calculations were performed for 

contaminated sand pads located beneath two tanks. 

The piping inventory was based on analytical data 

collected during closure of tank WM-182. The 

radionuclide inventories for the tanks, sand pads, 

and piping are given in Section 2.3. A full 

description of the development of the radionuclide 

inventory is given in Appendix A. 

The analysis of performance begins with 

developing and selecting scenarios for evaluation. 

Several deterministic site-specific scenarios have 

been developed. (In a deterministic model, 

relationships are fixed and no probabilities are 

entered, so a given input produces one predictable 

output.) A scenario is defined here as a description 

of all the features, events, and processes 

influencing the waste disposal system 

performance. Each scenario consists of several 

scenario modules that have a corresponding 

conceptual model and mathematical 

representation. 

Scenarios are developed by preparing a 

comprehensive list of features, events, and 

processes. This list is then screened to remove 

processes or events that will not influence 

performance, are deemed physically improbable, 

or are extremely rare. This approach may lead to 

the creation of more than one scenario. In this PA, 

the scenarios consist of the groundwater transport 

model, radon release model, and volatile 

radionuclide release model. Associated with each 

of these scenarios is a dose model. The scenario 

“features” lists were used to develop conceptual 

models by making assumptions about the 

magnitude, timing, consequences, and detailed 

mechanisms of the events and processes. The 

conceptual models were used to formulate the 

mathematical models ultimately used to simulate 

site performance. 

The approach used throughout the model 

development and parameter selection process was 

to make reasonable assumptions and select best 

estimate parameter values. Where parameter 

values were not well-defined or it was necessary to 

limit model-complexity, conservative assumptions 

have been made. The conceptual models, scenario 

models, dose models, and associated dose models 

are described in Section 3. The results of the 

models for each scenario are presented in 

Section 4. 

The intruder scenario also was evaluated for 

the TFF PA. This scenario was developed based 

upon the physical environment at the TFF. The 

development of the intruder scenario, 

mathematical representations, results, and 
1-6

sensitivity/uncertainty analysis are described in 

Section 5. 

The remainder of the TFF PA presents an 

interpretation and evaluation of (a) the 

performance analysis results compared to the 

performance objectives and of (b) the sensitivity 

and uncertainty of the analyses to ensure a 

reasonable expectation of facility performance. 

1.1.2 Conservative Parameters Used in 

Performance Assessment 

Every effort has been made in this PA to use 

very conservative inventory and modeling 

parameters and demonstrate that performance 

criteria are met, especially in the radionuclide 

inventory development, sand pad inventory, 

degradation analysis, and performance models. In 

determining the radionuclide inventory, recent 

analytical results from sampling Tanks WM-182, 

WM-183, and WM-188 were considered. 

Generally, the highest radionuclide concentrations 

were found in the WM-188 analytical data. These 

concentrations were used in the radionuclide 

inventory calculations rather than an average of 

the analytical data from the three tanks. In 

addition, the same radionuclide inventory is 

attributed to each of the 300,000-gal tanks. 

For the sand pad inventory, which is necessary 

because of accidental releases in 1962 that 

contaminated the sand pads in Tanks WM-185 and 

WM-187, several conservative assumptions were 

used. For example, the surface available for 

diffusive transport is assumed to instantaneously 

remove all radionuclides that reach the surface 

(the liquid/sand pad interface). This assumption 

results in greater mass transfer than if a 

concentration gradient is assumed to exist in the 

sand pad. Also, the sand pad is assumed to be 

saturated from the release and air does not occupy 

the pore space, providing an increased mass of 

diffusing radionuclides into the sand pad. 

Complete flushing of the sand pad is assumed to 

occur annually, completely draining the liquid and 

removing radionuclides from the sand pad. 

Finally, a conservatively high effective diffusion 

coefficient value for each radionuclide from the 

liquid to the sand pad was used. 

The degradation analysis also incorporates 

several conservative assumptions. For instance, 

the degradation rate parameters for the vault, grout 

between the vault and tank, piping, and the tank 

and grout inside the tank are 100, 100, 500, and 

500 yr, respectively. These degradation step 

changes are within the ranges predicted by the 

degradation analysis, even for the minimum 

degradation case, which indicates 100, 500, 1,750, 

and 1,750 yr, respectively. (A summary of the 

degradation analysis is provided in Table 7-9.) 

Other conservative assumptions used in the 

degradation analysis include determining the rates 

of modeled chemical attacks from experiments 

involving degrading instead of intact concrete and 

not taking credit for the chemical barrier provided 

by the chemically-reducing grout, vault, and tank. 

Only the physical barrier to flow and transport is 

modeled, while the chemical barrier is probably 

more significant in affecting radionuclide release 

rates. 

In addition, the vault failure, modeled as 

occurring from the expansion reaction in 

reinforcement corrosion, is assumed to be 100 yr, 

but no credit is taken for the non-aggressive 

corrosion environment surrounding the vault. 

Finally, the corrosion rates that were determined 

for coupons placed in the tank liquid and used in 

corrosion rate calculations are expected to be 

greater than corrosion rates in the grouted tank and 

piping and in the water contacting the vault wall. 

The performance models used in this PA were 

designed to reflect the physical setting of the 

INEEL site to the extent possible, which 

contributes to the integrity of the PA. The site 

comprises a very complex geography, including 

the influence of the Big Lost River. Comparison of 

the modeled results to actual conditions was 

favorable. 

The numerical model developed to simulate 

water and radionuclide transport was structured to 

use reliable and verified information where 

available and incorporated conservative ranges for 

less-certain data. Where appropriate, existing 

hydrologic data were used extensively to describe 

subsurface hydraulic conditions. For example, the 

distribution of basalt units and interbeds closely 
1-7

followed the geologic cross sections presented in 

several USGS reports (Anderson 1991; Anderson 

and Bartholomay 1995; Anderson, Ackerman, and 

Liszewski 1996; Anderson, Liszewski, and Cecil 

1997; Anderson, Kuntz, and Davis 1999). In cases 

where data from several references conflicted, 

attempts were made to confirm the data used in the 

model from different sources. For example, if the 

drilling logs indicate that a selected interbed 

consists of both fine and coarse grain sediments, 

separate hydraulic information such as the 

hydraulic conductivity or unsaturated hydraulic 

characteristic was evaluated to confirm the 

lithologic information. If no information was 

available to confirm lithologic or hydraulic data, 

then a conservative approach was adopted that 

resulted in using the value that yielded the highest 

transport rates. 

The calibration of the PORFLOW water 

transport model was conducted under existing site 

conditions. These conditions included water from 

the percolation ponds and the Big Lost River. The 

model was calibrated against existing water level 

data that defined the location and depth of perched 

water. Once the calibration was completed and a 

reasonable match obtained, the base flow model 

from radionuclide transport simulations contained 

only the Big Lost River input. The percolation 

ponds will be removed from service prior to 

the1000-yr simulation period. The existing model 

simulates the cascading of water from 

discontinuities in the perched zone. This approach 

provides a dynamic model of the flow in the 

unsaturated zone. This modeling scenario is 

considered to be conservative since the 

percolations would increase the perched zone’s 

ability to restrict downward contaminant transport. 

The modeling assumed that the system would 

begin failing at 100 yr; at 500 yr the entire vault 

would fail. At 500 yr, the entire vault structure 

was assumed to be a porous media with the 

permeability the same as the surrounding 

alluvium. This assumption is very conservative 

because the grout and concrete would most likely 

fracture. Transport would occur as unsaturated 

flow in the fractures with diffusion limited 

transport from the surrounding porous blocks to 

the fracture. Transport under these conditions is 

significantly less that the model scenario. 

At every stage of model development, a 

conservative approach was used to select hydraulic 

and transport properties. This approach resulted in 

a dynamic model capable of a wide range of 

simulations for testing a wide range of hydraulic 

and transport parameters in a realistic manner. 

Further, implementing conservative values for 

less-certain data yielded transport rates that are at 

the upper range for a realistic numerical 

simulation. 

The PA was conducted in accordance with the 

quality assurance requirements given in 

Appendix B. 

1.2 General Facility Description 

1.2.1 Tank Farm Facility 

The TFF is part of INTEC and includes 

11 belowground 300,000-gal and four 

belowground 30,000-gal stainless-steel tanks. 

Aboveground structures in the TFF include the 

TFF Control House (CPP-628) as the most 

significant building, several other ancillary 

buildings, condenser pits, valve boxes, and tank 

and vault sump riser covers. A perimeter fence 

encloses the TFF (see Figure 1-2). Gates are 

located on the west and north sides of the fence. 

A detailed description of the INTEC TFF is 

provided in Appendix C. The Computer Interface 

Building is not associated with the TFF. Portions 

of the piping associated with WM-182 and 

WM-183 will be decontaminated and capped or 

otherwise sealed during this closure. Condenser 

pits, valve boxes, and tank and vault sump riser 

covers will be closed during closure of WM-182 

and WM-183. 

From 1953 to 1988, DOE processed spent 

nuclear fuel at INTEC. The process was designed 

to recover the highly enriched uranium in the fuel 

using a three-step solvent extraction process. The 

first solvent extraction cycle resulted in a highly 

radioactive liquid that was considered HLW and 

was stored at the TFF. Subsequent extraction 

cycles and decontamination activities generated a 

liquid waste that was concentrated by evaporation 

and also stored at the TFF. 
1-8

Figure 1-2. A map of the INTEC Tank Farm Facility. The numbers shown are building or structure 

numbers.
1-9

Construction of the TFF began in 1951 with 

Tanks WM-180 and WM-181. Tanks WM-182 

through WM-184 were completed in 1955, Tanks 

WM-185 and WM-186 were completed in 1957, 

and Tanks WM-187 and WM-188 were completed 

in 1959. The last tanks, WM-189 and WM-190, 

were constructed in 1964. Construction of the four 

30,000-gal tanks was completed in 1955. The 

HLW was removed from the 30,000-gal tanks and 

the tanks were flushed in 1974 and 1975. The 

30,000-gal tanks were reused in 1982 and 1983 

but the waste was Process Equipment Waste 

Evaporator condensate (DOE-ID 2000). 

Eight of the tanks (WM-180, WM-182, 

WM-183, WM-185, and WM-187 through 

WM-190) contain stainless-steel cooling coils to 

cool the waste contents in the tanks. Risers 

provide access to each tank. Each tank has four or 

five 12-in.-diameter risers. Tanks WM-184 

through WM-190 also has two 18-in.-diameter 

risers (see Figure 1-3). Most risers also house 

equipment, such as radio frequency probes for 

level measurement, corrosion coupons, or waste 

transfer equipment (steam jets and airlifts). Two 

steam jets are located inside each tank, except for 

Tanks WM-189 and WM-190. These two tanks 

each have one steam jet and one airlift pump. A 

single steam jet can transfer waste out of a tank at 

approximately 50 gpm, and an airlift can transfer 

waste out of a tank also at approximately 50 gpm 

(INEEL 2000a). 

Each vault contains at least one sump and 

steam jet pump to remove any liquid waste or 

surface water that may leak into the vault. The 

sumps have liquid-level detectors and high-level 

alarms. 

Each 300,000-gal tank is contained in a 

concrete vault. The vaults are located 

approximately 45 ft (14 m) belowground and are 

configured in one of three basic designs: 

monolithic octagonal, pillar and panel octagonal, 

or monolithic square (see Figures 1-4 through 

1-6). The 6-in. thick concrete vault roofs are 

covered with approximately 10 ft (3 m) of soil to 

provide radiation shielding. Tanks WM-182 and 

WM-183 are contained in pillar and panel 

octagonal vaults. Because vaults of this design 

were constructed with prefabricated components, 

these vaults are not considered as robust as vaults 

of monolithic design (Palmer et al. 1999). 

Liquid wastes transferred to, from, and among 

the tanks are managed through a system of piping, 

valves, and monitoring boxes. The liquid waste is 

routed through waste transfer valves located in 

underground, stainless steel-lined concrete boxes, 

referred to as valve boxes. Liquid releases 

resulting from decontamination efforts or leakage 

of liquids from the valve boxes and piping or 

valve and liquid in piping encasements (secondary 

containment for piping) is drained to tanks 

(INEEL 1998). 

1.2.2 Closure Concept 

In general, the closure strategy for the TFF 

involves the removal of waste from the tanks. The 

piping (except the abandoned piping that was 

previously flushed and decontaminated) and tank 

vaults will be decontaminated. Confirmatory 

sampling and analysis will then be performed, 

followed by grouting of the tanks and vaults. 

Grout also will be pushed through decontaminated 

waste lines. Grouting activities will use 

chemically-reducing grout to further minimize 

any potential leaching of radionuclides. An 

overview of the TFF closure concept is shown in 

Figure 1-7. 

Each TFF tank will be closed using the 

following general sequence: 

1. Remove liquid waste in tanks 

2. Empty tank to heel with existing jets 

3. Flush tank and piping 

4. Empty to heel with existing jets 

5. Wash tank and remove waste 

6. Video and sample heel residuals 

7. Evaluate tank residuals against closure 

performance objectives 

8. Displace heel with grout 
1-10

Figure 1-3. Vault and dome of Tank WM-185 showing the concrete beams and concrete risers on top. 

Figure 1-4. Monolithic octagonal vault for Tank WM-180. 
1-11

Figure 1-5. Pillar and panel octagonal vault for Tank WM-183. 

Figure 1-6. Monolithic square vault for Tank WM-190. 
1-12

Figure 1-7. Simplified tank closure sequence. 

9. Fill piping with grout 

10. Fill tank vault with grout 

11. Fill tank with grout. 

Steps 1 through 3 are considered part of the 

tank deactivation. Steps 4 through 7 are required 

to remove waste and decontaminate a given tank. 

These steps may be repeated, as necessary, until 

the decontamination sequence is no longer 

effective (i.e., additional decontamination is no 

longer economically practical). The last four steps 

will remove remaining tank residuals to the extent 

readily feasible, stabilize the residuals, and leave 

the wastes in a solid physical form. 

Contaminated soils associated with the TFF 

are not addressed in the TFF closure or this PA. 

Instead, past releases of contaminants from the 

TFF are being addressed under CERCLA as 

described in the FFA/CO signed in 1991 (IDHW, 

EPA, and DOE-ID 1991). However, the potential 

interaction of contaminants in the surrounding 

TFF soils with potential future releases from the 

TFF tanks and associated systems are addressed in 

the TFF Composite Analysis Report (DOE-ID 

2003b). 

Closure of the TFF tanks will be performed in 

phases; Tanks WM-182 and WM-183 will be 

closed in the first phase. The closure of these two 

tanks will serve as a proof-of-process 

demonstration of the waste removal, 

decontamination, and sampling techniques for the 

closure of the remaining TFF tanks. This 

demonstration is necessary because of the 

uniqueness and complexity of the TFF closure. 

The closure is unique because similar tanks have 

not been closed previously at the INEEL. It is 

complex because of the nature of the waste and the 

configuration of the tanks. High radiation fields in 

the tanks and associated equipment preclude 

manual decontamination of most areas so remotehandling techniques must be used. Furthermore, 

access to the tanks and vaults is available only 

through risers, which prevents the use of routine 

decontamination and sampling procedures. 

Therefore, the closure strategy may be refined for 

subsequent closure phases based on information 

obtained during the closure of Tanks WM-182 and 

WM-183. 

The closure of Tanks WM-182 and WM-183 

will be executed in three phases: the planning 

phase, the pre-closure phase, and the closure 

demonstration phase. The planning phase 

encompasses activities such as document 
1-13

preparation, reviews, and notifications. Pre-closure 

phase activities include site preparation and 

decontamination studies. The closure 

demonstration phase includes evaluation of results 

from pre-closure phase decontamination studies, 

and tank decontamination and grouting. The preclosure and closure activities are described in 

detail in the conceptual closure approach report for 

closure of Tanks WM-182 and WM-183 (INEEL 

1999a). 

The final closure of the TFF will not occur 

until all tanks have been closed and the RI/FS for 

OU 3-14 completed. However, the requirements 

for and the general features of the cover design for 

the TFF can be described. The TFF cover will, at a 

minimum, meet the standards described in 

10 CFR 61, “Licensing Requirements for Land 

Disposal of Radioactive Waste,” (2003). 

The general features of a cover for the closed 

TFF are similar to a conceptual cover design for 

the Idaho CERCLA Disposal Facility and are 

based on the functions and sequence of engineered 

layers that are commonly called the “Hanford 

Barrier” (Ward and Gee 2000). The Hanford 

Barrier was developed to provide long-term 

protection (1,000 yr) of certain types of waste in 

semiarid environments and exceeds RCRA cover 

design requirements, including life expectancy and 

hydraulic conductivity. The main function of a 

surface barrier is to ensure that buried wastes are 

contained and protected from environmental and 

biotic forces. Surface barriers have been identified 

as a critical component in management of buried 

wastes and other sources of subsurface 

contamination. Barrier technology, particularly for 

long-term deployment, remains largely unproven 

at the field scale (Myers and Duranceau 1994). 

The overriding technical purpose of the cover 

is to minimize movement of precipitation into the 

volume of contaminated materials. Other 

important roles are to discourage human intruders 

and to minimize the long-term maintenance that 

might be needed to ensure that the overriding 

purpose is not compromised by subsidence, 

intrusion from plants and animals, or by other 

natural forces. 

The evapotranspiration part of the cover 

removes as much of the precipitation falling on the 

cover as possible through evaporation and 

transpiration. To the extent that this layer is 

successful in carrying out this purpose, the 

likelihood of water reaching the residual 

contamination and mobilizing contaminants will 

be greatly reduced. 

Because the evapotranspiration layer cannot 

ensure that no water will pass through the upper 

part of the cover, additional protection to prevent 

water from reaching the residual contamination is 

needed. A layer of low-permeability clay is 

frequently used to provide that protection; 

however, the clay must remain moist to prevent 

cracking and a consequent significant increase in 

permeability. The capillary break portion of the 

cover keeps the clay from drying out due to a loss 

of moisture and provides, at the bottom of the 

capillary break section, a drain for any excess 

moisture that would come through from above. 

Part of the capillary break also serves as a barrier 

to intrusion and to erosion if natural forces remove 

the upper layers of the cap. The nominal design for 

the Hanford Barrier uses low-permeability asphalt, 

not clay, as the water barrier, but the other 

purposes of the layers in the capillary barrier 

dictate that the clay not be eliminated. 

Any water that is not removed by 

evapotranspiration or drained off by through the 

sand layer at the bottom of the capillary break 

portion of the cover is prevented from entering the 

zone of residual contamination by the infiltration 

barrier. Figure 1-8 shows this part of the cover as -

consisting of a low-permeability asphalt layer and 

a base layer. Typical thicknesses for these layers 

are 4 to 6 in. each. 

In most disposal facilities, a major purpose of 

the fill layer is to help provide a physical buffer 

between the waste and the infiltration barrier, 

thereby protecting the barrier from any possible 

damage from objects in the waste and from minor 

waste settling. This function is not needed, 

however, because the tank vaults being covered 

are already overlain with considerable soil, the 

vaults do not present any protruding physical 

features at their tops, and the vaults are unlikely to 
1-14

Figure 1-8. Typical cover layering. 

experience any settling. A second purpose for the 

fill is to define the vertical contours of the cover. 

The ability of any cover over residual waste to 

fulfill its design goals effectively is influenced by 

many factors. Subsidence of some kinds of waste 

can cause local low spots in the cover and trap 

infiltrating water instead of letting it run off 

harmlessly. This is very unlikely to happen with 

the fully grouted high-level waste tanks because 

there will be no voids under the cover that can 

cause subsidence. 

The effects of wind and water erosion and 

plant and animal intrusion are difficult to predict. 

Any cover placed over the tanks, however, will be 

designed to minimize these effects over the long 

term based on site-specific parameters. 

In projecting the performance of waste 

management systems, the presence of completely 

man-made materials such as geotextile filter fabric 

or a flexible membrane liner is sometimes 

conservatively ignored. At other times those 

components are assumed to perform for hundreds 

of years. No information is available about the 

durability of the man-made asphalt layer in the 

Hanford Barrier but as long as it is covered to 

prevent exposure to ultraviolet rays it should be 

expected to perform as designed for at least 

hundreds of years. 

1.2.3 Waste Characterization and 

Certification Program 

During and at the conclusion of the preclosure activities for Tanks WM-182 and 

WM-183, samples of tank residuals will be 

collected to determine the concentrations of 

hazardous constituents (to support HWMA/RCRA 

closure) and radionuclides (to support DOE Tier 1 

and Tier 2 closure) remaining in the tanks. A 

phased approach will be used for sampling the 

tanks. Existing equipment will be used to measure 

concentrations of gamma-emitting radionuclides 

that are being removed from the tank. This realtime monitoring of the waste stream will be 

trended as it is moved from Tanks WM-182 and 
1-15

WM-183. As the concentrations are reduced and 

begin to stabilize, personnel will collect the data 

set to determine the decontamination factor 

attained. The sampling and analysis approach is 

described in detail in the Sampling and Analysis 

Plan for the Post-Decontamination 

Characterization of the WM-182 and WM-183 

Tank Residuals (Portage Environmental 2002). 

Tank sample results will be used to evaluate the 

final status of the tank contamination and to 

compare to the PA inventory to ensure that a 

bounding analysis was performed for the Tier 1 

Closure Plan (DOE-ID 2003a). 

1.2.4 General Land Use Patterns 

The INEEL Comprehensive Facility and Land 

Use Plan (DOE-ID 1997a) and the INTEC Final 

Record of Decision (ROD) (DOE-ID 1999) 

describe the land use for the INEEL and INTEC. 

Land use at the INEEL is currently governmentcontrolled industrial use. The term controlled 

means that unrestricted public access to INTEC 

and the INEEL is not available. Presently, access 

to INEEL facilities requires proper clearance, 

training, or escort and controls to limit the 

potential for unacceptable exposures. A security 

force is used to limit access to approved personnel 

and visitors. These controls are estimated to be in 

place for a minimum of 100 yr. 

Future land use scenarios are identified in the 

Long-Term Land Use Future Scenarios for the 

Idaho National Engineering Laboratory (DOE-ID 

1995a). This document was developed using a 

stakeholder process that involved a public 

participation forum, a public comment period, and 

the INEEL Citizens Advisory Board. Following 

review and comment by the public participation 

forum, the document underwent a 30-day public 

comment period, and was subsequently submitted 

to the Citizens Advisory Board for review and 

recommendations. 

To date, no recommendations have been 

received for residential use of any portion of the 

INEEL until at least year 2095. Until 2095, the 

reasonably anticipated future use for the TFF is as 

a government-controlled industrial facility (DOEID 1999). 

Planning assumptions in the INEEL 

Comprehensive Facility and Land Use Plan 

(DOE-ID 1997a) are that the INEEL will remain 

under government control for a minimum of 

100 yr. This PA assumes that the institutional 

control period begins at the end of closure for the 

TFF. The anticipated date of completion for the 

TFF closure project is 2016. Therefore, the end of 

institutional control is assumed to be 2116 for the 

purposes of dose calculations in this PA. 

1.3 Related Documents 

As an interim status hazardous waste 

management unit, the TFF closure must comply 

with applicable HWMA/RCRA requirements 

(State of Idaho 1983; 42 USC 6901, 1976). 

However, the TFF also is a HLW facility regulated 

by DOE, and must meet the requirements of DOE 

Order 435.1 (DOE 2001b) and its associated 

manual (DOE 2001a) and guidance (DOE 1999a). 

Three closure paths are allowed by DOE 435.1-1: 

decommissioning, the CERCLA process, and 

closure by an approved closure plan. The TFF will 

be closed under an approved DOE closure plan. 

DOE requires a two-tiered approach to closure 

plan development, review, and approval. The 

Tier 1 Closure Plan (DOE-ID 2003a) defines and 

bounds the parameters of the TFF closure actions. 

It must be approved by the Deputy Assistant 

Secretaries of Waste Management and 

Environmental Restoration before physical closure 

activities may begin. This PA supports the Tier 1 

Closure Plan. The Tier 2 Closure Plan will give 

specific analyses of the TFF closure. It should 

show that the performance objectives and other 

commitments made in the Tier 1 Plan can be met 

and maintained. The Tier 2 Closure Plan will be 

approved by DOE-ID. 

The Composite Analysis for Tank Farm 

Closure (DOE-ID 2003b) is a planning tool used 

to ensure that the combined effect of all sources of 

residual radioactive material that could contribute 

to the radioactive dose calculated will not 

compromise the requirements to protect the public. 

In addition, IDAPA 58.01.05.009 [40 CFR 

265.197(c)(l), 2002] specifies that if a tank system 

does not have adequate secondary containment, 
1-16

then the owner or operator of a tank system must 

prepare both a closure plan for clean closure and a 

contingent closure plan for closure as a landfill. 

The decision to close the TFF to clean closure 

standards under HWMA/RCRA or as a landfill 

will be made when all the tanks have been closed 

and the RI/FS for OU 3-14 is complete. Because 

this decision has not been made, both a closure 

plan and a contingent plan have been written. The 

Idaho Hazardous Waste Management 

Act/Resource Conservation and Recovery Act 

Closure Plan for Idaho Nuclear Technology and 

Engineering Center Tanks WM-182 and WM-183

describes the strategy for clean closure of the tanks 

(DOE-ID 2001a). The Contingent Landfill Closure 

and Post-Closure Plan for Idaho Nuclear 

Technology and Engineering Center Tanks 

WM-182 and WM-183 discusses the landfill 

closure strategy (DOE-ID 2001b). 

Because the TFF was used to store HLW, a 

Waste-Incidental-to-Reprocessing (WIR) 

determination reportd

 is being prepared in 

accordance with DOE Order 435.1. In the WIR 

report, the TFF residuals are evaluated to 

determine whether the residuals can meet the 

incidental waste criteria of LLW, rather than HLW 

(DOE 2001a). It is anticipated that the WIR 

determination will allow the residuals to be 

managed as LLW. 

Several other important related documents 

define the scope of the TFF PA and provide 

resources and background information relevant to 

the analyses. 

Under the terms of a 1992 Consent Order (and 

subsequent modifications) between IDHW and 

DOE, DOE must permanently cease use of tanks 

in the TFF or bring the tanks into compliance with 

secondary containment requirements as set forth 

by IDAPA 58.01.05.009 (40 CFR 265.193, 2002). 

The Consent Order further specifies that this 

compliance cannot be achieved through an 

equivalency demonstration or by obtaining a 

 

d. DOE-ID, 2002, Idaho Nuclear Technology and 

Engineering Center Tank Farm Facility Residual-WasteIncidental-to-Reprocessing Report, DOE/ID-10777, Draft B, 

December. 

variance as provided by IDAPA 58.01.05.009 [40 

CFR 265.193(d)(4) and (h)]. DOE plans to close 

the TFF tanks because high radiation fields would 

make compliance with secondary containment 

requirements difficult, and a need for such storage 

is not evident after 2012. 

The TFF is located within the boundaries of 

the INTEC facility at the INEEL. Past releases of 

hazardous and radiological contaminants from 

INTEC, including the TFF, are being addressed 

under CERCLA (42 USC 9601–9675, 1980), as 

described in the FFA/CO signed by DOE, EPA, 

and IDHW in 1991 (IDHW, EPA, and DOE-ID 

1991). Also, an RI/FS (Rodriguez et al. 1997) and 

a ROD (DOE-ID 1999) have been completed for 

Waste Area Group 3 (WAG 3), which 

encompasses INTEC. Past releases from leaking 

pipes and associated soil contamination are being 

addressed by this CERCLA action and therefore 

are not considered in this PA. The Composite 

Analysis for Tank Farm Closure (DOE 2002b) 

considers the contribution to the groundwater all 

pathways dose from all radionuclide sources at the 

INEEL. 

The Idaho High-Level Waste & Facilities 

Disposition Final Environmental Impact Statement 

(EIS) analyzes the potential environmental 

consequences of managing two waste types at the 

INEEL: high-level waste in calcine form and 

liquid mixed waste (historically known as sodiumbearing waste and newly generated waste). This 

EIS evaluates several alternatives to the 

disposition of the TFF. 

The INTEC TFF Tanks WM-182 and WM-183 

Closure Study provides the closure methodology 

for the TFF.e

 This document provides information 

on the proposed process for closure of the tanks, 

vaults, and associated ancillary equipment 

including piping. 

Two additional reports that support this PA are 

the Idaho Chemical Processing Plant Safety 

Document (INEEL 1998) and the Idaho Nuclear 

 

e. LMITCO, 1998, INTEC TFF Tanks WM-182 and WM-183 

Closure Study (Draft), August 31. 
1-17

Technology and Engineering Center Safety 

Analysis Report (INEEL 2001a). 

1.4 Performance Criteria 

In July 1999, DOE issued Order 435.1, 

“Radioactive Waste Management.” This Order and 

its associated manual and guidance set forth the 

authorities, responsibilities, and requirements for 

the management of DOE’s inventory of HLW, 

transuranic waste, and low-level waste. DOE 

Manual 435.1-1 requires that 

A site-specific radiological performance 

assessment shall be prepared and 

maintained for DOE low-level waste 

disposed of after September 26, 1988. The 

performance assessment shall include 

calculations for a 1,000 year period after 

closure of potential doses to 

representative future members of the 

public and potential releases from the 

facility to provide a reasonable 

expectation that the performance 

objectives identified in this Chapter are 

not exceeded as a result of operation and 

closure of the facility. (DOE 2001a)

This requirement is applicable to the TFF closure 

plan. 

Performance assessments involve estimating 

radiological exposure to future members of the 

public. These exposure estimates are then 

compared to specific performance measures to 

determine whether there is a reasonable 

expectation that the performance measures will be 

met in the future. Performance measures consist of 

specific performance objectives identified in 

DOE Manual 435.1-1 and other performancerelated factors required by DOE Manual 435.1-1 

for deactivated HLW facility/site closure actions 

that contain residual LLW. 

This section describes each performance 

criterion used to assess the performance of the 

TFF closure facility/site and the points of 

assessment. These criteria include the performance 

objectives in DOE Manual 435.1-1, Section 

IV.P.(1), other performance criteria derived from 

the requirements in DOE Manual 435.1-1, Section 

IV.P.(2), and the points of assessment. 

Performance criteria not explicitly called out in the 

manual (e.g., site-specific regulatory agency 

agreements) also are discussed. 

1.4.1 Public Protection Performance 

Objectives 

The first applicable performance objective, 

from DOE Manual 435.1-1, IV P.(1)(a), states 

Dose to representative members of the 

public shall not exceed 25 mrem 

(0.25 mSv) in a year total effective dose 

equivalent from all exposure pathways, 

excluding the dose from radon and 

progeny in air. 

This performance objective is interpreted as 

requiring the performance analysis to provide a 

reasonable expectation that the “all-pathways” 

dose to a hypothetical future member of the public 

will not exceed 25 mrem (effective dose 

equivalent [EDE]) in a year, excluding doses from 

inhalation of radon and its short-lived progeny. 

“All pathways” include any and all modes by 

which a receptor at the point of public access 

could be exposed, including the air pathway. The 

analysis is to cover 1,000 yr following closure of 

the disposal facility. Analysis beyond 1,000 yr to 

calculate the maximum dose and the time of that 

dose shall be included in the sensitivity and 

uncertainty analyses as a means of increasing 

confidence in the outcome of the modeling. 

Normally, the point of compliance for this 

performance objective should be at the point of 

highest calculated dose beyond a 100-m buffer 

zone surrounding the waste. A larger or smaller 

buffer zone may be used with justification (DOE 

2000). 

The second performance objective, from DOE 

Manual 435.1-1, IV.P. (1)(b), states 

Dose to representative members of the 

public via the air pathway shall not 

exceed 10 mrem (0.10 mSv) in a year total 

effective dose equivalent, excluding the 

dose from radon and its progeny. 
1-18

Consistent with the National Emission 

Standards for Hazardous Air Pollutants (40 CFR 

61, Subpart H, 2002), 220Rn, 222Rn, and their 

progeny need not be included in the air pathway 

analysis for comparison with the 10 mrem/yr EDE 

performance objective; separate controls for the 

emission of radon are discussed below. For the air 

pathway dose analysis, the point of compliance 

should be the point of highest calculated dose 

beyond a 100-m buffer zone surrounding the 

waste. With justification, a larger or smaller buffer 

zone may be used. The 10 mrem/yr limit should be 

recognized to refer to all sources, not just the 

disposal facility. Therefore, if the performance 

assessment assumes a point of compliance that 

corresponds to the future land use boundary, a 

limit that is a fraction of the 10 mrem/yr dose limit 

should be used in recognition of the potential 

presence of other sources (DOE 2000). 

The third performance objective, from DOE 

Manual 435.1-1, IV.P.(1)(c), states 

Release of radon shall be less than an 

average flux of 20 pCi/m2

/s (0.74 Bq/m2/s) 

at the surface of the disposal facility. 

Alternately, a limit of 0.5 pCi/l (0.0185 

Bq/l) of air may be applied. 

For radon, a separate limit is applied. In most 

cases, the limit to be applied should be an average 

ground surface emanation rate of 20 pCi/m2

/s 

directly over the disposal unit. There may be 

special cases involving the disposal of material 

that radiologically resembles uranium or thorium 

mill tailings in isolated locations that warrant 

using an alternative limit. The alternative limit is 

an incremental increase of 0.5 pCi/L of radon in 

the air concentration at the point of assessment 

(DOE 2000). 

Table 1-1 gives a summary of the DOE 

performance objectives and the corresponding 

points of compliance. 

1.4.2 Water Resource Impact 

Assessment 

DOE Order 5820.2A (1998), superseded by 

DOE Order 435.1, contained a performance 

objective for protection of groundwater resources. 

DOE Manual 435.1-1 does not contain a specific 

performance objective (e.g., dose or concentration 

standard) for water resource impacts. The 

approach in DOE Manual 435.1-1 was chosen by 

DOE for consistency with Nuclear Regulatory 

Commission (NRC) methods for LLW disposal 

and radiation protection principles articulated by 

the National Council on Radiation Protection and 

Measurements and the International Commission 

on Radiological Protection (ICRP). In accordance 

with these principles, it is appropriate to assign a 

Table 1-1. Preliminary performance assessment components, objectives, and points of compliance.a

Component Performance Objective Point of Compliance 

All pathways 25 mrem in a year, not including doses 

from radon and progeny 

The point of highest projected dose or 

concentration beyond a 100-m buffer 

zone surrounding the disposed waste 

Air pathway 10 mrem in a year, not including doses 

from radon and progeny 

The point of highest projected dose or 

concentration beyond a 100-m buffer 

zone surrounding the disposed waste 

For 220Rn, 222Rn, and their progeny either 

(1) an average flux of ≤ 20 pCi/m2

/s or 

Radon Disposal facility surface 

(2) an air concentration of ≤ 0.5 pCi/L 

unless constrained by applicable laws, 

regulations, or agreements 

The point of highest projected dose or 

concentration beyond a 100-m buffer 

zone surrounding the disposed waste 

a. Source: DOE 2000. 
1-19

fraction (e.g., 25 mrem) of the 100 mrem/yr 

public dose performance measure to a particular 

practice (e.g., radioactive waste disposal) but it is 

not recommended to further fraction performance 

objectives to specific pathways (e.g., 

groundwater). Thus, exposure by water pathways 

is included in the all-pathways analysis but there 

is no specific performance objective for exposure 

by water pathways. In the case of the air pathway, 

the 10 mrem/yr performance objective is based on 

a specific federal regulatory requirement. There is 

no comparable requirement for water resources 

(DOE 2000). 

DOE Manual 435.1-1, IV.P.(2)(g), states 

For the purposes of establishing limits on 

radionuclides that may be disposed of 

near-surface, the performance assessment 

shall include an assessment of impacts to 

water resources. 

For water resources protection, impacts were 

assessed on a site-specific basis in accordance 

with a hierarchical set of criteria. This approach 

recognizes that there are no federal requirements 

for protection of water resources for a radioactive 

waste disposal facility or closure action. The sitespecific hierarchical approach, rather than 

mandating specific performance measures for all 

sites, is consistent with the EPA strategy for 

groundwater protection. EPA recognizes that 

groundwater protection is a regional and local 

matter (EPA 2001). Accordingly, the hierarchy for 

establishing water resources protection is as 

follows (from the DOE Deactivated High-Level 

Waste Facility Closure Manual, DOE 2000): 

• First, the closure action must comply with any 

applicable state or local law, regulation, or 

other legally applicable requirement for water 

resource protection. 

• Second, the closure action must comply with 

any formal agreement applicable to water 

resource protection that is made with 

appropriate state or local officials. 

• Third, if neither of the above conditions 

applies, the site should select assumptions for 

use in the performance assessment based on 

criteria established in the site groundwater 

protection management program and any 

formal land-use plans. 

• If none of the above conditions apply, the site 

may select assumptions for use in the 

performance assessment for the protection of 

water resources that are consistent with the 

use of water as a drinking water source. 

For assessments addressing the use of 

groundwater as a drinking water source, the point 

of assessment normally should be the location of 

the highest groundwater concentration outside a 

100-m buffer zone. A larger or smaller buffer 

zone may be used with justification (DOE 2000). 

In terms of protecting the groundwater as a 

resource, assuming some volume averaging based 

on projected use may be appropriate. Applying the 

performance measure at an assumed wellhead 

mixed with a reasonable volume of groundwater, 

based on site-specific assumptions regarding 

groundwater use, is appropriate. This assumes 

mixing is consistent with state or local laws, 

regulations, or agreements (DOE 2000). 

1.4.2.1 Sole Source Aquifer 

Designation. The Eastern Snake River Plain 

Aquifer has been designated by EPA as a sole 

source aquifer (58 FR 138, 1991). After sole 

source designation, any federal financial 

assistance projects are subject to EPA approval to 

ensure that these projects do not contaminate the 

aquifer and create a significant hazard to public 

health. However, the INEEL is operated by direct 

federal funding and is not funded through federal 

financial assistance projects (Maheras et al. 1997). 

Therefore, the designation of the Eastern Snake 

River Plain Aquifer as a sole source aquifer has 

no regulatory impact on the performance 

objectives used in performance assessments at the 

INEEL.

1.4.2.2 Groundwater Protection. DOE 

Order 435.1 and the supporting manual and 

guidance do not specify radiation dose or 

concentration limits that would constitute 

performance objectives for protection of 

groundwater (DOE 2001a, 2001b, 1999a). 

DOE Manual 435.1 uses a hierarchy approach that 
1-20

requires an analysis of potential applicable state 

and local requirements or agreements for 

groundwater protection.

Maheras et al. (1997) assessed the federal, 

state, and local requirements for groundwater 

protection at the INEEL for the RWMC PA. Five 

different options for the groundwater protection 

performance objective were evaluated. 

Option 1. Groundwater protection would be 

defined by the performance objective specified in 

IDHW Rules, IDAPA 58.01.11, “Groundwater 

Quality Rule” (2002). IDAPA 58.01.11.200, 

“Ground Water Quality Standards” (1997), states 

that the concentration of radioactive materials or 

radioactivity in potable water supplies shall not 

exceed the values listed in 10 CFR Part 20, 

Appendix B, Table 2, Column 2, “Effluent 

Concentrations” (2003). These concentrations are 

equivalent to a radiation dose of 50 mrem/yr EDE 

based on a water consumption rate of 2 L/d 

(0.5 gal/d). 

Option 2. Groundwater protection would be 

defined as compliance with the 25 mrem/yr total 

effective dose equivalent (TEDE) all-pathways 

standard contained in DOE Guide 435.1. Instead 

of being based solely on the consumption of 

drinking water, the evaluation of compliance 

would be based on all applicable exposure 

pathways (i.e., an “all-pathways” analysis). 

Option 3. Groundwater protection would be 

defined as compliance with the maximum 

contaminant levels (MCLs) for radionuclides 

specified in 40 CFR 141 (2002), promulgated in 

65 FR 236 (2000). EPA groundwater standards for 

radionuclides are compared, from 1976 to the 

present, in Table 1-2. The current radionuclide 

standards include (a) a limit on the concentration 

of 5 pCi/L for 226Ra and 228Ra combined, (b) a 

limit on the concentration of 15 pCi/L for gross 

alpha particle activity, not including radon or 

uranium, (c) a limit on the dose equivalent to the 

total body or any given internal organ of 

4 mrem/yr from man-made beta particle and 

photon radioactivity, and (d) a limit of 30 μg/L for 

uranium. The definition of man-made 

radionuclides excludes the beta-emitting progeny 

of 232Th, 235U, and 238U. In addition, the definition 

of gross alpha activity in 40 CFR 141 (2002) 

specifies that the gross alpha activity is to be 

inferred from measurements made from a dry 

sample. Therefore, any gaseous 222Rn originally 

present in the sample would not be present when 

the sample was counted. Consequently, the shortlived alpha-emitting 222Rn progeny and 218Po and 

214Po also would not be present and would be 

excluded from the gross alpha activity. 

40 CFR 141 also specifies that the concentrations 

that yield 4 mrem/yr for 3

H and 90Sr should be 

calculated based on a water intake of 2 L/d 

(0.5 gal/d) and the 168-hr data listed in Maximum 

Table 1-2. Comparison of MCLs for radionuclides. 

Radionuclide 

1976 Interim 

Standards 

1991 Proposed 

Standards 2000 Standards 

Combined 226Ra, 228Ra 

(pCi/L) 

5 20 5 

Total Gross α 

(pCi/L) 

15 with 226Ra 

(excluding U and Rn) 

15 without 226Ra 15 without 226Ra 

βγ emitters (mrem) 4 4 (EDE) 4

Uranium No standard 20 μg/L

(kidney toxicity)a

170 pCi/L

(radiotoxicity) 

30 μg/L 

a. Drinking water equivalent level of 100 μg/L and relative source contribution factor of 20%. 
1-21

Permissible Body Burdens and Maximum 

Permissible Concentration of Radionuclides in Air 

or Water for Occupational Exposure (Department 

of Commerce 1963). For 3

H and 90Sr, the 

concentrations that yield 4 mrem/yr total body 

dose are specified to be 20,000 pCi/L and 8 pCi/L, 

respectively. 

Option 4. The current MCLs specified in 

40 CFR 141 (2002) of 5 pCi/L for 226Ra and 228Ra 

(combined) and 15 pCi/L for gross alpha activity, 

including 226Ra but excluding radon and uranium, 

would be retained. The 4 mrem/yr dose limit for 

beta particle and photon radioactivity from manmade radionuclides also would be retained but the 

4 mrem/yr would now be an EDE, not a total body 

dose. The EDE would be calculated using a water 

intake of 2 L/d and ICRP-30 dosimetry, as 

implemented in EPA (1988), not using the ICRP-2 

dosimetry that served as the basis for the data 

listed in the Department of Commerce report 

(1963). The concentrations that yield 4 mrem/yr 

total body dose specified in 40 CFR 141 of 

20,000 pCi/L for 3

H and 8 pCi/L for 90Sr would be 

retained. 

Option 5. A dose limit of 4 mrem/yr (EDE) 

for all radionuclides would be used as the 

performance objective for the protection of 

groundwater resources, without the use of the 

MCLs. A water intake of 2 L/d and ICRP-30 

dosimetry, as implemented in EPA (1988), would 

be used to calculate the EDE. 

The reasoning for this option is based on an 

analysis of the 1976 EPA interim water quality 

standards (41 FR 298) and the proposed standards 

that have been published. In regard to standards 

for beta/gamma-emitting radionuclides, EPA 

clearly intended to replace the 1976 interim limits 

based on the dose limit for whole body or any 

organ by a limit based on EDE. This intention was 

shown in the proposed standards for 1991. Thus, 

for these radionuclides, the assumed groundwater 

protection requirement for the TFF disposal 

facility is the same as the proposed 1991 revision 

of the EPA drinking water. The only exception is 

that dose (not risk) is evaluated in the PA. 

Replacement of dose limits for whole body or any 

organ by a limit based on EDE is in agreement 

with the current radiation protection policies of 

DOE, NRC, and EPA. 

EPA has used concentration limits, rather than 

dose limits, for alpha-emitting radionuclides so the 

standard is expressed in terms of quantities that 

could be measured directly by operators of public 

drinking water systems. This approach has not 

been used for man-made, beta/gamma-emitting 

radionuclides because EPA has presumed that only 

naturally-occurring, alpha-emitting radionuclides 

would be of concern for public drinking water 

supplies. EPA states in 65 FR 236 (2000) that “a 

newly proposed MCL expressed mrem EDE could 

result in a more consistent risk level within the 

Agency’s target risk range.” However, in the final 

rule, they ratified the current standard since it was 

protective of public health. Commenters that 

supported EPA’s final rule stated that there was no 

appreciable occurrence of man-made beta emitters 

in drinking water, so it was not a pressing public 

health concern to revise the MCL. 

The final rule states that 

EPA agrees that review of the MCL for 

beta particle and photon radioactivity is a 

priority and, as previously stated..., the 

Agency intends to review this standard 

within the general time frame established 

for the DOE submission of the licensing 

application for the Yucca Mountain site. 

EPA’s discussion in the final rule regarding 

the standards indicates that they realize the 

difficulty of applying the current MCLs to waste 

management and cleanup decisions. However, 

since the main focus of the MCLs was for water 

treatment plants, EPA chose to maintain the 

standards and review them at a later date. 

In a prospective analysis such as a 

performance assessment, there is clearly no intent 

to demonstrate compliance with a groundwater 

protection requirement by actual measurement at a 

water treatment plant. Therefore, there is no need 

to express the performance objective for alphaemitting radionuclides in terms of concentration 

limits. Indeed, using a single dose limit for all 

radionuclides provides a consistent and transparent 

approach. 
1-22

1.4.2.3 Selection of Groundwater 

Performance Objectives. The various 

groundwater protection performance objectives 

available for use in the TFF PA are summarized in 

Table 1-3. The selection of a site-specific 

groundwater protection performance objective was 

based on guidance in DOE Guide 435.1.

The performance objective Option 5 discussed 

in Section 1.4.2.2 for groundwater protection (i.e., 

4 mrem/yr EDE for all radionuclides) is more 

restrictive than the EPA’s 2000 revisions of the 

drinking water standards for radionuclides. In the 

EPA revised standards, the dose limit of 

4 mrem/yr applies only to beta/gamma-emitting 

radionuclides; the concentration limits for alphaemitting radionuclides are in addition to this dose 

limit. However, a 4 mrem/yr EDE in the Option 5 

performance objective applies to all radionuclides. 

To illustrate this point, the current EPA standard 

for radium of 5 pCi/L corresponds to an EDE of 

about 5 mrem/yr. The inclusion of alpha-emitting 

radionuclides in the dose limit provides a 

performance objective that is more restrictive than 

EPA’s revisions of drinking water standards and 

provides a common basis for determining 

concentration limits in groundwater for all 

radionuclides. 

The INEEL FFA/CO does not specifically 

address the issue of groundwater protection for 

LLW. Risk assessments done as part of the 

FFA/CO have used MCLs as applicable or 

relevant and appropriate requirements. The use of 

Option 5 does not preclude the use of MCLs. The 

MCLs for those radionuclides specified in 

40 CFR 141 also can be calculated in addition to 

the 4 mrem/yr EDE specified in Option 5. The use 

of MCLs is not considered necessary for a 

performance assessment. Option 5 provides a 

performance objective that is more protective of 

the public health than that offered by Option 4 and 

the use of MCLs. 

The use of Option 5 as the performance 

objective for groundwater protection simplifies the 

assessment of the groundwater pathway. The need 

to separate radionuclides in terms of dose or 

concentration is not deemed necessary. The use of 

Option 5 allows for a single dose limit to be 

applied to all radionuclides and provides a 

consistent approach with other potential pathways 

(i.e., air). Radionuclide concentrations will be 

available from the groundwater analysis to 

compare to the MCLs, which can then be 

compared to other groundwater programs. 

Table 1-3. Summary of groundwater protection performance objective options. 

Option 

Total Dose 

(mrem/yr) 

Man-made BetaGamma Dose 

(mrem/yr) 

Gross Alpha 

Concentration

(pCi/L) 

226Ra and 228Ra 

Concentration 

(pCi/L) 

Uranium 

Concentration

(μg/L) 

1 50a —b —b —b —b

2 25a —b —b —b —b

3 — 4c,d 15e,f 5g

 — 

4 — 4a,d 15e,f 5g

 — 

5 4a

 —b —b —b —b

a. Based on ICRP-30 dosimetry. 

b. Incorporated into total dose (EDE). 

c. Based on ICRP-2 dosimetry. 

d. For 3

H and 90Sr, the concentrations that yield 4 mrem/yr total body dose are specified to be 20,000 pCi/L and 8 pCi/L, 

respectively. 

e. Excluding uranium and radon isotopes. 

f. Excludes beta-gamma progeny of 238U, 235U, and 232Th. 

g. 226Ra and 228Ra, combined.
1-23

1.4.3 Intruder Analysis 

DOE Order 5820.2A, which has been 

superseded by DOE Order 435.1, contained a 

performance objective for the dose to individuals 

who inadvertently intrude into the LLW disposal 

facility after loss of active institutional control. 

DOE Manual 435.1-1 does not contain a specific 

performance objective for inadvertent intruders. 

As with the water resource impact assessment 

(Section 1.4.2), the approach in DOE Manual 

435.1-1 was chosen by DOE for consistency with 

NRC methods for LLW disposal and radiation 

protection principles articulated by the National 

Commission on Radiological Protection and the 

ICRP. In accordance with these principles, it is 

appropriate to assign a fraction (e.g., 25 mrem) of 

the 100-mrem/yr public dose performance 

measure to a particular practice (e.g., radioactive 

waste disposal) but it is not recommended to 

further fraction performance objectives to 

derivative activities (e.g., inadvertent intrusion). 

Thus, inadvertent intrusion is considered in the 

performance assessment but there is no specific 

performance objective for inadvertent intrusion 

(DOE 2000). 

DOE Manual 435.1-1, IV.P.(2)(h), states 

For purposes of establishing limits on the 

concentration of radionuclides that may 

be disposed of near-surface, the 

performance assessment shall include an 

assessment of impacts calculated for a 

hypothetical person assumed to 

inadvertently intrude for a temporary 

period into the low-level waste disposal 

facility. For intruder analyses, 

institutional controls shall be assumed to 

be effective in deterring intrusion for at 

least 100 years following closure. The 

intruder analyses shall use performance 

measures for chronic and acute exposure 

scenarios, respectively, of 100 mrem 

(1 mSv) in a year and 500 mrem (5 mSv) 

total effective dose equivalent excluding 

radon in air. 

Intruder analyses are to be performed as one 

of the mechanisms for establishing concentration 

limits for waste considered acceptable for nearsurface disposal. DOE intends to exercise control 

of the closure sites until it can be safely released 

pursuant to DOE Order 5400.5 (1993) (or 

10 CFR 834 when promulgated). Hence, intrusion 

is an accidental, temporary event. However, for 

purposes of conducting intruder analyses, the 

intrusion event should be considered to occur 

because of a lapse in institutional controls that 

would be remedied within a few years. The focus 

of the intruder analysis should be on the selection 

of reasonable scenarios and reasonably 

conservative parameters (DOE 2000). 

Intrusion is assumed to occur no sooner than 

100 yr following facility closure and should not be 

analyzed beyond 1,000 yr post-closure. The onset 

of intrusion can be extended beyond 100 yr if 

adequate justification is provided (e.g., continued 

DOE presence for facility decommissioning). 

Passive controls, such as marker systems and 

engineered features of the disposal system, may be 

effective in deterring accidental intrusion into the 

closure site and may be used as justification for 

extending the onset of intrusion (DOE 2000). 

Considering regional social customs and 

regional construction practices (e.g., well-drilling 

and excavation), the 500-mrem EDE should be 

used in assessing acute exposure from individual 

events that reasonably could occur at the site. 

Different individual events may be considered as 

appropriate for the site-specific conditions (DOE 

2000). 

The 100-mrem/yr EDE should be used in 

assessing chronic exposure from residing at or 

frequently visiting the disposal site. In the analysis 

of chronic exposure of a hypothetical intruder, 

doses should be assumed to come from external 

exposure to and inhalation and ingestion of 

materials exhumed from the site. Exposure may 

occur through a variety of pathways but need not 

include the consumption of contaminated 

groundwater or the irrigation of crops with 

contaminated groundwater. Groundwater 

consumption and crop irrigation are excluded 

because the impacts of groundwater contamination 

are evaluated separately in the all-pathways 

analysis, the water resource protection analysis, or 

both. Similarly, intruder doses need not include 

consideration of doses from airborne radon and its 
1-24

short-lived progeny because these are dealt with in 

the air pathway analysis. Doses from the progeny 

of radon that are deposited in the disposed waste 

should be included in the intruder analyses (DOE 

2000). 

1.4.4 ALARA Analysis 

DOE’s approach to radiation protection for 

closure sites is based on two key components. One 

component is the performance objectives 

described in Section 1.4.1, which specify 

maximum doses for various pathways. The other 

component is the ALARA principle, which 

requires doses to be maintained “as low as 

reasonably achievable.” 

DOE Manual 435.1-1, IV.P.(2)(f), states 

Performance assessments shall include a 

determination that projected releases of 

radionuclides to the environment shall be 

maintained as low as reasonably 

achievable (ALARA). 

The goal of the ALARA process is to attain 

the lowest practical dose level after taking into 

account social, technical, economic, and public 

policy considerations (DOE 2000). Thus, expected 

doses can never exceed the performance objectives 

but also must be as far below the performance 

objective as can reasonably be achieved. 

Therefore, in addition to providing a reasonable 

expectation that the performance objectives 

described in Section 1.4.1 will not be exceeded, 

the performance assessment also needs to show 

that the closure action is being conducted in a 

manner that maintains releases of radionuclides to 

the environment ALARA. The TFF ALARA 

analysis is given in Appendix D. 

1.4.5 Summary of Performance Criteria 

The previous sections have identified all the 

applicable performance objectives for the TFF PA 

in support of the DOE Tier 1 Closure Plan. Tables 

1-4 thru 1-6 are summaries of the adopted 

performance objectives for the TFF PA for each 

time interval of interest. 

Table 1-4. Summary of adopted performance objectives for the period of active institutional control. 

Compliance 

Interval Pathway Compliance Point Performance Objective 

All pathways 

(excluding radon) 

Maximum point of impact at 

INEEL boundary 

25 mrem/yr 

Air emissions 

(excluding radon) 

Maximum point of impact at 

INEEL boundary 

10 mrem/yr 

Radon emissions TFF surface 20 pCi/m2

/s 

Period of active 

institutional control 

2016 to 2116 

Groundwater Maximum point of impact at 

INEEL boundary 

4 mrem/yr 

Table 1-5. Summary of adopted performance objectives for the post-institutional control period. 

Compliance 

Interval Pathway Compliance Point 

Performance 

Objective 

All pathways 

(excluding radon) 

Maximum dose beyond 100 m 25 mrem/yr 

Air emissions 

(excluding radon) 

Maximum dose beyond 100 m 10 mrem/yr 

Radon emissions TFF surface 20 pCi/m2/s 

Post-institutional 

control 

2116 to 3016 

Groundwater Maximum dose beyond 100 m 4 mrem/yr 
1-25

Table 1-6. Summary of adopted performance objectives for inadvertent intruders. 

Compliance 

Interval Pathway Compliance Point 

Performance 

Objective 

Post-institutional 

control 

2116 to 3016 

All pathways 

(excluding radon in 

air and groundwater) 

Point of maximum dose 500 mrem (acute) 

100 mrem/yr (chronic) 

1.5 Summary of Key 

Assumptions 

It is inherent in modeling exercises to make 

simplifying assumptions in the models to represent 

the real system of components. These simplifying 

assumptions have been made in several areas of 

this PA. The key assumptions are presented in the 

following sections and are discussed in further 

detail under the applicable sections of this PA. 

This PA assumes 2016 as the closure date for 

the TFF. DOE is now moving to have the TFF 

closed by 2012 because of the INEEL accelerated 

cleanup program. The 4-yr difference will have 

little impact on the PA conclusions for compliance 

to the performance objectives. Long-lived 

radionuclides contribute significantly to the 

groundwater dose. The groundwater predicted 

peak doses will not increase but will occur 4 yr 

earlier than predicted. The intruder scenario doses 

will increase slightly in all cases. The dominant 

contributors to dose in these scenarios are the 

relatively short-lived radionuclides 137Cs/137mBa 

and 90Sr/90Y. Reducing the decay will slightly 

increase the final inventory in 3012. Based simply 

on the difference in time of decay, the dose 

estimates in the intruder scenario would increase 

by approximately 1%. 

1.5.1 Radionuclide Inventory 

The tank radionuclide source inventory 

estimate is based on the most recent analytical 

results from sampling of Tanks WM-182, 

WM-183, and WM-188 and historical data 

regarding the contents of the eleven 300,000-gal 

tanks. The concentrations for radionuclides 

lacking current analytical data have been estimated 

using the ORIGEN2 model (Croff 1980). The 

model used sodium-bearing waste as the nuclear 

fuel waste stream and radionuclide concentrations 

based on closure in 2016. The inventory is 

designed to be conservative based on data of liquid 

and solid tank contents and recent sampling data. 

The conservative approach was maintained by 

using the highest radionuclide concentrations 

(Tank WM-188) in the radionuclide inventory 

calculations, rather than averaging the recently 

sampled tanks. The least conservative scenario 

appears to overestimate the amount of residual 

radioactivity remaining in the tanks after closure. 

The cleaning of the first of 10 tanks shows 

removal effectiveness greater than anticipated and 

similar to the tank cleaning mockup results. 

Assumptions for the development of the 

radionuclide inventory and supporting 

documentation are given in Appendix A. 

Contaminated sand pads are located beneath 

Tanks WM-185 and WM-187 because of previous 

operational upsets. The contamination of these 

sand pads has been evaluated for mechanisms of 

contaminant removal associated with the filling 

and draining of the vaults during rainfall or 

snowmelt. Radionuclides with low distribution 

coefficients were assumed to be removed from the 

sand pad, while those with high distribution 

coefficients were assumed to remain in the sand 

pads. This analysis also is presented in 

Appendix A. 

Contamination contained in piping associated 

with the transfer of material throughout the facility 

also was evaluated to determine the potential 

impact of this source on future doses. Appendix A 

gives an estimate of the final composition of 

contamination in the process lines after a series of 

washing and decontamination steps. 

1.5.2 Degradation Analysis 

An analysis of the degradation of the vaults, 

grout, tanks, and piping was completed for the 
1-26

TFF PA. This analysis, presented in Appendix E, 

contains several key assumptions necessary for the 

simplification of the analysis. The analysis 

indicated that the concrete vaults turn to rubble 

approximately 500 yr after closure. The grout and 

tanks were predicted to turn to rubble after 

5,000 yr. Using the degradation analysis, key 

assumptions were made in the conceptual model 

such that the concrete vault was assumed to be 

completely degraded at 100 yr and the tanks and 

grout at 500 yr. These assumptions show that a 

very conservative analysis is being conducted in 

relation to the results of the degradation analysis. 

The piping associated with the tank systems at 

the TFF was modeled based on the assumption 

that the stainless-steel piping and associated grout 

inside and enveloping the pipe in the secondary 

containment degrade similar to the tank 

environment and release contaminants beginning 

500 yr after closure. The degradation analysis for 

the piping is presented in Appendix E. 

1.5.3 Groundwater Model 

Based on the availability of geotechnical 

information and the inherent uncertainty regarding 

this information, several assumptions were 

necessary for the model simulations. Listed below 

are the key hydrologic assumptions inherent in the 

groundwater and contaminant transport modeling. 

• The hydrostratigraphy presented by numerous 

United States Geological Survey (USGS) 

reports is accurate and reliable, and thus forms 

the basis for numerical simulations of 

groundwater flow and contaminant transport. 

• Literature-based values for recharge indicate 

that the Big Lost River and the INTEC 

percolation ponds are the major sources of 

recharge to the unsaturated zone and aquifer at 

INTEC. 

• Infiltration from the Big Lost River and the 

INTEC percolation ponds is constant over 

time. The amount of seepage from the Big 

Lost River was based on stream gaging 

studies. INTEC plant discharge records were 

used for the percolation pond estimates. 

• The percolation ponds were only used in the 

calibration of perched water bodies at INTEC. 

The ponds will be closed in the future; 

therefore, the ponds are not used the flow 

model for transport of radionuclides at INTEC 

from the TFF. The water infiltration from the 

ponds was shut off in the model after the 

assumed TFF closure date of 2016. 

• Water usage from the Big Lost River and 

flood diversion of the river were assumed to 

be constant over the time of the simulations. 

However, separate analysis for flooding of the 

Big Lost River was evaluated. 

• The fundamental assumption for simulating 

water transport and developing the initial 

hydrologic conditions for contaminant 

transport simulations is that there are no 

perched water zones due solely to infiltrating 

water from precipitation. 

• The interbeds are the controlling geological 

units responsible for the formation of perched 

water zones whether the hydraulic 

conductivity of a sedimentary interbed might 

be smaller than that of an overlying basalt 

flow or alterations in the baked zone exist 

between two flows that contribute to reduced 

vertical hydraulic conductivity. 

• Because of the lack of direct information on 

the hydraulic conductivity of the interbed 

units, values were assigned in modeling 

simulations based on lithologic properties, 

residual moisture contents, and the existence 

of perched water in nearby monitoring wells. 

• Hydraulic conductivity values for the basalt 

units were based on the thickness. Thin basalt 

units were assigned higher hydraulic 

conductivity values. 

• Fractured basalts were treated as a porous 

medium with low residual moisture contents 

and unsaturated hydraulic characteristic curves 

that yield high water releases at low capillary 

pressures. 

• For calibration involving seepage losses from 

the Big Lost River and the INTEC percolation 

ponds, the location and extent of perched 
1-27

water zones is controlled by variations in 

hydraulic conductivity. 

• Based on the volume of water available from 

the Big Lost River seepage losses, compared 

with estimates from other potential sources of 

recharge, the source of the perched water in 

the northern portion of the INTEC facility was 

due primarily to the Big Lost River seepage. 

• Similarly, the INTEC percolation ponds were 

the principle source of water for the perched 

zones in the southern portion of INTEC. 

• Since there was a lack of site-specific 

information, such as Kd values for specific 

radionuclides or dispersion coefficients for 

geologic materials underlying the site, 

information from similar sites or previously 

conducted PA investigations was incorporated 

into the model. 

1.6 Summary 

This document has been prepared to describe 

the methods and results of the PA for the TFF. The 

results of the analyses indicate compliance with 

the established performance objectives and 

provide reasonable assurance that public health 

and safety will be protected. 

The basic concept of a performance 

assessment is to develop a simplified 

representation of the physical setting at the INEEL 

using research data, assumptions, documented 

information, and modeling techniques in an 

attempt to understand contaminant transport 

within the physical system and along potential 

pathways of human exposure. This PA begins with 

site-specific physical data that are used to 

formulate a conceptual model and provide a listing 

of the performance objectives. Information in 

Section 2 includes description of geography, 

demography, meteorology, climatology, ecology, 

geology, and hydrology. Also included are the 

design features of the TFF and the waste 

characteristics. The specific processes in the 

conceptual model are then described 

mathematically. 

The PA includes scenarios in which an 

inadvertent intruder receives both an acute dose 

and a chronic dose. Section 5 defines the intruder 

scenarios and gives the results of intruder scenario 

analysis. 

The results of the PA are described in 

Section 6. The results show reasonable assurance 

that the potential dose will be below the regulatory 

standards. 

Section 7 provides an evaluation of the PA 

results with respect to the performance objectives. 

A PA does not produce an absolute answer but 

rather a bounding approximation. This 

approximation is dependent on the accuracy of the 

input data, the conceptual and mathematical 

models, and the computer codes. A 

sensitivity/uncertainty analysis was prepared as 

part of this PA. 

Section 7 contains a sensitivity/uncertainty 

analysis based on several scenarios involving the 

sensitive parameters for the transport of 

radionuclides. Different inventories, sorption 

coefficients, hydraulic diffusivity values, and 

infiltration rates are varied in numerical 

simulations for four different scenarios and the 

results are also discussed. Selection criteria are 

discussed for the radionuclide inventory, 

radionuclide screening analysis, TFF degradation 

analysis, modeled releases, and the groundwater 

analysis. Four scenarios were established for this 

PA: best, realistic, conservative, and worst-case. 

The conservative scenario is shown in Section 3 as 

the compliance scenario and is compared to the 

performance objectives. The criteria for the 

selection and quantification of four radionuclide 

inventories and transport scenarios are also 

presented in this section. 
2-1 

2. CLOSURE FACILITY AND SITE CHARACTERISTICS

This section provides descriptive information 

and data for the INEEL site, environment, and 

TFF site, and the characteristics of the residual 

waste. This description gives the basis for the TFF 

conceptual model and an understanding of the 

method of analysis. The emphasis of this section is 

on the characteristics that are important to the 

performance of the closure action. 

2.1 Site Characteristics 

2.1.1 Geography and Demography 

2.1.1.1 Closure Action Location. The 

INEEL is located in southeastern Idaho, on the 

west-central portion of the Eastern Snake River 

Plain (ESRP) (Figure 2-1). Included in its 890 mi2

(2,300 km2

) of area are portions of five Idaho 

counties (Bingham, Bonneville, Butte, Clark, and 

Jefferson) (see Figure 2-2). The nearest INEEL 

boundaries are 22 mi (35 km) west of Idaho Falls, 

23 mi (37 km) northwest of Blackfoot, 44 mi 

(71 km) northwest of Pocatello, and 7 mi (10 km) 

east of Arco, Idaho. There are no permanent 

residents within an 11-mi (18-km) radius of 

INTEC and there is a relatively low population 

base in the surrounding areas. The significant 

communities in close proximity to the INEEL are 

Atomic City (1 mi [2 km]), population 25; Arco 

(7 mi [10 km]), population 1,026; Idaho Falls 

(22 mi [35km]), population 50,730; Blackfoot 

(23 mi [37 km]), population 10,419; and Pocatello 

(44 mi [71 km]), population 51,466. The INEEL 

also lies approximately equidistant from three 

larger metropolitan areas: Salt Lake City, Utah 

(211 mi [340 km]), population 181,743; 

Boise, Idaho (257 mi [414 km]), population 

185,787; and Butte, Montana (214 mi [344 km]) 

population 33,892 (Sehlke and Bickford 1993, 

Department of Commerce 2001).

The INEEL HLW Program is conducted at 

INTEC, located in the south-central portion of the 

INEEL. INTEC occupies approximately 200 acres 

(80 ha) and consists of more than 150 buildings. 

Primary facilities include storage, treatment, and 

laboratory facilities for spent nuclear fuel, mixed 

HLW, and mixed transuranic waste/sodiumbearing waste (Figure 2-3). Located outside the 

INTEC perimeter fence are parking areas, a 

helicopter landing pad, the wastewater treatment 

lagoon, various pits, and percolation ponds. These 

areas occupy approximately 55 acres (22 ha). 

INTEC began operations in 1953 and was, 

historically, a fuel reprocessing facility for defense 

projects, research, and storage of spent nuclear 

fuel. In 1992, DOE decided to end the fuelreprocessing mission at INTEC. This decision led 

to the phase-out of fuel dissolution, solvent 

extraction, product denitration, and other 

processes. The current mission of INTEC is to 

safely receive and store spent nuclear fuel and 

radioactive wastes, prepare spent nuclear fuel for 

shipment to an off-site repository, develop new 

technologies for waste and waste management for 

DOE, and remediate past environmental releases. 

2.1.1.2 Closure Site Description. The 

TFF lies within the boundaries of the INTEC 

facility in the northern portion of the facility. The 

HLLW generated from fuel reprocessing 

operations were stored in stainless-steel storage 

tanks at the TFF. The TFF comprises nine 

300,000-gal tanks, two 318,000-gallon tanks, and 

four 30,000-gal tanks (Figure 2-4). All of the 

tanks, except for the 30,000-gal tanks, are 

contained within concrete vaults under the ground 

surface. The 30,000-gal tanks were directly buried 

in the ground without concrete vaults.

The principal surface materials at the INEEL 

are basalt, alluvium, lakebed or lacustrine 

sediments, slope wash sediments and talus, silicic 

volcanic rocks, and sedimentary rocks. The natural 

plant life consists mainly of sagebrush and various 

grasses. The vegetation of the INEEL is limited by 

soil type, meager rainfall, and extended drought 

periods. Only a few deciduous trees, located 

principally along the Big Lost River, exist on the 

INEEL. The most prominent ground cover is a 

mixture of sagebrush (Artemisia tridentata) and a 

variety of grasses. Lanceleaf rabbitbrush 

(Crysothamnus viscidiflorus) covers about 80% of 

the INEEL and can be found in any given area. 

The soil at the INEEL TFF site is previously 

disturbed sandy gravel; the flat terrain precludes 

erosion (INEEL 2001a. 
2-2 

Figure 2-1. Idaho map showing the location of the INEEL on the Eastern Snake River Plain.
2-3 

Figure 2-2. INEEL vicinity map. 

INEEL

LINCOLN

MINIDOKA

POWER

BINGHAM

BLAINE

BUTTE

CLARK

FREMONT

JEFFERSON

MADISON

BONNEVILLE

CARIBOU

BANNOCK

Camas National

Wildlife Refuge

Mud Lake Wildlife

Management Area

Mud Lake

Lost River Range

Lemhi Range

Bitterroot Range

Craters of the Moon 

Wilderness Area

INEEL 

Diversion Dam

Atomic City

Rigby

Idaho Falls

Fort Hall 

Indian 

Reservation

Pocatello

American Falls

Reservoir

Shelley

Howe

American Falls

Snake River

Big Lost River

Arco

Blackfoot

INEEL

Vicintiy Map

Miles

Kilometers

0

0

30

40
2-4 

Figure 2-3. Major facilities at INTEC. 

Figure 2-4. Location of the TFF tanks and vaults at INTEC. 

6 16

HEPA BU IL D IN G

C PP-6 49

C P P -7 39

CPP- 692

C PP-6 04

C PP-6 04

VES-WL-133

C P P -6 94

CPP-763

C P P -6 34

C PP-6 28

C PP-6 19

C PP-6 18

CPP-717-A

CPP-717-BCPP-717-CCPP-717-D

CPP-659

CPP-654

T-6

BEECH STREET

FIR STREET

WM-180 WM-181

WM-185 WM-186

WM-184

WM-187 WM-189

WM-188 WM-190

WM-182 WM-183

LO DG E POLE STREET

FI R STREET

REDWOOD STREET
2-5 

INTEC is located on an alluvial plain 

approximately 200 ft (60 m) from the Big Lost 

River channel, near the channel intersection with 

the INEEL’s Lincoln Boulevard. INTEC is 

surrounded by a stormwater drainage ditch system 

(DOE-ID 1998). Stormwater runoff from most 

areas of INTEC flows through ditches to an 

abandoned gravel pit on the northeast side of 

INTEC. From the gravel pit, the runoff infiltrates 

the ground. The system is designed to handle a 

1-in-25-yr, 24-hr storm event. Because the land is 

relatively flat (slopes of generally less than 1%) 

and annual precipitation is low, stormwater runoff 

volumes are small and are generally spread over 

large areas where they may evaporate or infiltrate 

the ground surface. 

The INEEL is located in the Pioneer Basin. 

This closed drainage basin includes three main 

streams: Big Lost River, Little Lost River, and 

Birch Creek. These three streams drain the 

mountain areas to the north and west of the 

INEEL, although most flow is diverted for 

irrigation in the summer months before it reaches 

the site boundaries. Flow that reaches the INEEL 

infiltrates the ground surface along the length of 

the streambeds in the spreading areas at the 

southern end of the INEEL and, if the stream flow 

is sufficient, in the ponding areas (playas or sinks) 

in the northern portion of the INEEL. During dry 

years, there is little or no surface water flow on the 

INEEL. Because the Pioneer Basin is a closed 

drainage basin, water does not flow off the 

INEEL. Rather, the water either infiltrates the 

ground surface to recharge the aquifer or is 

consumed by evapotranspiration. 

The Big Lost River flows southeast from 

Mackay Dam, past Arco, and onto the Snake River 

Plain (see Figure 2-2). Near the INEEL’s 

southwestern boundary, a diversion dam prevents 

flooding of downstream areas during periods of 

heavy runoff by diverting water to a series of 

natural depressions or spreading areas (DOE-ID 

1995b). During periods of high flow or low 

irrigation demand, the Big Lost River continues 

northeastward past the diversion dam, passes 

within 200 ft (60 m) of INTEC, and ends in a 

series of playas 15 to 20 mi (24 to 30 km) 

northeast of INTEC, where water infiltrates the 

ground surface. 

Flow from Birch Creek and the Little Lost 

River infrequently reaches the INEEL. In the 

summer months, the water in Birch Creek and the 

Little Lost River is diverted for irrigation prior to 

reaching the INEEL. During periods of unusually 

high precipitation or rapid snowmelt, water from 

Birch Creek and the Little Lost River may enter 

the INEEL from the northwest and infiltrate the 

ground. 

Flood studies at the INEEL have examined the 

flooding potential at INEEL facilities due to 

failure of the Mackay Dam, 45 mi (72 km) 

upstream from the INEEL (Koslow and Van 

Haaften 1986). The USGS made preliminary 

estimates of the 100-yr flood plain along the Big 

Lost River (Berenbrock and Kjelstrom 1998). 

DOE commissioned additional studies to refine the 

100-yr flood plain and delineate the 500-yr flood 

plain, including a two-dimensional model analysis 

and a paleohydrologic and geomorphic assessment 

of the flood risk along the Big Lost River (Ostenaa 

et al. 1999). There is no record of any historical 

flooding at INTEC since 1952. Section 2.1.5.3 

discusses in detail the potential for flooding at 

INTEC. 

The Snake River Plain has a relatively low 

rate of seismicity, whereas the surrounding basin 

and range has a fairly high rate of seismicity 

[Woodward-Clyde Federal Services (WCFS) 

1996]. The primary seismic hazards from 

earthquakes to INEEL facilities consist of the 

effects from ground shaking and surface 

deformation (surface faulting and tilting). Other 

potential seismic hazards such as avalanches, 

landslides, mudslides, and soil liquefaction are not 

likely to occur at the INEEL because the local 

geological conditions and terrain are not 

conducive to these types of hazards. Based on the 

seismic history and geologic conditions, 

earthquakes greater than moment magnitude of 5.5 

and associated strong ground shaking and surface 

fault ruptures are not likely to occur within the 

Snake River Plain [Woodward-Clyde Consultants 

(WCC) 1990; WCFS 1996]. However, moderate to 

strong ground shaking could affect the INEEL 

from earthquakes in the surrounding basin and 

range. Section 2.1.4.2 provides detailed 

discussions on seismology. 
2-6 

Volcanic hazards include the effects of lava 

flows, fissures, uplift, subsidence, volcanic 

earthquakes, and ash flows or airborne ash 

deposits. Most of the basalt volcanic activity 

occurred from 4 million to 2,100 yr ago in the 

INEEL area. The most recent and closest volcanic 

eruption occurred at the Craters of the Moon 

National Monument, 26.8 mi (43.1 km) southwest 

of INTEC (Kuntz, Covington, and Schorr 1992). 

Based on a probability analysis of the volcanic 

history in and near the south-central INEEL area, 

the Volcanism Working Group estimated that the 

conditional probability that basaltic volcanism 

would affect a south-central INEEL location is 

less than once per 100,000 yr or longer 

(Volcanism Working Group 1990). The 

probability is associated primarily with the Axial 

Volcanic Zone and the Arco Volcanic Rift Zones. 

INTEC is located in a lesser lava flow hazard area 

of the INEEL, more than 5 mi (8 km) from the 

Axial Volcanic Zone and any volcanic vent 

younger than 400,000 yr. Section 2.1.4.3 provides 

additional discussions on the volcanic hazards. 

2.1.1.3 Population Distribution. 

Population growth surrounding the INEEL (i.e., 

within a seven-county region comprising Bannock, 

Bingham, Bonneville, Butte, Clark, Jefferson, and 

Madison counties) has paralleled statewide growth 

from 1960 to 1990. During this time, the regional 

population increased an average of approximately 

1.40% annually, while the annual growth rate for 

the state was 1.53% (Department of Commerce 

1995). From 1990 to 1999, state population 

growth accelerated to nearly 2.45% per year, but 

the regional population growth dropped to 1.32% 

(Department of Commerce 2000). Population 

growth for the state is projected to slow after the 

year 2000 (Department of Commerce 1997). 

Table 2-1 lists 1990 and 2000 census data for the 

counties surrounding the INEEL and growth 

projections for 2010 and 2025. The projections are 

based on an annual growth rate of 1.05%, derived 

from the average of the projected state growth 

rates from 2000 to 2025 (Department of 

Commerce 1997).

In the year 2000, Bannock and Bonneville 

counties had the largest populations in the region, 

and together accounted for 46.9% of the total 

regional population (Department of Commerce 

2001). Butte and Clark are the most sparsely 

populated counties and together contained 1.57% 

of the regional population in 2000. The largest 

cities in the region are Pocatello (Bannock 

County) and Idaho Falls (Bonneville County), 

with year 2000 populations of approximately 

51,466 and 50,730, respectively. 

Figures 2-5 through 2-7 show population 

densities, based on the 1990 census, for the years 

2000 through 2020 at 10-yr intervals for the 50-mi 

(80-km) radius around INTEC (INEEL 1999b). 

The nearest populated area to the INEEL is 

Atomic City, population about 25 (Department of 

Commerce 2001), located approximately 1 mi 

(2 km) from the southern INEEL boundary and 

about 11 mi (18 km) from INTEC. 

No permanent residents live within a 10-mi 

(20-km) circle centered at INTEC on the INEEL 

(DOE 2002). No cities or towns are within 10 mi 

(20 km) of the INTEC TFF. However, several 

INEEL facilities, such as the Central Facilities 

Area (CFA), the Test Reactor Area (TRA), and 

RWMC are within 10 mi (20 km) of the INTEC 

TFF. Also, the Experimental Breeder Reactor I 

(EBR-I), a National Historic Landmark, is located 

southwest and within 10 mi (20 km) of the INTEC 

TFF. 

The INEEL has become a major part of the 

eastern Idaho community and economy. During 

fiscal year 2001, the INEEL employed an average 

of over 8,000 people and had a budget of nearly 

$1 billion (INEEL 2001b). The INEEL contractors 

are the largest employers in the area and one of the 

five largest in the state. The 2001 impact analysis 

indicates that INEEL activities and its workforce 

support an additional 10,250 jobs statewide 

(INEEL 2001b). Therefore, over 18,000 Idaho 

jobs can be attributed to the facility. The INEEL 

workforce is projected to decrease by 

approximately 450 people by fiscal year 2005 

(INEEL 2000b). 

2.1.1.4 Uses of Adjacent Lands. The 

INEEL occupies approximately 890 mi2

(570,000 acres or 230,000 ha) of land in Bingham, 

Bonneville, Butte, Clark, and Jefferson counties in 

southeastern Idaho. Approximately 2% of this land 

(11,400 acres [4,610 ha]) has been developed to
2-7 

Table 2-1. Regional population of the INEEL; selected years 1990–2025. 

County 1990 a 2000 b 2010 2025 

Bannock 66,026 75,565 83,886 98,116 

Bingham 37,583 41,735 46,329 54,187 

Bonneville 72,207 82,522 91,608 107,147 

Butte 2,918 2,899 3,217 3,762 

Clark 762 1,022 1,135 1,328 

Jefferson 16,543 19,155 21,264 24,870 

Madison 23,674 27,467 30,491 35,663 

Total 219,713 250,365 277,930 325,073 

a. Source: Department of Commerce 1990. 

b. Source: Department of Commerce 2001. 

Figure 2-5. 2000 population distribution for the area surrounding INTEC (INEEL 1999b). 

216

2 
2-8 

Figure 2-6. 2010 projected population distribution for the area surrounding INTEC (INEEL 1999b). 

Figure 2-7. 2020 projected population distribution for the area surrounding INTEC (INEEL 1999b). 
2-9 

support INEEL facility and program operations 

associated with energy research and waste 

management activities (DOE 1995). INEEL 

operations are performed within the Site’s primary 

facility area (CFA, TRA, INTEC, etc.), which 

occupies 2,032 acres (822.3 ha). A 345,000-acre 

(140,000 ha) security and safety buffer zone is 

located around the core development area, which 

also accommodates environmental research and 

ecological and socio-cultural preservation. 

Approximately 6% of the INEEL 

(34,000 acres [14,000 ha]) is devoted to utility 

rights-of-way and public roads, including 

Highway 20 (which runs east and west and crosses 

the southern portion of the INEEL), Highway 26 

(which runs southeast and northwest intersecting 

Highway 20), and Idaho State Highways 22, 28, 

and 33 (which cross the northeastern part of the 

INEEL) (DOE 1995). 

Up to 340,000 acres (140,000 ha) of the 

INEEL is leased for cattle and sheep grazing 

(DOE 1995); the Bureau of Land Management 

(BLM) administers grazing permits. However, 

grazing of livestock is prohibited within 0.5 mi of 

any primary facility boundary and within 2 mi 

(3 km) of any nuclear facility. In addition, 

900 acres (400 ha) located at the junction of Idaho 

State Highways 28 and 33 are used by the U.S. 

Sheep Experiment Station as a winter feedlot 

(DOE-ID 1997a). Figure 2-8 shows land use in the 

vicinity of the INEEL. 

On July 17, 1999, the Secretary of Energy and 

representatives of the U.S. Fish and Wildlife 

Service, BLM, and Idaho State Fish and Game 

Department designated 73,263 acres (29,648 ha) 

of the INEEL as the Sagebrush Steppe Ecosystem 

Reserve (DOE 1999c). In 1995, the National 

Biological Service listed the ungrazed sagebrush 

steppe ecosystem in the Intermountain West and 

big sagebrush (Artemisia tridentata) in Idaho’s 

Snake River Plain as critically endangered (Noss, 

LaRoe, and Scott 1995). The INEEL Sagebrush 

Steppe Ecosystem Reserve was designated to 

ensure this portion of the ecosystem receives 

special, scientifically-controlled consideration. 

Conservation management in this area is intended 

to maintain the current vegetation and provide the 

opportunity for study of an undisturbed sagebrush 

steppe ecosystem. Traditional rangeland uses, such 

as livestock grazing, which exist in a portion of the 

area, are continuing under this management 

designation. The designated INEEL Sagebrush 

Ecosystem Reserve is located in the northwest 

portion of the area. The southern boundary of the 

Reserve, which runs east and west along section 

lines, is about 11 mi (18 km) north of INTEC at 

the closest point (DOE 2002). 

Recreational uses of the INEEL include public 

tours of general facility areas and EBR-I, a 

National Historic Landmark. Controlled hunting 

also is permitted on the INEEL, but is restricted to 

0.5 mi inside the Site boundary. These restricted 

hunts are intended to assist the Idaho Department 

of Fish and Game in reducing crop damage caused 

by wild game on adjacent private agricultural 

lands. The INEEL also is designated as a National 

Environmental Research Park, functioning as a 

field laboratory set aside for ecological research 

and evaluation of the environmental impacts from 

nuclear energy development (DOE 2002). 

The INEEL does not lie within any of the land 

boundaries established by the Fort Bridger Treaty 

of 1868. The entire INEEL is land occupied by 

DOE; therefore, the provision in the Fort Bridger 

Treaty that allows the Shoshone-Bannock Tribes 

to hunt on unoccupied lands of the United States 

does not presently apply to any INEEL land (DOE 

2002). 

Land use at the INEEL is in a state of 

transition. Emphasis is moving toward radioactive 

and hazardous waste management, environmental 

restoration and remedial technologies, and 

technology transfer, resulting in more development 

of the INEEL within some facility areas and less 

development in others. DOE has projected land 

use scenarios at the INEEL for the next 25, 50, 75, 

and 100 yr. Future development is projected to 

take place in the central portion of the INEEL 

within existing facility areas. 

For further review, see the Long-Term Land 

Use Future Scenarios for the Idaho National 

Engineering Laboratory (DOE-ID 1995a) and the 

Idaho National Engineering and Environmental 

Laboratory Comprehensive Facility and Land Use 

Plan (DOE-ID 1997a). 
2-10 

Big Lost River

Birch Creek

Little Lost River

Big Lost River Playas

Figure 2-8. Selected land use of the INEEL and surrounding vicinity. 

Approximately 75% of the land adjacent to the 

INEEL is owned by the federal government and 

administered by BLM. Land uses on this federallyheld land consist of wildlife management, mineral 

and energy production, grazing, and recreation. 

The State of Idaho owns approximately 1% of the 

adjacent land. This land also is used for wildlife 

management, grazing, and recreation. The 

remaining 24% of the land adjacent to the INEEL 

is privately owned and is primarily used for 

grazing and crop production (INEEL 1999b). 

Small communities and towns located near the 

INEEL boundaries include Mud Lake to the east; 

Arco, Butte City, and Howe to the west; and 

Atomic City to the south. The larger communities 

of Idaho Falls/Ammon, Rexburg, Blackfoot, and 

Pocatello/Chubbuck are located to the east and 

southeast of the INEEL Site. The Fort Hall Indian 

Reservation is located southeast of the INEEL 

Site. 

All county plans and policies encourage 

development adjacent to previously developed 
2-11 

areas to minimize the need to extend infrastructure 

improvements and to avoid urban sprawl. Because 

the INEEL is remotely located from most 

developed areas, its lands and adjacent areas are 

not likely to experience residential and 

commercial development and no new development 

is planned near the INEEL Site. However, 

recreational and agricultural uses are expected to 

increase in the surrounding area in response to 

greater demand for recreational areas and the 

conversion of rangeland to cropland (Sehlke and 

Bickford 1993). 

The four most prominent tourist/recreation 

areas or attractions in the INEEL area include: 

• Yellowstone National Park, which is 

approximately 72.5 mi (117 km) northeast of 

the INEEL and 99.5 mi (160 km) from INTEC 

• EBR-I, which is situated on the INEEL 

• Craters of the Moon National Monument, 

which is located approximately 19 mi (31 km) 

southeast of the INEEL 

• The resort areas of Ketchum and Sun Valley, 

which are approximately 59.5 mi (95.8 km) 

west of the INEEL and 72 mi (120 km) from 

INTEC (INEEL 1999b). 

Other recreation and tourist attractions in the 

region surrounding the INEEL Site include Hell’s 

Half Acre Wilderness Study Area, Black Canyon 

Wilderness Study Area, Camas National Wildlife 

Refuge, Market Lake State Wildlife Management 

Area, North Lake State Wildlife Management 

Area, Targhee and Challis National Forests, 

Sawtooth National Recreation Area, Sawtooth 

Wilderness Area, Sawtooth National Forest, Grand 

Teton National Park, Jackson Hole recreation 

complex, and the Snake River. 

The Idaho National Engineering and 

Environmental Laboratory Comprehensive 

Facility and Land Use Plan (DOE-ID 1997a) and 

the Final Record of Decision, Idaho Nuclear 

Technology and Engineering Center (DOE-ID 

1999) describe the land use for the INEEL and 

INTEC. Land use at the INEEL is currently 

government-controlled industrial use. The term 

controlled means that unrestricted public access to 

INTEC and the INEEL is not available. Presently, 

access to INEEL facilities requires proper 

clearance, training, or escort and controls to limit 

the potential for unacceptable exposures. A 

security force limits access to approved personnel 

and visitors. These controls are estimated to be in 

place for the next 100 yr. 

Future land use scenarios are identified in the 

Long-Term Land Use Future Scenarios for the 

Idaho National Engineering Laboratory (DOE-ID 

1995a). This document was developed using a 

stakeholder process that involved a public 

participation forum, a public comment period, and 

the INEEL Citizens Advisory Board. Following 

review and comment by the public participation 

forum, the document underwent a 30-day public 

comment period and was subsequently submitted 

to the Citizens Advisory Board for review and 

recommendations. To date, no recommendations 

have been received for residential use of any 

portion of the INEEL until at least the year 2095. 

Until 2095, the reasonably anticipated future use 

for the TFF is as a government-controlled 

industrial facility (DOE-ID 1999). 

Planning assumptions in the INEEL 

Comprehensive Facility and Land Use Plan

(DOE-ID 1997a) are that the INEEL will remain 

under government control for at least the next 

100 yr. Future government management and 

control becomes increasingly uncertain with time. 

No residential development will be allowed to 

occur within INEEL boundaries during the next 

100 yr. 

INTEC was one of the facilities that had a 

future use scenario projected in the Long-Term 

Land Use Future Scenarios for the Idaho National 

Engineering Laboratory (DOE-ID 1995a). The 

scenarios are broken down into 25-yr increments. 

• Present: Interim storage of spent nuclear fuels, 

disposition of fuels, managing and improving 

waste, and water management techniques. 

• 25-yr: Continued use as industrial area, 

planned new waste treatment facility. 

• 50-yr: Approaching end of useful life if no 

new mission identified, decontamination and 
2-12 

decommissioning with all or selected areas for 

restricted industrial use. 

• 75-yr: Standby mode for restricted industrial 

use, reuse permitted but no new development 

outside existing fence line. 

• 100-yr: Continuation as a restricted industrial 

area. 

2.1.2 Meteorology and Climatology 

Meteorological data have been collected 

periodically at over 45 locations on and near the 

INEEL since 1949. The longest and most complete 

record of air temperature and precipitation 

observations (over 35 yr) at the INEEL was 

collected from the weather station at CFA. The 

CFA station is located approximately 3 mi (5 km) 

south of INTEC. Differences in climate between 

the CFA monitoring station and INTEC are 

minimal. INTEC and CFA are at approximately 

the same terrain elevation and have the same 

exposure to wind, snow, and cloud cover. 

The National Oceanic and Atmospheric 

Administration (NOAA) Air Resources 

Laboratories conducts most of the meteorological 

monitoring within 50 mi (80 km) of the INEEL. 

An overview of climatological data is available 

from data summaries collected from the CFA 

monitoring station. A summary of the climatology 

of the INEEL is available in Climatography of the 

Idaho National Engineering Laboratory (Clawson, 

Start, and Ricks 1989). 

2.1.2.1 Temperature. Temperatures at the 

INEEL vary widely over the course of the year. 

Records for CFA indicate that the highest and 

lowest daily maximum temperatures range from 

101°F (38.3°C) to −47°F (−44°C), respectively. 

The average annual temperature at the INEEL 

exhibits a gradual seven-month increase, 

beginning with the first week in January and 

continuing through the third week in July. During 

the summer months of April through October, the 

average monthly temperature varies from 41 to 

68°F (5 to 20°C). The temperature then decreases 

over the course of five months until the minimum 

average temperature is reached again. 

2.1.2.2 Wind. The prevailing wind direction 

at INTEC and at most locations on the INEEL is 

southwesterly (DOE-ID 1995a). In summer, a 

sharp reversal in wind direction occurs daily; 

winds from the southwest predominate during 

daylight hours, and northeasterly winds 

predominate at night. The reversals normally 

occur shortly after sunrise and sunset. The wind 

rose diagrams in Figure 2-9 show the annual wind 

flow at the INEEL at three of the meteorological 

monitoring sites from 1988 to 1992. Multi-year 

wind roses exhibit little variability and are 

representative of typical patterns.

The average monthly wind speed varies from 

approximately 3.1 mph (4.9 km/h) in December to 

9.3 mph (15 km/h) in April and May (DOE-ID 

1995a). The highest hourly average speed was 

recorded at 51 mph (82 km/h), from the westsouthwest. Strong wind gusts can occur during 

thunderstorm activity. On average, strong wind 

gusts occur 2 or 3 days per month during June, 

July, and August. 

The highest instantaneous speed, recorded 

18 ft (5.5 m) aboveground at CFA, was 78 mph 

(130 km/h), with the wind from the westsouthwest. Calm conditions prevail 11% of the 

time. 

Atmospheric particulate matter is routinely 

monitored using low-volume air sampling stations 

at various locations across the INEEL and a total 

suspended particulate monitor at CFA. In 1989 and 

1990, the INEEL mean value from the low-volume 

air samplers ranged from 17 to 20 μg/m3

, and the 

CFA annual average ranged from 28 to 40 μg/m3

(Hoff et al. 1990, 1991). 

2.1.2.3 Precipitation. The average annual 

precipitation at CFA is 8.7 in. (22 cm). The 

highest recorded annual amount of precipitation 

recorded was 14.4 in. (36.6 cm) in 1963, and the 

lowest amount was 4.5 in. (11 cm) in 1966. A 

precipitation peak of approximately 1.2 in./mo 

(3.0 cm/mo) is associated with thunderstorms in 

May and June each year (DOE-ID 1995a). Other 

months generally receive one-half or less of this 

amount. 
2-13 

Figure 2-9. Annual average wind direction and speed at INEEL meteorological monitoring stations. 

Snowfall is a substantial contributor to total 

annual precipitation. Snowfall and snow depth 

records are available from the CFA monitoring 

station. Snowfall ranges from 6.7 to 60 in./yr 

(17 to 200 cm/yr) with an annual average of 28 in. 

(71 cm). Although snow falls mostly during 

November through April, it does occasionally fall 

during May, September, and October (Clawson, 

Start, and Ricks 1989). The maximum average 

monthly snowfall is 6.4 in. (16 cm), occurring in 

December. The water content of melted snow 

probably contributes between one-quarter and onethird of the average annual precipitation (Sehlke 

and Bickford 1993). 

2.1.2.4 Evaporation. The potential annual 

evaporation from a saturated ground surface at the 

INEEL is approximately 36 in. (91 cm), with 80% 

of the evaporation occurring between May and 

October. During July, the warmest month of the 

year, the daily potential evaporation rate is 

approximately 0.2 in. (0.5 cm) (Hull 1989). 

Evaporation occurring during the remainder of the 

year is small. Actual evaporation rates are much 

lower than potential rates because the ground 

surface is rarely saturated. Transpiration by the 

native vegetation of the Snake River Plain is 

estimated at 5.9 to 9.1 in./yr (15 to 23 cm/yr). 

From late winter to spring, precipitation is most 

likely to infiltrate into the ground because of the 
2-14 

low evapotranspiration rates (Mundorff, 

Crosthwaite, and Kilburn 1964). For evaporation 

from surface water bodies (ponds), a pan 

evaporation rate of approximately 42.9 in./yr 

(109 cm/yr) has been estimated (Clawson, Start, 

and Ricks 1989).

2.1.2.5 Relative Humidity. Average 

relative humidity at CFA ranges from a monthly 

average minimum of 15% in August to a monthly 

average maximum of 89% in December and 

February. On a daily basis, humidity reaches a 

maximum just before sunrise, at the time of the 

lowest temperature, and a minimum late in the 

afternoon, near the time of the highest 

temperature. 

2.1.2.6 Tornadoes. Most of the tornado 

activity in the U.S. occurs east of the Rocky 

Mountains. In Idaho, tornadoes have been reported 

only in the spring and summer seasons (April 

through August). In the 42-year period of 1916 

through 1957, 19 tornadoes were reported in 

Idaho. With expanding population and better 

surveillance methods, the average number of 

tornadoes per year will probably continue to 

increase slowly in Idaho. Compared to areas in the 

Midwest, however, the frequency will remain very 

small. With very few tornadoes occurring in the 

state per year, the chances of any one location 

being struck are remote.

When taken in context with maximum 

atmospheric moisture content, surrounding 

geography, and other statistics, the national 

tornado statistics allow a realistic assessment of 

tornado risk and establish a value for the 

maximum credible tornado expected at the 

INEEL. For the years 1950 to the present, the 

NOAA record indicates there have been a total of 

five funnel clouds sighted within the boundaries of 

the INEEL. The calculated return period for a 

tornado on the INEEL with wind speeds exceeding 

120 mph (190 km/hr) is 1E+06 years (Coats and 

Murray 1985). 

2.1.2.7 Dust Devils. Although tornadoes are 

rare at the INEEL, the whirling winds of the less 

violent dust devils are common. These dust devils 

pick up dust and pebbles and can overturn, blow 

down, or carry off insecure objects. They usually 

occur on warm, sunny days with little or no wind. 

The dust cloud may be several hundred yards in 

diameter and extend several hundred feet into the 

air.

2.1.2.8 Hurricanes and Tropical Storms. 

Because of the moderating influence of the Pacific 

Ocean and the isolating influence of surrounding 

mountains, neither hurricanes nor tropical storms 

occur at the INEEL.

2.1.2.9 Precipitation Extremes. The 

highest precipitation extremes at the INEEL 

(regardless of location) are cited in Table 2-2 

(NOAA 1984). The greatest amounts recorded 

during 1-hr and 24-hr periods are listed monthly 

and annually. The high hourly amounts during 

May and June were the result of heavy 

thunderstorms passing over the rain gauge. The 

maximum for one hour was 1.15 in. (2.92 cm) at 

Test Area North (TAN). Precipitation amounts 

greater than 1 in./d (3 cm/d) have occurred during 

10 of the calendar months within the period of 

record. Some months have had multiple 

occurrences.

2.1.3 Ecology 

2.1.3.1 Flora. In 1975, the INEEL Site was 

dedicated as one of five DOE National 

Environmental Research Parks. It is an outdoor 

laboratory used to study ecological relationships 

and the effects of human activities on natural 

systems. In addition, it provides a unique setting 

for scientific investigation because the public has 

been excluded from much of the area for the past 

25 yr. Ecological data collected from the Idaho 

National Environmental Research Park provide a 

basis for analyzing environmental changes over 

time and assessing the effect of human influence 

on the environment. 

Research on the flora and fauna of the INEEL 

Site has largely been conducted by or in 

conjunction with the DOE Radiological and 

Environmental Sciences Laboratory (RESL). The 

physical aspects of the INEEL Site and its flora 

and fauna are typical of cold, high altitude, 

sagebrush ecosystems found in many parts of the 

western United States. The following discussion of 

the flora and fauna at the INEEL is from the 
2-15 

Table 2-2. Greatest precipitation measured at the INEEL (regardless of location).a

Month 

1-hrb 

(in.) 

24-hrc 

(in.) 

January 0.18 1.08 

February 0.18 0.96 

March 0.17 0.61 

April 0.24 1.51 

May 1.00 1.78 

June 1.15 1.73 

July 0.24 1.33 

August 0.45 1.44 

September 0.55 1.55 

October 0.34 1.12 

November 0.25 1.02 

December 0.23 1.18 

Maximum 1.15 1.78 

a. Source: NOAA 1984. 

b. From January 1950 through December 1964; hourly amounts were not available from 1965 through 1982. 

c. From January 1950 through December 1982.

Environmental Resource Document for the Idaho 

National Engineering Laboratory (Irving 1993). 

This report contains additional detailed 

information and references to specific ecological 

studies. The common and scientific names for the 

flora discussed here are presented in Table 2-3. 

For ease of reading, only the common names are 

used in this discussion. 

Extensive surveys of INEEL vegetation were 

carried out in 1952, 1958, and 1967 using 150 

permanent transects established and maintained 

for this purpose (Harniss and West 1973). Reports 

by McBride et al. (1978) and Jeppson and Holte 

(1978) also have described vegetation. 

The common vegetation type, found on 

approximately 80% of the INEEL Site, is a 

mixture of big sagebrush, green rabbitbrush, and 

perennial grasses. Most of the trees on the INEEL 

Site are scattered along the Big Lost River and the 

Twin Buttes area. Figure 2-10 depicts the 

distribution of vegetation at the INEEL. 

Vegetation in low-lying areas and along playa 

borders consists primarily of alkaline-tolerant 

species including shadscale saltbush, nuttal 

saltbush, and winterfat. Important associated 

grasses are bottlebrush squirreltail, giant wildrye, 

and Indian ricegrass. 

Prickly pear, painted milkvetch, and 

skeletonweed are common in sandy areas in the 

north. Willows, baltic rush, and povertyweed grow 

along the Big Lost River channel. 

Several studies have been conducted at the 

INEEL on the plant rooting depths, especially for 

the RWMC Subsurface Disposal Area (SDA). 

Vegetation studies of plant uptake of radionuclides 

at the INEEL have focused primarily on 

(a) determining if deep-rooted plants are a 

mechanism for waste pit intrusion and subsequent 
2-16 

Table 2-3. Flora at the INEEL Site. 

Common Name Latin Name 

Cactus Family—Cactaceae 

Coryphantha Coryphantha sp. 

Prickly pear cactus Opuntia polyacantha 

Goosefoot Family—Chenopodiaceae 

Shadscale saltbush Atriplex confertifolia 

Nuttall saltbush Atriplex nuttallii 

Winterfat Ceratoides lanata 

Summer cypress Kochia scoparia 

Povertyweed Monolepsis nuttalliana 

Russian thistle Salsola kali 

Composite or Aster Family—Compositae 

Big sagebrush Artemisia tridentata 

Threetip sagebrush Artemisia tripartite 

Hoary false-yarrow Chaenactis douglasii 

Green rabbitbrush Chrysothamnus viscidiflorus 

Skeleton weed Lygodesmia grandiflora 

Common dandelion Taraxacum officinale 

Gray horsebrush Tetradymia canescens 

Goatsbeard or yellow salsify Tragopogon dubius 

Mustard Family—Cruciferae 

Flixweed tansy mustard Descurainia Sophia 

Grass Family—Gramineae 

Crested wheatgrass Agropyron cristatum 

Bluebunch wheatgrass Agropyron spicatum 

Cheatgrass Bromus tectorum 

Giant wildrye Elymus cinereus 

Indian ricegrass Oryzopsis hymenoides 

Bottlebrush squirreltail Sitanion hystrix 

Rush Family—Juncaceae 

Baltic rush Juncus balticus 

Pea Family—Leguminosae 

Painted milkvetch Astragalus ceramicus Sheld. var. apus Barneby 

Thistle milkvetch Astragalus kentrophyta Gray var. kentrophyta

Woolly-pod milkvetch Astragalus purshii Dougl. var. ophiogenes Barneby 

Phlox Family—Polemoniaceae 

Large-flowered gymnosteris Gymnosteris nudicaulis Greene 

Longleaf phlox Phlox longifolia 

Buckwheat Family—Polemoniaceae 

Buckwheat Oxytheca dendroides,

a

 Nutt. 

Willow Family—Salicaceae 

Willows Salix sp. 

Parsley Family—Umbelliferae 

Desert parsley Lomatium sp. 

a. Source: Hitchcock and Cronquist 1974. 
2-17 

Figure 2-10. Approximate distribution of vegetation at the INEEL (INEEL 1999b). 
2-18

uptake of radionuclides, and (b) analyzing 

inventories of radionuclides in aerial portions of 

plants. Aerial portions of plants are important 

because they can potentially transport subsurface 

contaminants through dispersal of leaves, 

consumption by herbivores, and use by birds as 

nesting materials. 

One RWMC SDA study compared 

radionuclide uptake by crested wheatgrass 

(rooting depth 3 to 4.9 ft [0.9 to 1.5 m]) to that of 

Russian thistle (rooting depth 3 to 16 ft [0.9 to 

4.9 m]). The study showed higher radionuclide 

concentrations in the deeper-rooted species 

(Arthur 1982). Examples of other deep-rooting 

species are rabbitbrush and sagebrush. General 

examples of shallow-rooting plant types are 

grasses and annual forbs. 

Reynolds and Fraley (1989) found that the 

roots of big sagebrush extended to a depth of 

88.7 in. (225 cm), green rabbitbrush to a depth of 

74.9 in. (190. cm), and Great Basin wildrye had 

roots up to 78.8 in. (200. cm) deep at the SDA. 

Maximum lateral spread of the roots of both big 

sagebrush and Great Basin wildrye was 35.5 in. 

(90.2 cm) and occurred at a depth of 15.8 in. 

(40.1 cm). In addition, studies indicate root 

penetration of up to 5.2 ft (1.6 m) for sodar and 

crested wheatgrass at the INEEL (Markham 

1987). 

A survey of rare plants on the INEEL Site was 

initiated in 1980 (Cholewa and Henderson 1984). 

The survey identified the following plants: painted 

milkvetch and woolly-pod milkvetch, which were 

under federal review for endangered or threatened 

status; coryphantha, large-flowered gymnosteris, 

and oxytheca, on the Idaho State Watch List; and 

thistle milkvetch, which was previously unknown 

to occur in Idaho. Since then, painted and woollypod milkvetch have been removed from 

endangered or threatened candidate status 

(Mosely and Groves 1990). 

2.1.3.2 Fauna. The INEEL supports a 

variety of wildlife including small mammals, 

birds, reptiles, and a few large mammals. The 

common and scientific names for the fauna 

discussed here are presented in Table 2-4. For 

ease of reading, only the common names are used 

in this discussion.

The mammals include chipmunks, ground 

squirrels, several species of mice, kangaroo rats, 

cottontail rabbits, bats, jackrabbits, and coyotes. 

Commonly occurring game animals are sage 

grouse, mourning dove, elk, pronghorn, and mule 

deer. 

Limited data are available on the number of 

game animals seasonally inhabiting the INEEL 

and the harvest of these animals by hunters. 

Pronghorn inhabit the INEEL during the entire 

year; however, many are migratory and summer to 

the north of the INEEL. Pronghorn often bear 

young within the INEEL. 

Aquatic life on the INEEL is limited and 

depends mainly upon the flow of the Big Lost 

River. During several months of the year, and 

even during some entire years, the river does not 

flow. However, during spring runoff and periods 

of high rainfall, the diversion system (at the 

southern boundary of the INEEL) and the Big 

Lost River sinks (at the northern boundary of the 

INEEL) support water flow when water 

accumulates. This normally occur less than 2 or 3 

months in the spring. Fish species observed in the 

Big Lost River on the INEEL include rainbow 

trout, mountain whitefish, eastern brook trout, 

Dolly Varden char, Kokanee salmon, and 

shorthead sculpin (Overton et al. 1976). 

An investigation of amphibians and reptiles 

within the INEEL was conducted from May 

through September 1975 (Reynolds et al. 1986). 

The Great Basin spadefoot toad was the only 

amphibian recorded, typically associated with the 

Big Lost River, the Big Lost River sinks, and the 

spreading areas near the RWMC. The sagebrush 

lizard and the short-horned lizard are common; 

the sagebrush lizard is the most abundant reptile. 

The western skink and the leopard lizard also have 

been observed. Four species of snakes, including 

the Great Basin rattlesnake and Great Basin 

gopher snake, were recorded. The western 

terrestrial garter snake and the desert-striped 

whipsnake are present in lesser numbers and have 

more restricted distributions. 
2-19

Table 2-4. Fauna at the INEEL. 

Common Name Latin Name 

Fisha

Rainbow trout Salmo gairdneri 

Eastern brook trout Salvelinus fontinalis 

Dolly varden char Salvelinus malma 

Kokanee salmon Oncorhynchus nerka 

Mountain whitefish Prosooium williamsoni 

Shorthead sculpin Cottus confuses 

Reptiles and Amphibiansb

Leopard frog Rana pipiens 

Great Basin spadefoot toad Spea intermontanus 

Leopard lizard Gambelia wislizenii 

Sagebrush lizard Scel6porus graciosus 

Short-horned lizard Phrynosoma douglassi 

Western skink Eumeces skiltonianus 

Desert-striped whipsnake Masticophis taeniatus 

Great Basin gopher snake Pituophi-s melanoleucus 

Western Terrestrial garter snake Thamnophis elegans 

Great Basin rattlesnake Crotalus viridis 

Mammalsc

Family—Canidae

Coyote Canis latrans 

Family—Felidae

Bobcat Lynx rufus 

Family—Antilocapridae’

Pronghorn Antilocapra americans 

Family—Cervidae

Mule deer Odocoileus hemionus 

Elk Cervus canadensis 

Family—Vespertilionidae

Big-brown bat Eptesicus fuscus 

Townsend’s big-eared bat Plecotus townsendii 

Family—Leporidae

White-tailed jackrabbit Lepus townsendii 

Black-tailed jackrabbit L. californicus 

Nuttall’s cottontail Sylvilagus nuttallii 

Family—Sciuridae

Least chipmunk Tamias minimus

Townsend’s ground squirrel Spermophilus townsendii 

Family—Geomyidae

Northern pocket gopher Thomomys talpoides

Family—Heteromyidae

Ord’s kangaroo rat Dipodomys ordii 

Family—Cricetidae

Western harvest mouse Reithrodontomys megalotis 

Deer mouse Peromyscus maniculatus 
Table 2-4. (continued). 

2-20 

Common Name Latin Name 

Birdsd

Family—Accipitridae

Golden eagle Aquila chrysaetos 

Ferruginous hawk Buteo regalis 

Bald eagle Haliaeetus leucocephalus 

Family—Falconidae

Merlin Falco columbarius 

Prairie falcon Falco mexicanus 

Peregrine falcon Falco peregrinus 

Family—Phasianidae

Sage grouse Centrocercus urophasianus 

Family—Scolopacidae

Long-billed curlew Numenius americanus 

Family—Strigidae

Burrowing owl Athene cunicularia 

Family—Columbidae

Mourning dove Zenaida macroura 

Family—Corvidae

Blackbilled magpie Pica pica 

Family—Mimidae

Sage thrasher Oreoscoptes montanus 

Family—Turdidae

American robin Turdus migraturius 

Family—Tyrannidae

Say’s phoebe Sayornis saya 

Family—Alaudidae

Horned lark Eremophila alpestris 

Family—Emberizidae

Western meadowlark Sturnella neglecta 

Sage sparrow Amphispiza belli 

Brewer’s sparrow Spizella breweri 

a. Source: Simpson and Wallace 1978. 

b. Source: Nussbaum, Brodie, and Storm 1983. 

c. Source: Jones, Carter, and Genoways 1979. 

d. Source: American Ornithologist’s Union 1983.

A total of 740 insect species have been recorded 

at the INEEL; 226 of these species have not yet 

been identified beyond the family level. The 

majority of the abundant species belong to the 

orders Hymenoptera (wasps and ants) and Diptera 

(flies). About half of the abundant species are 

parasitic or predatory. 

Over 185 species of birds have been recorded 

on the INEEL, and about 60 of these species 

probably breed on the INEEL. Many of the bird 

species are relatively uncommon, and only a few 

species are very abundant. The most common 

species are the Brewer’s sparrow, sage thrasher, 

sage sparrow, horned lark, sage grouse, mourning 

dove, western meadowlark, blackbilled magpie, 

and robin (Reynolds et al. 1986). 

Species on the INEEL that merit special 

consideration because of their sensitivity to 

disturbance or their threatened status include the 

ferruginous hawk, merlin, long-billed curlew, 
2-21 

Townsend’s big-eared bat, common loon, white 

pelican, great egret, and trumpeter swan (Reynolds 

et al. 1986; Mosely and Groves 1990). The bald 

eagle and peregrine falcon are on the Federal 

Endangered Species List and occasionally visit the 

INEEL. 

Studies have been performed on burrowing 

characteristics of small mammals such as ground 

squirrels, deer mice, and voles (Arthur, Grant, and 

Markham 1983; Markham 1987; Reynolds and 

Laundre 1988). Results of the studies indicate 

burrows no deeper than 140 cm (55 in.) at the 

INEEL. 

2.1.4 Geology, Seismology, and 

Volcanology 

2.1.4.1 Regional and Site-Specific 

Geology. The bulk of the geology sections in this 

report are taken primarily from previous work 

conducted by Maheras et al. (1994) and Anderson, 

Kuntz, and Davis (1999). The INEEL is located on 

the west-central part of the ESRP, a northeasttrending structural basin about 200 mi (300 km) 

long and 50 to 70 mi (80 to 100 km) wide 

(Figure 2-11). The INEEL is underlain by a 

sequence of Tertiary and Quaternary volcanic 

rocks and sedimentary interbeds that are more than 

10,000 ft (3,000 m) thick (Whitehead 1992). The 

volcanic rocks consist mainly of basalt flows in 

the upper part of the sequence and rhyolitic ashflow tuffs in the lower part. Basalt and sediment 

generally range in age from about 200 thousand to 

4 million years before present (Anderson, 

Liszewski, and Cecil 1997), and underlie the plain 

at depths ranging from about 2,200 to 3,800 ft 

(670 to 1,200 m) below ground surface (bgs).

Hundreds of basalt flows, basalt-flow groups, 

and sedimentary interbeds underlie the INEEL. 

Basalt makes up about 85% of the volume of 

deposits in most areas. A basalt flow is a solidified 

body of rock formed by the surficial outpouring of 

molten lava from a vent or fissure (Bates and 

Jackson 1980). A basalt-flow group consists of 

one or more distinct basalt flows deposited during 

a single, brief eruptive event. All basalt flows of a 

group erupted from the same vent or several 

nearby vents represent the accumulation of one or 

more lava fields from the same magma and have 

similar geologic ages, paleomagnetic properties, 

potassium contents, and natural-gamma emissions 

(Anderson and Bartholomay 1995). The basalt 

flows consist mainly of medium- to dark-gray 

vesicular to dense olivine basalt. Individual flows 

generally range from 10 to 50 ft (3 to 20 m) thick, 

and are locally interbedded with scoria and thin 

layers of sediment. Sedimentary interbeds are as 

thick as 50 ft (20 m) and consist of well-sorted to 

poorly-sorted deposits of clay, silt, sand, and 

gravel. In places, the interbeds contain or consist 

mainly of scoria and basalt rubble. Sedimentary 

interbeds accumulated on the ancestral land 

surface for hundreds to hundreds of thousands of 

years during periods of volcanic quiescence and 

are thickest between basalt-flow groups. 

At least 178 basalt-flow groups and 103 

sedimentary interbeds underlie the INEEL above 

the effective base of the aquifer (Anderson, 

Ackerman, and Liszewski 1996; Anderson, 

Liszewski, and Cecil 1997). Basalt-flow groups 

and sedimentary interbeds are informally referred 

to as A through S5. Basalt-flow groups LM 

through L, and related sediment, range in age from 

about 200 to 800 thousand years old and make up 

the unsaturated zone and the uppermost areas of 

the INEEL. Most wells in the southern and eastern 

parts of the INEEL are completed in basalt-flow 

groups AB through I and the related sediments. 

Flow groups AB through I and related sediments 

range in age from about 200 to 640 thousand years 

old and make up a stratigraphic section 

characterized by horizontal to slightly inclined 

layers. Anderson, Liszewski, and Cecil (1997) 

estimated the geologic ages of basalts and 

sediments in the unsaturated zone and the Snake 

River Plain Aquifer to be from about 200 thousand 

to 1.8 million years; average accumulation rates 

are reflective of a subsidence rate of 164 ft 

(50.0 m)/100,000 yr. 

The nomenclature for the stratigraphy 

underlying the INTEC facility and the surrounding 

area is based on work presented by Anderson 

(1991). A north-south geologic cross section, 

illustrated in Figure 2-12, forms the framework for 

the conceptual model and numerical simulations 

discussed in Sections 3 and 4. 
2-22 

Figure 2-11. General geologic map of the INEEL (Kuntz et al. 1994). 
2-23 

Figure 2-12. North-south geological cross section. 
2-24 

The stratigraphy of the aquifer at and near 

INTEC is dominated by thick, massive basalt 

flows of flow group I and thin, overlying flows of 

flow groups B through H. The basalt flows, as 

interpreted, appear to be relatively uniform in 

thickness beneath INTEC. Significant changes in 

the flow thickness are often related to changes in 

the lithology of the flow or are caused by the flow 

margins in which the flow appears as a lobe of 

basalt. The lithologic changes that may cause a 

change in the flow thickness are the existence of 

pyroclastic deposits on or within a flow, or a flow 

being very vesicular, and thus, more susceptible to 

the effects of erosion. 

Based on the Anderson (1991) geologic cross 

section, the unsaturated zone and upper regional 

aquifer underlying INTEC are comprised of 19 

basalt-flow groups, 11 sedimentary interbeds, and 

surficial alluvium. The characteristics of these 

individual flow groups and their impact on the site 

hydrology are discussed in detail in Section 3. The 

sediments, as interpreted, appear to be primarily 

made up of sands and silts with some small clay 

lenses. The majority of the sediments are thin 1- to 

5-ft (0.3- to 2-m) layers of silt between the major 

basalt flows. Sediments were most likely 

deposited in eolian or fluvial type environments. 

Two major sediment sequences are shown on the 

cross sections: the upper sequence associated with 

the CD, thick D, and DE2 sands and silts and the 

lower sediments associated with the DE6, DE7, 

and DE8 stratigraphic units. 

The cross sections show a very thick sequence 

of sediments, particularly in the northern end of 

the north-south section, which are stratigraphically 

shown as the CD, D, and DE2 units. These 

sediments appear to be a thick sequence of sands 

over and underlain by silts and clays. The 

sediments associated with the DE6, DE7, and 

DE8stratigraphic units appear to be made up of 

gravels, silts, and clays. These sediments were 

most likely deposited in a fluvial environment and 

may indicate a braided stream deposit. This is the 

last major sediment deposit above the Snake River 

Plain Aquifer. 

Holocene surficial geology and archaeology 

suggest that fluvial and eolian deposition and 

tectonic subsidence in the INEEL area have been 

in approximate net balance for at least the past 

10,000 yr. A reversal of the long-term, regional 

pattern of ESRP subsidence, sedimentation, and 

volcanism into an erosional rather than a 

depositional regime would require major changes 

from the Holocene tectonic or climatic 

configuration of the ESRP. Worldwide geologic 

evidence indicates that the Quaternary epoch 

(approximately the past 2 million years) has been a 

time of major climatic fluctuations. During colder 

and wetter periods, glaciers occupied highelevation areas. Lowland areas such as the ESRP 

received thick, widespread loess blankets. 

Lowland areas also were periodically impacted by 

local catastrophes (such as the large, latePleistocene, glacial outburst flood(s) that traveled 

down the Big Lost River Valley), eroded upland 

surfaces on the ESRP and deposited sediment in 

the INTEC area. If the future ESRP climate were 

to become warmer and more arid, the probable 

consequences would be decreased vegetation and 

increased eolian transport of fine-grained 

sediment, mainly as longitudinal dunes of fine 

sand. 

Future climate fluctuations on the ESRP (to 

either colder/wetter or warmer/drier conditions) 

are not expected to erode the INTEC land surface. 

Quaternary geologic and Holocene archaeological 

data suggest the INEEL area will probably 

continue its long-term history of regional 

subsidence and net accumulation of sedimentary 

and volcanic materials, although sedimentation 

patterns on the ESRP will change in response to 

future climate fluctuations. 

INTEC soil cover erosion could occur as a 

consequence of faulting and uplift but this erosion 

would involve a major change in the Quaternary 

tectonic configuration of the ESRP. Therefore, this 

scenario is improbable within the next 10,000 yr, 

considering 

• The regional seismicity and tectonic history of 

the INEEL area 

• The absence of Quaternary tectonic faults on 

the ESRP in the vicinity of the INEEL 
2-25 

• The long response time for significant erosion 

to occur as a result of protracted faulting and 

uplift. 

In summary, the following impacts from 

volcanic and tectonic activity are relevant to 

INTEC radiological performance assessment: 

• During the past 4 million years, the ESRP and 

the INTEC area have undergone regional 

subsidence, basaltic volcanism, and fluvial and 

eolian sedimentation. Erosion has not been a 

significant process on the ESRP. 

• Surficial- and subsurface-geologic data 

indicate that the INTEC area has both 

subsided and accumulated basalt lava flows 

and sediments at an average rate of 0.01 in./yr 

(0.25 mm/yr). Significant uplift or erosion has 

not interrupted this long-term trend. 

• Lava inundation or magma intrusion 

associated with volcanism from the nearby 

Arco Volcanic Rift Zone is improbable 

considering the volcanic history of the area. 

Lava inundation or magma intrusion would 

not likely result in the release of radionuclides 

to the environment. 

Detailed discussions of the site geology and 

stratigraphy are presented in Anderson (1991) and 

Anderson, Liszewski, and Cecil (1997). 

2.1.4.2 Seismology. The seismically active 

Intermountain seismic belt surrounds the ESRP 

and Centennial Tectonic seismic belts. The 

Intermountain seismic belt is a zone of 

concentrated seismicity that extends from 

northwestern Montana through eastern Idaho and 

Utah into southern Nevada. The Centennial 

seismic belt, also a seismically active zone, 

extends from the Hebgen Lake, Montana, area 

westward into central Idaho.

The INEEL, USGS, Montana Bureau of Mines 

and Geology, United States Bureau of 

Reclamation, and the University of Utah 

Seismograph Stations have compiled earthquake 

data for 1884 to 1989 (shown in Figure 2-13). The 

distribution of epicenters indicates that the ESRP 

is devoid of earthquakes relative to the active areas 

surrounding it, with the possible exception of the 

1905 earthquake located at Shoshone, Idaho. 

Historical records suggest that the epicenter for the 

1905 earthquake is not located within the ESRP, 

but rather near the Idaho-Utah border. 

A large earthquake, in the vicinity of the 

INEEL but outside the ESRP, occurred in the 

Centennial seismic belt on October 28, 1983, with 

a surface-wave magnitude (M) of 7.3. The 

earthquake resulted from slippage along the Lost 

River fault—a northwest rupture along a normal 

fault with relative vertical movement downward to 

the west. The epicenter for this event was located 

in the Thousand Springs Valley near the western 

flank of Borah Peak, approximately 55 to 60 mi 

(89 to 100 km) from INEEL facilities. There was 

substantial damage to masonry structures in the 

local communities of Mackay and Challis near the 

epicentral area. Although the earthquake ground 

motions were felt at the INEEL Site, only minor 

nonnuclear building damage occurred in the form 

of hairline cracks and settlement (Gorman and 

Guenzler 1983). INTEC did not experience 

structural failures or waste spills as a result of the 

earthquake and waste storage facilities did not 

show evidence of permanent movement or 

resulting damage. Peak ground accelerations 

ranging from 0.022 to 0.078 g were recorded at 

several INEEL facility areas. The INEEL was 

located in Modified Mercalli Intensity Zone VI 

during the earthquake (Jackson 1985). 

The largest earthquake in the region occurred 

on August 17, 1959, at Hebgen Lake, Montana, 

located approximately 120 mi (190 km) northeast 

of the INEEL Site. The event had a surface-wave 

magnitude of 7.5; it was felt at the INEEL, but 

caused no damage there. 

The INEEL has maintained a seismic network 

for monitoring earthquake activity on and around 

the ESRP since December 1971. Currently, the 

seismic network consists of 24 seismic stations 

and 21 strong-motion accelerographs. The seismic 

stations continually record seismic data and their 

data are used to calculate the locations and 

magnitudes of microearthquakes (M < 3.5) that 

occur locally. When triggered, the strong-motion 

accelerographs record earthquake ground motions 

from local, moderate to large, earthquakes. 
2-26 

Figure 2-13. Seismicity map showing the ESRP in relation to the Intermountain seismic belt and the 

Centennial Tectonic belt (WCFS 1996). 

From 1972 to 1990, approximately 15 

microearthquakes occurred within the ESRP, 

indicating that infrequently-occurring, small 

magnitude earthquakes (M < 1.3) are characteristic 

of ESRP seismicity (Jackson et al. 1993; Pelton, 

Vincent, and Anderson 1990). These data are in 

agreement with the historical earthquake data 

compiled for the surrounding region (Figure 2-14). 

Recent modifications to the seismic network, such 

as placing sensors in 18- to 21-m (59- to 69-ft) 

boreholes, will lower the magnitude threshold of 

detecting microearthquakes within the ESRP. 

Because the seismically-active Intermountain 

and Centennial seismic belts surround the ESRP, 

and several Quaternary faults are located near the 

western boundary of the INEEL, seismic hazard 

assessments are being updated for all facility areas 

at the INEEL. These assessments are being 

performed to quantitatively estimate peak ground 

motions that INEEL facilities may experience 

from nearby large magnitude earthquakes. Most of 

the INEEL is located in Seismic Zone 2B and a 

small portion is located in Zone 3. The seismic 

design levels for INEEL facilities exceed those 

required for these classifications. 

Uplift and erosion of the INTEC area could 

result from faulting and uplift of the southern Lost 

River Fault if the fault encroached southward onto 

the ESRP to a position several kilometers west of 

INTEC. Assuming immediate initiation of this 

faulting and maximum-uplift rates from the most 

recently active fault segments of the nearby Basin 

and Range Province (1 to 2 m [3 to 7 ft]/1,000 yr), 
2-27 

Figure 2-14. Earthquake epicenters in the INEEL region, 1850–1995 (INEEL 1999b). 
2-28 

significant uplift and erosion of the INTEC area 

would require 10,000 to 100,000 yr. 

Additional information on seismic hazards for 

the region is presented in Environmental Resource 

Document for the Idaho National Engineering 

Laboratory (Irving 1993). 

2.1.4.3 Volcanology. Most of the INEEL is 

underlain by a 0- to 0.6-mi (0- to 1-km) thick 

sequence of late Tertiary and Quaternary basalt 

lava flows and interbedded sediments. Based on 

drill-hole information, regional mapping along the 

margins of the ESRP, and geophysical 

information, the basalt/sediment sequence is 

underlain by an older section (up to several 

kilometers thick) of late Tertiary rhyolitic volcanic 

rock. These two volcanic sequences are a 

consequence of the passage of the Yellowstone 

mantle plume (hotspot) through the INEEL area of 

the ESRP in late Tertiary time (Malde 1991). The 

Tertiary rhyolitic volcanic rocks were erupted at 

6.5 to 4.3 million annum (Ma) when the hotspot 

resided beneath the INEEL area. They consist 

mostly of ash-flow tuffs erupted during large, 

violent explosive episodes and large rhyolitic lava 

flows. They are analogous to the ash-flow tuffs 

and lava flows that erupted from calderas in the 

Yellowstone Plateau at 2.0 to 0.6 Ma.

These types of large-scale explosive eruptions 

can occur only directly over the mantle hotspot 

because large inputs of heat into the lower and 

middle crust are required to generate such large 

volumes of rhyolitic magma. Because the hotspot 

is now situated beneath Yellowstone National 

Park, recurrence of this type of volcanic activity in 

the INEEL area is nearly impossible. Residual heat 

in the upper mantle after passage of the hotspot 

has continued to produce basaltic magmas that 

have risen to the surface and erupted onto the 

subsiding ESRP. Basaltic eruptions in the INEEL 

area began at about 4 Ma, soon after passage of 

the hotspot, and have continued. The most recent 

activity occurred along the Great Rift about 

2,100 yr ago. 

Basalt vents on the ESRP include broad, 

nearly circular, low-relief shield volcanoes, small 

spatter cones, and spatter ramparts along eruptive 

fissures. Lava fields related to single vents range 

in surface area from 2 to 400 km2

 (0.7 to 154 mi2) 

and in volume from 0.05 to 7 km3

 (0.01 to 1.7 mi3) 

(Kuntz, Covington, and Schorr 1992). Volcanic 

vents are not randomly distributed on the ESRP; 

they are concentrated in northwest-trending linear 

zones known as volcanic rift zones (Figure 2-15). 

In addition, vents are concentrated in a northeasttrending zone, known as the Axial Volcanic Zone, 

along the central axis of the ESRP. The Axial 

Volcanic Zone is a constructional highland caused 

by more voluminous magma output along the axis 

of the ESRP. 

Based on radiometric age determinations of 

basalt lava flows, the Arco Volcanic Rift Zone 

north of the Big Southern Butte was active 

between 600 and 100 kilo annum (Ka) (Kuntz, 

Covington, and Schorr 1992). The Cerro Grande 

and North and South Robbers flows (10,500 to 

12,000 Ka) near the Big Southern Butte occur at 

the intersection of the Arco Volcanic Rift Zone 

and the Axial Volcanic Zone. Except for 

volcanism along the Great Rift, all of the 

Holocene volcanic fields of the ESRP occur along 

the Axial Volcanic Zone (Figure 2-15). 

Recurrence of volcanism in the ESRP has a greater 

likelihood of occurring along the Great Rift or the 

Axial Volcanic Zone. 

Additional information on the site volcanism 

is presented in “Quarternary Volcanism, 

Tectonics, and Sedimentation in the Idaho 

National Engineering Laboratory Area” (Hackett 

and Smith 1992). 

2.1.5 Hydrology 

2.1.5.1 Surface Water. The following 

sections discuss the surface water hydrology of the 

region. References for further information on the 

surface water features and resources of the INEEL 

site include Lamke (1969), Carrigan (1972), Tullis 

and Koslow (1983), and Bennett (1990).

Most of the INEEL and all of INTEC are in 

the Pioneer Basin. The Pioneer Basin is a closed 

topographic depression on the ESRP that receives 

intermittent runoff from the Big Lost River, Little 

Lost River, and Birch Creek drainage (see 

Figure 2-16). The Pioneer Basin is not crossed by 

any perennial streams because of the permeability 
2-29 

INEEL

Figure 2-15. Volcanic rift zones at the INEEL. 
2-30 

INEEL

Figure 2-16. Surface water features of the INEEL and surrounding vicinity. 
2-31 

of alluvium and underlying rock of the basin, 

which causes the water to infiltrate into the 

ground. In addition, much of the water from the 

tributary drainage basins is diverted for irrigation 

upstream of the INEEL. The largest stream, the 

Big Lost River, enters the INEEL near the 

southern end from the west. During exceptionally 

wet years, the Big Lost River flows in a large arc 

north to the foot of the Lemhi Mountain Range, 

where it ends in a series of playas (sinks). The 

only other naturally occurring stream on-Site is 

Birch Creek, which enters from the north. This 

stream is usually dry, except during heavy spring 

runoff when water may flow onto the INEEL. The 

Little Lost River approaches the INEEL from the 

northwest through Howe and ends in a playa just 

off the INEEL Site. 

The Big Lost River discharges an average of 

211,000 acre-ft/yr (260E+08 m3

/yr) below 

Mackay Dam, 30 mi (50 km) northwest of Arco. 

The largest recorded annual flow of the Big Lost 

River for the entire period of record occurred 

in 1984 and amounted to 476,000 acre-ft/yr 

(5.87E+08 m3

/yr), which was measured below 

Mackay Dam. The second largest annual flow 

occurred in 1965 and amounted to roughly threequarters of the 1984 record. 

Other than these intermittent streams, playas, 

and man-made percolation, infiltration, and 

evaporation ponds, there is little surface water at 

the INEEL Site. Surface water that reaches the 

INEEL is not used for consumptive purposes (e.g., 

irrigation, manufacturing, or drinking). In 

addition, there are no identified future uses of 

surface water that reach the INEEL. 

The northwest boundary of INTEC is closest 

to the Big Lost River channel, approximately 

200 ft (60 m) away. This is near the point where 

the channel intersects with Lincoln Boulevard on 

the INEEL. The Big Lost River is the principle 

natural surface water feature on the INEEL and 

the only stream with potential impacts to the TFF. 

The Big Lost River flows southeast from Mackay 

Dam, through the Big Lost River Basin past Arco, 

Idaho, and onto the ESRP. Stream flows are often 

depleted before reaching the INEEL by irrigation 

diversions and infiltration losses along the river. 

When flow in the Big Lost River does reach the 

INEEL, it is either diverted to the flood diversion 

facilities or flows northward across the INEEL in 

a shallow, gravel-filled channel to its terminus in 

the Big Lost River playas. In the playas, its flow is 

lost to evaporation and infiltration recharging the 

Snake River Plain Aquifer. See Table 2-5 for 

monthly discharge values of the Big Lost River at 

Lincoln Boulevard near INTEC. 

Major control on the Big Lost River upstream 

of the INTEC site includes the Mackay Dam and 

the INEEL flood diversion facility, each of which 

is discussed in greater detail below. 

2.1.5.1.1 Mackay Dam—Mackay 

Dam, located about 45 mi (72 km) upstream from 

the INEEL, impounds water from the Big Lost 

River for irrigation purposes downstream. Mackay 

Dam is a 1,430-ft (436-m) long, 79-ft (24-m) high 

earth-filled dam built for the Big Lost River 

Irrigation District. The dam was completed in 

1917 and has a storage capacity of 44,500 acre-ft 

(5.49E+07 m3

) and a surface area of 1,241 acres 

(502.2 ha) at a water-surface elevation of 

6,066.5 ft (1,849.1 m) (Table 2-6). An ungated 

overflow spillway with a weir length of 75 ft 

(23 m) at an elevation of 6,066.5 ft (1,849.1 m) 

above mean sea level (amsl) is located near the 

west abutment of the dam. The spillway is 

designed for a discharge of 3,250 ft3

/s (92.0 m3/s) 

with 4 ft (1 m) of freeboard on the dam. The outlet 

works also are located near the west abutment and 

extend through the embankment and under the 

spillway to form an outlet channel. The outlet 

works consist of five motor-operated slide gates 

measuring 4 × 8 ft, mounted in an upstream 

control tower. The arched-roof outlet tunnel 

measures 10 × 10 ft, and reaches 500 ft 

downstream into a 10-ft diameter steel pipe, 

which extends to the outlet. At the outlet, the pipe 

branches into six 4-ft diameter pipes, emptying 

into a stilling basin at the toe of the dam. The total 

discharge capacity of Mackay Dam is less than 

10,000 ft3

/s (300 m3/s). Water from the Big Lost 

River is impounded for the irrigation of about 

57,500 acres (23,300 ha) of land downstream 

from the reservoir and for recreational 

opportunities. Another 10,200 acres (4,130 ha) of 

land upstream from the reservoir also is irrigated 

with Big Lost River water (INEEL 1999b).
2-32 

Table 2-5. Monthly discharge of the Big Lost River at Lincoln Boulevard near INTEC (ft3).a

Year Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov 

1965 0 0 2,380 10,300 15,400 29,600 31,100 16,900 10,900 0 0 1966 0 0 0 3,660 981 0 0 0 0 0 0 1967 0 0 0 0 2,030 20,180 18,376 4,400 9,050 8,740 0 1968 0 0 2,280 3,390 16 524 0 1,053 1,130 3,290 4,500 1969 0 0 0 3,960 33,000 33,500 21,800 4,780 9,840 6,710 3,290 1970 0 0 501 1,650 793 13,800 17,700 1,510 6,080 5,280 4,750 1971 0 0 0 10,600 12,300 17,200 20,800 7,760 13,400 14,400 13,100 1972 0 0 1,540 4,920 504 1,710 861 84 2,990 3,520 3,099 1973 0 0 0 2,830 405 0 0 0 0 0 0 1974 0 0 3,240 5,520 6,940 16,200 9,390 1,170 1,160 3,760 4,200 1975 0 0 0 3,180 12,000 12,100 18,700 3,560 6,520 8,210 7,990 1976 0 0 333 1,450 1,660 1,120 0 0 300 620 1,100 1977 0 0 0 0 0 0 0 0 0 0 0 1978 0 0 0 0 0 0 0 0 0 0 0 1979 0 0 0 0 0 0 0 0 0 0 0 1980 0 0 0 0 0 0 1,140 0 0 0 0 1981 0 0 0 1,300 5,092 7,560 0 0 0 0 0 1982 0 0 0 5,930 17,200 13,400 15,100 4,820 8,190 10,500 5,740 1983 600 600 900 12,800 15,800 18,900 18,200 9,780 7,320 6,200 5,660 1984 1,200 1,200 1,200 2,200 2,230 4,550 3,950 5,790 5,140 5,980 8,710 1985 3 0 0 7,170 6,430 0 0 0 9,950 10,707 1,275 1986 0 96 537 8,370 14,825 20,315 2,900 1,016 14,753 8,220 1,190 1987 0 0 531 491 0 0 0 0 0 0 0 1988 0 0 0 0 0 0 0 0 0 0 0 1989 0 6 0 0 0 0 0 0 0 0 0 1990 0 0 0 0 0 0 0 0 0 0 0 1991 0 0 0 0 0 0 0 0 0 0 0 1992 0 0 0 0 0 0 0 0 0 0 0 1993 0 0 0 0 0 5,116 0 0 0 0 0 1994 0 0 0 0 0 0 0 0 0 0 0 

a. Sources: INEEL 1999b. 
2-33

Table 2-6. Characteristics of Mackay Dam and the INEEL Diversion Dam and reservoir.a

Characteristic Mackay Dam INEEL Diversion Dam 

Dam crest elevation (ft amsl) 6,076.0 5,064.7 

Dam crest length (ft) 1,430 500 

Dam height (ft) 79 22 

Spillway Ungated overflow crest, 75 ft long None 

Spillway crest elevation (ft amsl) 6,066.5 NA 

Gate centerline elevation (ft amsl) 6,036.6 (upper) 

6,007.8 (lower) 

5,045.6 

Dam base elevation (ft amsl) 5,997.0 5,042.6 

Spillway maximum capacity (ft3

/s) 6,588 NA 

Gate maximum capacity (ft3

/s) 2,960 1,121 

Reservoir capacityb

 (acre-ft) 55,091 at 6,076.0 ft 

44,500 at 6,066.5 ft 

8,750 at 6,030.0 ft 

500 at 6,010.0 ft 

58,000 at 5,050.0 ft 

18,200 at 5,040.0 ft 

a. Sources: Koslow and Van Haaften 1986; INEEL 1999b. 

b. It has been estimated that Mackay Reservoir has lost 22% of its mid- and late-season irrigation capacity because of 

sedimentation of the reservoir. Reservoir capacity given for the INEEL Diversion Dam is for the spreading areas; no water is 

held immediately behind the Diversion Dam. 

2.1.5.1.2 INEEL Flood Diversion 

Facility—The INEEL Flood-Diversion Facility 

(FDF) includes a diversion dam, dikes, and 

spreading areas located about 10 mi (20 km) 

upstream from INTEC. The FDF was constructed 

in 1958 and enlarged in 1984 to reduce the threat 

of flood on the INEEL from the Big Lost River. 

The FDF controls or divides the flow in the Big 

Lost River between the spreading areas to the 

south and the playas to the north where the water 

can be temporarily stored until it infiltrates into the 

ground. This prevents floodwaters from impacting 

INTEC and other INEEL facilities. The FDF has 

an elevation between 5,030 and 5,064.7 ft (1,530 

and 1,543.7 m) amsl. INTEC lies at about 4,917 ft 

amsl (1,499 m) amsl, and the playas, located about 

18 mi amsl (29 km) downstream from INTEC, lie 

between about 4,780 and 4,790 ft amsl (1,457 and 

1,460 m amsl).

The FDF’s diversion dam consists of a small 

earthen diversion dam and headgate that diverts 

water from the main channel, through a connecting 

channel, and into a series of four natural 

depressions, called spreading areas. Flow in the 

diversion channel is uncontrolled at discharges 

that exceed the capacity of the culverts. The 

diversion channel is capable of carrying 7,200 ft3

/s 

(2.0E+02 m3

/s) from the Big Lost River channel 

into the spreading areas. Two low swales located 

southwest of the main channel will carry an 

additional 2,100 ft3

/s (59 m3/s) for a combined 

diversion capacity of 9,300 ft3

/s (260 m3/s). The 

capacity of the spreading areas is about 

58,000 acre-ft (7.2E+07 m3

) at an elevation of 

5,050 ft (1,540 m) amsl. Runoff from the Big Lost 

River has never been sufficient to exceed the 

capacity of the spreading areas and overflow the 

weir. Gates placed on two 6-ft diameter corrugated 

steel culverts control flow downstream onto the 

INEEL. At full capacity, the culverts are capable 
2-34

of handling up to 900 ft3/s (30 m3/s) of flow 

through the diversion dam. As stated previously, 

there are no users of the surface water that reaches 

the INEEL (INEEL 1999b). 

2.1.5.2 Floods. A study of recorded 

discharge data suggests a history of lowmagnitude floods. The data are from several 

USGS stream-flow stations along the Big Lost 

River upstream of the INEEL. Flooding in the 

basin is associated with peak flows during the 

snowmelt season and occasional flooding caused 

by ice jams in the stream channel. Big Lost River 

flows seem to be attenuated because of the gravel, 

deep alluvium, and permeable basalt found in the 

channel bed. These streamflow losses, combined 

with controlled streamflow, diversion canals, and 

irrigation use, significantly impact the natural 

flood peaks. Downstream on the INEEL, the local 

semi-arid climate, relief, and geology combine to 

regulate local runoff. Local flooding in the past 

has been associated with unseasonably warm 

temperatures and rain on frozen ground as the 

following local flood history describes (INEEL 

1999b).

2.1.5.2.1 Flooding in 1965—A 

record snowpack occurred in the Big Lost River 

Basin in the winter of 1964–65. The maximum 

runoff occurred in late June. The Mackay 

Reservoir was full and most of the runoff was 

passed down to the basin and through the FDF on 

the INEEL. During the flood peak, June 29, 1965, 

approximately 1,800 ft3

/s (51 m3/s) was diverted to 

the spreading areas from a peak flow of 2,215 ft3

/s 

(62.72 m3

/s). The Big Lost River overflowed its 

banks above Arco through most of June. On the 

INEEL, the flood was controlled by the FDF and 

by the storage and infiltration in the river 

channels, playas, and sinks. The water did not 

reach the end of the Big Lost River channel at the 

Birch Creek playa during this flood. This flood is 

significant because it exhibited the largest crest 

and largest water volume to be discharged onto the 

INEEL in 65 years of record, yet caused no 

damage to INEEL facilities.

2.1.5.2.2 Flooding in 1984—High 

stream flows in the Big Lost River and a severe 

cold spell during the winter of 1983–84 caused ice 

jams that imposed a danger of localized flooding. 

Ice buildup in the spreading areas resulted in 

waters backing up in the diversion channel and 

ultimately threatening to overtop a dike. The high 

stream flows in the Big Lost River in 1983 and 

1984 were largely the result of the Borah Peak 

earthquake of October 28, 1983. The earthquake 

created new springs upstream of Mackay 

Reservoir, which significantly increased the 

inflows to the reservoir. Outflows from the 

reservoir also were increased to reduce the storage 

behind the dam. In response to this flood threat, 

the diversion area was upgraded to provide 

additional flood control, increasing the diversion 

channel flow capacity of 2,500 ft3

/s (71 m3/s) to 

over 9,000 ft3

/s (300 m3/s). Downstream, INEEL 

facilities were not threatened or damaged by the 

accumulation of ice in the diversion channel.

Generally during the winter months there is no 

flow in the Big Lost River downstream on the 

INEEL; however, if there is flow, nearly all the 

flow is diverted to the FDF to avoid the 

accumulation of ice in the main channel, reducing 

the possibility of flooding downstream. In a 

review of the historical information, no flooding or 

inundation from storms or runoff has caused 

flooding of the INTEC site. 

2.1.5.3 Potential Flooding 

Considerations. As discussed in the following 

sections, the bounding flood scenario for the 

INEEL TFF is the overtopping failure of the 

Mackay Dam from a general storm probable 

maximum precipitation (PMP) (INEEL 1999b). A 

detailed discussion of this general storm PMP may 

be found in the following sections. The INTEC 

TFF site elevation is near or at the extent of 

floodwaters 4,917 ft (1,498 m) amsl predicted for 

this hypothetical dam failure scenario.

2.1.5.3.1 Probable Maximum 

Flood on Streams and Rivers—The probable 

maximum flood (PMF) represents the hypothetical 

flood that is considered to be the most severe flood 

event reasonably possible, based on the hydrometeorological application of maximum 

precipitation and other hydrologic factors. Either 

an unusually severe storm or some catastrophic 

event, such as a dam failure, may cause the PMF. 

For conservatism in safety and design, the 

bounding scenario used for INEEL facilities is a 
2-35

PMF-induced overtopping failure of the Mackay 

Dam caused by an extreme precipitation event, the 

general storm PMP. Figure 2-17 represents the 

PMF hydrograph, and Figure 2-18 is the 

inundation map for the PMF-induced failure of the 

Mackay Dam. Table 2-7 lists information on the 

peak water surface elevation, peak flow, water 

velocity, and time of arrival at several downstream 

locations for this dam failure scenario. The 

following discussion was taken from the TMI-2 

Safety Analysis Report (INEEL 1999b), which is 

applicable to the assessment of the INTEC TFF 

maximum flood. 

2.1.5.3.2 Probable Maximum 

Precipitation—The general storm PMP for the 

drainage basin above Mackay Dam resulted from a 

48-hr general storm in June, preceded three days 

before by an antecedent storm with a magnitude of 

40% of the 48-hr storm. This scenario provides for 

no flow losses to the ground in order to be 

conservative and represent situations in which the 

ground may be frozen or fully saturated. The peak 

flow for the PMF is 82,100 ft3

/s, occurring 154 hr 

after the beginning of the storm. The PMF 

estimate falls within the 50,000–200,000 ft3

/s 

Myers envelope curve used by the U.S. Army 

Corps of Engineers (INEEL 1999b). The PMF 

peak flow is almost 20 times higher than the 

highest flow of 4,420 ft3

/s recorded at Howell 

Ranch, a USGS station located approximately 

17 mi (27 km) northwest of the dam. The PMF is 

based on the maximum potential for critical hydrometeorological conditions to occur, not on 

probabilities or historical flood frequencies.

2.1.5.3.3 Precipitation Losses—

The Big Lost River leaves the mountains at Arco. 

Below this point, the topography and 

drainage characteristics change along the river. 

The area is a low, flat plain with basalt bedrock. 

The drainage from most of the area in Pioneer 

Basin is integrated with the Big Lost River. 

Locally, some depressions in the basalt receive 

intermittent runoff. There is seldom enough 

precipitation in this area to exceed the infiltration 

capacity of the soil to create intermittent streams 

to the Big Lost River.

2.1.5.3.4 Runoff Model—The 

combined Big Lost River Basin and Pioneer Basin 

range in elevation from 4,784 ft (1,458 m) to over 

12,600 ft (3,840 m) amsl. Thus, this area has over 

7,546 ft (2,300. m) of relief, resulting in large 

differences in temperature and climate at any 

given time. The low land in the Pioneer Basin is 

subjected to periods of warm wind, rain, and 

snowmelt during the winter months. These 

conditions cause runoff and minor flooding in the 

lower basins during regional storms and 

Figure 2-17. Probable maximum flood hydrograph (INEEL 1999b). 

15 

15 

INEEL

LINCOLN

MINIDOKA

POWER

BINGHAM

BLAINE

BUTTE

CLARK

FREMONT

JEFFERSON

MADISON

BONNEVILLE

CARIBOU

BANNOCK

Camas National

Wildlife Refuge

Mud Lake Wildlife

Management Area

Mud Lake

Lost River Range

Lemhi Range

Bitterroot Range

Craters of the Moon 

Wilderness Area

INEEL 

Diversion Dam Atomic City

Rigby

Idaho Falls

Fort Hall 

Indian 

Reservation

Pocatello

American Falls

Reservoir

Shelley

Howe

American Falls

Snake River

Big Lost River

Arco

Blackfoot

INEEL

Vicintiy Map

Miles

Kilometers

0

0

30

40

A B

C D Spreading Areas

Box Canyon

Mackay 

Reservoir

Little Lost River

Big Lost River 

Sinks
2-36

Figure 2-18. Probable maximum flood INEEL inundation map (INEEL 1999b). 

INEEL

LINCOLN

MINIDOKA

POWER

BINGHAM

BLAINE

BUTTE

CLARK

FREMONT

JEFFERSON

MADISON

BONNEVILLE

CARIBOU

BANNOCK

Camas National

Wildlife Refuge

Mud Lake Wildlife

Management Area

Mud Lake

Lost River Range

Lemhi Range

Bitterroot Range

Craters of the Moon 

Wilderness Area

INEEL 

Diversion Dam Atomic City

Rigby

Idaho Falls

Fort Hall 

Indian 

Reservation

Pocatello

American Falls

Reservoir

Shelley

Howe

American Falls

Snake River

Big Lost River

Arco

Blackfoot

INEEL

Vicintiy Map

Miles

Kilometers

0

0

30

40

A B

C D Spreading Areas

Box Canyon

Mackay 

Reservoir

Little Lost River

Big Lost River 

Sinks

Probable maximum flood area 

(Koslow and Van Haaften 1986) 
2-37

Table 2-7. Results of probable maximum flood induced overtopping failure of Mackay Dam. a,b,c

Location 

(approximate elevation, ft amsl) 

Streambed 

elevation 

(ft amsl) 

Peak water 

surface 

elevation 

(ft amsl) 

Peak flood 

flow 

(ft3/s) 

Peak water 

velocity 

(ft/s) 

Time of 

wave 

arrival 

(hr) 

Mackay Dam (6,076) 5,997 6,078 306,700 8.5 0.0 

Arco (5,310–5,410) 5,309 5,319 147,720 5.6 6.7 

INEEL Diversion (5,065) 5,043 5,073 71,850 1.0 10.0 

CFA (4,928–4,940) 4,935 4,942 67,830 3.4 12.8 

TRA (4,920–4,925) 4,919 4,924 67,170 2.8 13.2 

INTEC (4,914–4,930) 4,911 4,917 66,830 2.7 13.5 

NRF (4,845–4,850) 4,846 4,851 61,620 1.9 16.4 

TAN (4,780–4,795) 4,778 4,786 34,810 1.1 34.5 

a. Sources: Koslow and VanHaaften 1986; INEEL 1999b. 

b. Total flow to INEEL diversion spreading areas = 27,460 acre-ft. 

c. Total Mackay Reservoir release = 142,330 acre-ft. 

substantially increase the snowpack in the uplands. 

The largest documented runoff periods in the

lower parts of the basins have occurred in January, 

February, or March; the maximum runoff from the 

highlands is usually in May or June. Generally, 

frost leaves the ground in the Pioneer Basin and 

the valley floors of the mountain basins in March 

or April; the permeable soils and gravels can then 

accept surface water by infiltration before the bulk 

of the snow pack starts to melt. Most surface water 

reaching the Pioneer Basin from the tributary 

drainage basins eventually infiltrates beneath the 

soil and rock to the groundwater reservoir. The 

remainder is lost through evaporation.

2.1.5.3.5 Probable Maximum 

Flood Flow—The spillway of Mackay Dam is 

not adequate to pass the PMP safely, therefore 

overtopping and subsequent breaching of the dam 

because of this PMP storm were analyzed. During 

this overtopping failure, the inflow is sufficient to 

raise the water surface above 6,077 ft 

(1,852 m) amsl, 1 ft (0.3 m) above the crest of the 

dam. A trapezoidal breach was assumed to 

develop over a 1-hr period and extend to the base 

of the dam. The computer code DAMBRK 

(National Weather Service 1991), was used in the 

flood-routing analysis.

The peak flow resulting from the PMPinduced overtopping failure is 306,700 ft3

/s in the 

reach immediately downstream of the Mackay 

Dam. This peak flow attenuates to 71,850 ft3

/s at 

the INEEL Diversion Dam and 66,830 ft3

/s at 

INTEC. The flood wave reaches the INEEL 

Diversion Dam in 10 hours. Water velocities are 

approximately 1 to 3 ft/s downstream on the 

INEEL. The computer program DAMBRK 

identified the water levels at specified locations 

for the PMF-induced overtopping failure. The 

wind activity at the INEEL, coincident with the 

largest projected flood crest, could not produce 

waves that would exceed 0.5 ft (0.2 m) primarily 

because of the shallow depth of water surrounding 

most INTEC buildings (LITCO 1994). Thus, the 

static and dynamic effects of wave activity would 

be negligible. 

2.1.5.3.6 Potential Dam Failures—

Mackay Dam is classified as a “high hazard” dam 

by the State of Idaho with reference to the U.S. 

Army Corps of Engineers guidelines for safety 

inspection of dams (Army Corps of Engineers 

1995). This high hazard classification is based on 

the concentration of people and property 

downstream, the size of the dam, and its storage 

capacity, not on any aspect of the dam’s current 

condition or operation.
2-38

Mackay Dam is located in a region of 

historical seismicity as evidenced by the 1983 

Borah Peak earthquake. The performance of the 

dam during this earthquake demonstrated the 

stability of the embankment during moderate 

ground motion. However, Mackay Dam was built 

without any seismic design criteria; therefore, a 

seismically-induced dam failure has been analyzed 

to determine potential impacts at the INEEL. This 

analysis assumed a postulated seismic failure of 

Mackay Dam during an inflow to the reservoir 

equal to the 25-yr recurrence interval flood (peak 

flow 4,030 ft3

/s). Because a seismic event may 

potentially disrupt a significant part of the dam’s 

structure, the breech was assumed trapezoidal, 

extending to the bottom of the structure at 5,997 ft 

(1,828 m) amsl, and developing over a 1-hr period. 

The peak flow from the seismic dam failure in the 

reach immediately downstream of the dam is 

107,480 ft3

/s. This peak flow attenuates to 

45,410 ft3

/s at the INEEL Diversion Dam and to 

39,080 ft3

/s at INTEC. The leading edge of the 

wave reaches the INEEL Diversion Dam in about 

12 hours. Average water velocities on the INEEL 

are 1 to 3 ft/s. 

Other dam failure permutations examined 

include two hydraulic (piping) failures concurrent 

with 100-yr and 500-yr inflow floods to the 

reservoir. The INEEL Diversion Dam would be 

overtopped by the floodwaters released from the 

failure of Mackay Dam. This overtopping of the 

INEEL Diversion Dam would contribute to the 

flooding downstream on the INEEL. The 

DAMBRK analysis assumes that the INEEL 

Diversion Dam begins to fail when floodwaters 

reach 5,065 ft (1,544 m) amsl, an overtopping 

depth of 0.3 ft (9 cm). Because of the small size of 

this dam, the breach is assumed to be fully 

developed after 0.1 hr, an essentially instantaneous 

failure. 

The flood from dam failure would initially 

travel down a valley between basalt flows. The 

initial velocity would be high near the failure, but 

the average velocity would decrease to 

approximately 1 ft/s near the FDF. Water entering 

the FDF from this flood is much less than the 

actual capacity of the spreading area (Koslow and 

Van Haaften 1986). Water that bypasses the FDF 

would continue to spread out across the floodplain 

and have a peak water velocity of 2.7 ft/s at 

INTEC. This would result in flood water within 

the INTEC controlled area up to 4,917 ft 

(1,499 m). The existing ground elevation of the 

INTEC TFF site varies between 4,914 to 4,917 ft 

(1,498 to 1,499 m) amsl, with the majority of the 

tanks located in an area with an elevation of 

4,916 ft (1,498 m).a

An additional study of the MacKay Dam 

failure was conducted by Van Haaften, Koslow, 

and Naretto (1984) for the flooding analysis of the 

New Production Reactor site. A PMF event above 

Mackay Dam was assumed to result in a 1.1 m 

(3.7 ft) overtopping of the dam, with a 0.5 hr dam 

failure, 42.6 m (140 ft) breach in the dam, and no 

flow losses downstream. This flooding analysis 

scenario resulted in flooding of the INTEC facility 

with 1 to 2 m (3 to 7 ft) of water covering the site 

and a peak flow of 3292.5 m3

/s (116,272 ft3/s). 

The flood hydrograph for the base-case discharge 

shows a peak flow that lasts only 7.5 hrs 

(Figure 2-19). Since the INTEC facility is located 

at the boundaries of the flood, the area would be 

inundated for a lesser period of time. 

Figure 2-19. Flood hydrograph at INTEC. 

 

a. Based on E.K. Roemer, Portage Environmental, personal 

communication with Randy Lee, a senior Geographical 

Information System analyst at the INEEL, 2000, and GPS 

mapping of INTEC. 
2-39

The impact of the worst possible flooding 

condition on the INTEC facility is expected to be 

minimal. The elevation of the INTEC facility is 

near the highest elevation that floodwaters could 

potentially reach. Since the facility is near the 

edge of the floodwaters, surface water flow 

velocities would have minor erosional effects. One 

to two meters of water could cover the facility, but 

this would occur only for a short duration. A small 

wetting front infiltrating into the unsaturated zone 

would occur. However, based on infiltration rates 

measured by Dunnivant et al. (1998) and the short 

duration that ponded water could occur at the 

facility, the wetting front would only advance 16 

to 33 ft (4.9 to 10. m) into the underlying alluvial 

soils. 

2.1.5.3.7 Perched Water. There are 

several perched water zones underlying the 

INTEC facility, which can be divided into an 

upper and lower perched water zone. The upper 

basalt perched water zone was initially discovered 

in the late 1950s. Perched water was encountered 

in Wells USGS-50 and USGS-52 at 126 and 174 ft 

(38.4 and 53.0 m) bgs, respectively. The 

occurrence of this perched water was attributed to 

operational practices based on the presence of 

radioactive and chemical contaminants. Since 

then, numerous monitoring wells have been 

installed in the upper perched water zone to 

identify the source of recharge, delineate the 

perched water bodies, and determine the nature 

and extent of contamination.

A lower perched water zone also was 

identified in the basalt at depths between 340 and 

400 ft (1.0E + 02 and 100 m) bgs (Robertson, 

Schoen, and Barraclough 1974). This water was 

first discovered in 1956 while drilling Well 

USGS-40; perched water was encountered at a 

depth of 348 ft (106 m). An analysis of this 

perched water detected abnormally high total 

dissolved solids (303 mg/L), sodium (25 mg/L), 

and chloride (81 mg/L), indicating the water is of 

waste origin (Olmsted 1962). According to 

Robertson, Schoen, and Barraclough (1974), this 

was a reasonable level for the perched 

groundwater because of the presence of a clay bed 

“aquitard” at 370 ft (110 m) bgs. In the late 1950s, 

only wells drilled in the northern INTEC area 

encountered the lower perched groundwater zone. 

Since 1984, a lower perched groundwater zone 

also has formed in the southern INTEC area 

because of the disposal of process waste through 

the percolation ponds. The location of this lower 

perched water zone is indicated by Well MW-17 

and borehole neutron logs from Well USGS-51. 

Stratigraphy controls the hydrogeologic 

characteristics of the subsurface at INTEC, 

particularly in the formation and movement of 

perched groundwater. The formation of perched 

groundwater can be attributed to lithologic 

features contributing to contrasts in the vertical 

hydraulic conductivity of basalt layers and 

sedimentary interbeds in the unsaturated zone. 

Cecil et al. (1991) attributed four lithologic 

features to the formation of perched groundwater 

at INTEC. Perched groundwater can form where 

(a) a sedimentary interbed with a reduced vertical 

hydraulic conductivity underlies a more 

conductive basalt layer, (b) altered baked zones 

between two basalt flows reduce the hydraulic 

conductivity, (c) dense unfractured basalt having 

low vertical hydraulic conductivity is present, or 

(d) sedimentary and chemical filling of fractures 

near the upper contact of a basalt flow reduces the 

vertical hydraulic conductivity. 

Water movement in the basalt units located in 

the unsaturated zone is poorly understood. The 

presence of perched water at the surficial 

sediment-basalt contact indicates how water 

moves into the underlying fractured basalts. Water 

in a partially saturated medium is held under 

negative pressure or tension. This capillary tension 

prevents water from moving from a region with 

small pore spaces (such as the sediment interbeds) 

into a region of larger pore spaces (such as the 

fractured basalts) until the tension is nearly 

reduced to zero pressure. In other words, the 

overlying sediment interbeds reach saturation. As 

a result, water movement into the fractured basalts 

is likely dominated by fracture density (Maheras et 

al. 1994). Fracture spacing measurements 

documented by Knutson et al. (1992) indicate that 

near perimeter zones of basalt flows, the spacing 

between fractures decreases to 0.25 m (0.82 ft). 

A large-scale field infiltration test conducted 

by Dunnivant et al. (1998) showed that infiltrating 

water moved downward through isolated fractures 
2-40

and did not spread horizontally until reaching an 

interbed located at a depth of 180 ft (55 m). As 

water reached the interbed, a perched water body 

developed that was characterized by direct current 

resistivity measurements and confirmed by the 

recovery of water from monitoring wells. Water 

moved outward from the area coinciding with 

geological abnormalities and with topographic 

lows on the interbed surfaces. Depending on the 

amount of water and the horizontal extent of the 

interbed, water would continue to move laterally 

until a discontinuity in the interbed was reached, at 

which point the downward migration of water 

would continue. 

In modeling studies conducted by Maheras et 

al. (1994) for the RWMC area, water movement 

through the fractured basalts is assumed to be 

rapid, on the order of 98 ft/mo (30 m/mo). 

Conceptually, the model for the RWMC area 

excluded the fractured basalts and only included 

the sediment interbeds stacked together so that 

they were in direct contact with each other. The 

fractured basalts were assumed to have negligible 

water travel times and were not included in the 

simulations. 

Modeling conducted for the WAG 3 RI/FS

assumed that the basalts could be treated as an 

anisotropic “single porosity” media, where the 

matrix is neglected and only the fractures are 

considered for the transport of water (Rodriguez et 

al. 1997). For this modeling approach, a horizontal 

and vertical fracture permeability of 

90,000 mDarcy and 300 mDarcy, respectively, and 

an effective porosity of 5% were applied. 

Figure 2-20 illustrates the unsaturated moisture 

characteristic curve and unsaturated hydraulic 

conductivity as a function of moisture content 

based on data presented in Rodriguez et al. (1997). 

Figure 2-20. Basalt moisture curve. 
2-41

The unsaturated moisture curve assumes the 

rapid dewatering of the fractures in response to 

pressure decreases. The relative permeability 

decreases linearly in response to decreases in 

moisture content. This linear permeability 

response is consistent with a similar approach 

presented in the Handbook of Vadose Zone 

Characterization and Monitoring (Wilson, 

Everett, and Cullen 1995), where the effective 

conductivity for fractured media is a function of 

the fracture density. Because of the fractured and 

the porous media interaction, this type of 

formulation assumes instantaneous equilibrium of 

total head between the two media and essential 

features of the flow dynamics are overlooked. 

Such phenomena as capillary barriers (in which an 

unsaturated fracture impedes fluid flow) and rapid 

transient responses are not inherent in this 

approach (Wilson, Everett, and Cullen 1995). 

Consequently, the use of porous media unsaturated 

curves to represent flow in the fractured basalt is a 

rough approximation of unsaturated fluid flow 

characteristics. 

It is difficult to quantify this approach as 

conservative or non-conservative since both 

capillary barriers and rapid transient responses are 

not incorporated into the analysis. The net result is 

uncertainty in predicting unsaturated flow and 

contaminant transport in the fractured basalts that 

is difficult to quantify. 

Additional information on perched water 

zones at the INEEL is presented by Rightmire and 

Lewis (1987), Cecil et al. (1991, 1992), Magnuson 

and McElroy (1993), and McElroy (1993). 

2.1.5.3.8 Upper Perched Zone—

Based on the perched water data, it appears the 

upper perched groundwater bodies are formed by 

the relatively low vertical hydraulic conductivity 

in the sedimentary interbeds. Of particular 

importance to the formation of perched 

groundwater are the CD, D, and DE3 interbeds. 

Figure 2-21 shows the location of the upper basalt 

perched water zone that occurs between depths of 

100 to 190 ft (30 to 58 m).

The location and extent of the upper perched 

water zones are dependent not only on the 

stratigraphy but also on the location of the 

recharge source. It appears that water sources vary 

from the northern to the southern portions of the 

INTEC facility. Consequently, each area will be 

discussed in detail in the following sections. 

Northern INTEC. Twenty-three monitoring wells 

(including multiple completion wells) have been 

installed in the northern INTEC area to monitor 

the upper perched groundwater (i.e., groundwater 

that occurs less than 190 ft [58 m] bgs). Only a 

few of these wells still exist. Two perched 

groundwater bodies have been identified in the 

northern INTEC area. The upper perched 

groundwater body is present above the CD and D 

interbeds, and the lower perched groundwater 

body has been identified on the DE3 interbed. 

According to the lithology, the CD interbed occurs 

at depths between 113 and 119 ft (34.4 and 

36.3 m) bgs, the D interbed occurs at depths 

between 128 and 135 ft (39.0 and 41.1 m) bgs, and 

the DE3 interbed occurs at depths between 163 

and 170 ft (49.7 and 52 m) bgs. Based on available 

information, it appears that the perched 

groundwater between the CD and D interbeds is 

continuous over much of the northern INTEC area 

since these interbeds are only separated by 9 ft 

(3 m). 

In addition to these wells, Well MW-4 (1-in.-

diameter) is completed at the bottom of the D 

interbed and Well MW-10 (1-in. diameter) is 

completed in a fracture zone associated with the 

BC interbed. Historically, both of these wells have 

been dry. 

The extent of the upper basalt perched 

groundwater body is shown in Figure 2-21. Waterlevel elevations range from 4,797.3 to 4,845.3 ft 

(1,462.2 to 1,476.8 m) amsl and represent the 

average water table level throughout the 

monitoring period. The extent of the perched water 

above the CD and D interbeds is also illustrated in 

this figure. Perched groundwater is not known to 

occur above these interbeds outside the areas 

illustrated on the map. Where the perched water 

bodies overlap (i.e., in the vicinity of 

Wells CPP-33-4, CPP-33-1, and MW-5), the entire 

region between the CD and D interbeds is likely to 

be saturated. Otherwise, perched groundwater is 

only present above the associated interbed. 
2-42

LODG EP OLE S TRE E T

FI R S T REE T

RE DWOOD STREE T

ICPP-MON-A-021

ICPP-MON-A-022

ICPP-MON-P-001

ICPP-MON-P-002

ICPP-MON-P-003

ICPP-MON-P-004

ICPP-MON-P-005

ICPP-MON-P-006

ICPP-MON-P-007

ICPP-MON-P-008

ICPP-MON-P-009

ICPP-MON-P-010

ICPP-MON-P-011

ICPP-MON-P-013

ICPP-MON-P-014

ICPP-MON-P-015

ICPP-MON-P-016 ICPP-MON-P-017

ICPP-MON-P-018

ICPP-MON-P-020

ICPP-MON-P-024

USGS-040

USGS-041

USGS-042

USGS-043

USGS-044

USGS-045

USGS-046

USGS-047

USGS-048

USGS-049

USGS-051

USGS-052

USGS-059

USGS-067

USGS-122

USGS-123

CPP-33-2 CPP-33-1

CPP-33-3

CPP-33-4-1

PW-1

PW-2

PW-3

PW-5 PW-4

PW-6

CPP-33-2L CPP-33-1L

CPP-33-3L

CPP-33-5L

USGS-050

Figure 2-21. Upper perched water at INTEC. 
2-43

Based on the water-table configuration, it 

appears that multiple water sources are providing 

recharge to the upper basalt perched water body in 

the northern INTEC area. Seepage from the Big 

Lost River is a potential source of water in this 

area. The sewage treatment ponds, located east of 

the facility, provide approximately 1.25E+06 gal 

(4.73E+06 L) per month of recharge to the 

perched water body. This recharge has resulted in 

a water table elevation of 4,845 ft (1,477 m) amsl 

in the well completed near the sewage treatment 

ponds (Well CPP-MON-P-024). In the western 

portion of the perched water body and beneath the 

main portion of the facility, the recharge has 

produced a maximum water-table elevation of 

4,817.5 ft (1,468.4 m) in Well CPP-33-2. Between 

the eastern and western portions of the upper 

perched water body, the average groundwater 

elevation is 4,809.2 ft (1,465.8 m) in Well 

CPP-37-4. This water-table configuration indicates 

separate sources of water providing recharge to the 

eastern and western portions of the uppermost 

perched water body. The sewage treatment ponds 

have minimal, if any, impact upon the western 

portion of this perched water body. 

Occurring beneath the upper perched 

groundwater body, a deeper perched groundwater 

body has been identified at a depth of 

approximately 140 ft (43 m) bgs. This 

groundwater appears to be the result of the DE3 

interbed, which occurs between 163 and 169 ft bgs 

(49.7 and 51.5 m) in the northern INTEC area. 

Only three wells (MW-10 [2 in.], MW-12 [1 in.], 

and MW-20 [2 in.]) are completed in this perched 

groundwater body. Based upon monitoring data, 

the water-table elevation within this perched water 

zone varies from 4,769 to 4,779 ft (1,454 to 

1,457 m) amsl. 

Southern INTEC. The upper perched water 

bodies identified in the southern INTEC area are 

shown in Figure 2-21. The largest perched water 

body is the result of discharge to the percolation 

ponds and is monitored by Wells PW-1 through 

PW-6. In the vicinity of CPP-603, six wells 

(MW-7, MW-9, and MW-13 through MW-16) 

were installed to monitor perched water on the 

upper interbed that is present between 110 and 

130 ft (34 and 40 m) bgs. One triple completion 

well (MW-17) was installed to monitor for 

perched water on a deeper interbed occurring at 

approximately 190 ft (58 m) bgs. 

Wells PW-1 through PW-6 were installed 

adjacent to the percolation ponds to monitor the 

perched groundwater beneath the ponds. The 

hydrographs for these wells show similar 

fluctuation in the water level observed for Wells 

PW-1, PW-3, and PW-6, indicating these wells are 

effective in monitoring infiltration from the 

western percolation pond. The water-level 

fluctuation in Well PW-4 is opposite to the 

response observed in Wells PW-1, PW-3, and 

PW-6, indicating this well monitors infiltration 

from the eastern percolation pond. The water-level 

fluctuations in Wells PW-2 and PW-5 are fairly 

consistent, indicating these wells are influenced by 

discharge to either pond. 

Field aquifer tests were performed to 

determine the hydraulic conductivities for both 

basalts and sedimentary interbeds. Hydraulic 

conductivities determined in the field were fairly 

consistent, varying over only two orders of 

magnitude. Field hydraulic conductivities ranged 

from 0.11 to 8.3 ft/d (3.4 cm/d to 2.5 m/d), with an 

average of 3.3 ft/d (1.0 m/d). Significant 

differences in hydraulic conductivities were not 

observed between tests performed on basalts 

versus tests performed on sedimentary interbeds 

(i.e., interbeds CD, D, and DE2). The depths are 

approximately 110, 140, and 230 ft (34, 43, and 

70. m) bgs. A complete discussion of the sample 

collection and laboratory data package is described 

in Interim Data Results from the FY 93/94 

Perched Ground Water Investigation (INEL 

1994). 

The range of hydraulic conductivities 

determined from the field aquifer tests is within 

the range of hydraulic conductivities measured in 

the laboratory. The average hydraulic conductivity 

determined from the field tests is 3.3 ft/d 

(1.0 m/d), compared to an average of 1.96 ft/d 

(0.597 m/d) determined from the laboratory tests. 

Some of the difference between the two hydraulic 

conductivities may be attributed to the fact that the 

field tests measured horizontal hydraulic 

conductivity, whereas the laboratory tests 

measured vertical hydraulic conductivity. 

Typically, horizontal hydraulic conductivities are 
2-44

higher than the corresponding vertical hydraulic 

conductivities. 

Good correlation in the hydraulic conductivity 

values occurred from the boreholes where both 

field and laboratory measurements were 

performed. From the same zone in Well MW-4, 

the average hydraulic conductivity determined in 

the laboratory was 0.1 ft/d (3 cm/d), compared to 

the field-determined value of 0.11 ft/d (3.4 cm/d). 

Similarly in Well MW-6, the hydraulic 

conductivity determined in the laboratory was 

6.2 ft/d (1.9 m/d), compared to the fielddetermined value of 3.7 ft/d (1.1 m/d). These two 

wells are the only locations where both field and 

laboratory measurements were performed. 

2.1.5.3.9 Lower Basalt Perched 

Water Zone—A lower perched groundwater 

zone has been identified in the basalt between 320 

and 420 ft (98 and 130 m) bgs (see Figure 2-22). 

This water was first discovered in 1956 while 

drilling Well USGS-40; perched groundwater was 

encountered at a depth of 348 ft (106 m) 

(Robertson, Schoen, and Barraclough 1974). Since 

then, groundwater has been encountered in this 

zone during the drilling of Wells USGS-41, -43, 

-44, -50, -52, and MW-1, -17, and -18. Borehole 

neutron logs run in 1993 indicate perched water 

may still be present in this zone from Wells 

USGS-40, -43, -46, -51, and -52.

Only four wells are completed in the lower 

perched water zone that presently monitor waterlevel changes. Wells MW-1, MW-18, and 

USGS-50 are completed in the northern portion of 

the facility. Water in these wells is encountered at 

approximately 322, 407, and 383 ft (98.1, 124, and 

117 m) bgs, respectively. In the southern portion 

of the facility, only Well MW-17 is completed in 

this zone; water in this well is encountered at a 

depth of approximately 315 ft (96.0 m) bgs. Based 

on water quality information, the deep perched 

groundwater encountered by Well MW-17 is the 

result of discharge to the percolation ponds. 

Similar to the upper basalt perched water 

zone, the lower zone is thought to be formed by 

decreased permeability associated with 

sedimentary interbed layers. It appears that the 

lower perched groundwater has formed primarily 

on the DE8 interbed. The top of this interbed 

occurs beneath INTEC at depths ranging from 383 

to 426 ft (117 to 130 m) bgs. In the western 

portion of the facility, however, the DE6 interbed 

is responsible for creating perched groundwater 

associated with Wells USGS-40 and USGS-43. 

However, these data contain a high degree of 

uncertainty, since they consist of a combination of 

original driller’s logs (some dating back 40 yr), 

geophysical borehole logs, and monitoring wells 

that are completed in this zone. 

As shown in Figure 2-22, the lower perched 

water zone is probably not continuous beneath the 

entire facility and may actually consist of several 

individual perched water bodies. The north-south 

separation of the lower perched water bodies is 

based on the lack of perched groundwater (either 

through drilling or subsequent neutron logging) 

identified in Wells USGS-42, -45, -47, -48, and 

-49. Recharge to the southern perched water body 

is from wastewater discharged to the percolation 

ponds. Seepage from the Big Lost River is 

believed to be the source of recharge to the 

western portion of the northern perched water 

body. 

Water levels in this lower perched water zone 

have been monitored since the early 1960s in Well 

USGS-50. Other than during the late 1960s and 

early 1970s, the water level in this well has been 

fairly consistent. The water level during the stable 

period generally ranged between 4,530. and 

4,540. ft (1,381 and 1,384 m) amsl. In the late 

1960s/early 1970s, however, the water level 

increased by approximately 90 ft (30 m) in 

response to failure of the INTEC injection well. 

During this period, wastewater was discharged 

directly to the vadose zone from the INTEC 

injection well at a reported depth of 226 ft 

(68.9 m) bgs. 

The presence of a deep perched groundwater 

zone beneath the percolation ponds is indicated by 

the borehole neutron response from Wells 

USGS-51 and MW-17. Figure 2-23 shows the 

natural gamma log, stratigraphy, and epithermal 
2-45

LODG E P OLE S T REE T

FI RS T REE T

RE DWOO DS TREE T

ICPP-COR-A-023

ICPP-MON-A-021

ICPP-MON-A-022

ICPP-MON-P-001

ICPP-MON-P-002

ICPP-MON-P-003

ICPP-MON-P-004

ICPP-MON-P-005

ICPP-MON-P-006

ICPP-MON-P-007

ICPP-MON-P-008

ICPP-MON-P-009

ICPP-MON-P-010

ICPP-MON-P-011

ICPP-MON-P-013

ICPP-MON-P-014

ICPP-MON-P-015

ICPP-MON-P-016 ICPP-MON-P-017

ICPP-MON-P-018

ICPP-MON-P-020

ICPP-MON-P-024

USGS-040

USGS-041

USGS-042

USGS-043

USGS-044

USGS-045

USGS-046

USGS-047

USGS-048

USGS-049

USGS-051

USGS-052

USGS-059

USGS-067

USGS-122

USGS-123

CPP-33-2 CPP-33-1

CPP-33-3

CPP-33-4-1

PW-1

PW-2

PW-3

PW-5 PW-4

PW-6

CPP-33-2L CPP-33-1L

CPP-33-3L

CPP-33-5L

USGS-050

Figure 2-22. Lower perched water zone at INTEC. 
2-46

Figure 2-23. Natural gamma and neutron log for Well USGS-51. 
2-47

neutron logs run in 1984, 1985, and 1986 for 

Well USGS-51. Deflections of the neutron log to 

the left indicate increases in moisture content. As 

shown in this figure, two zones illustrate 

significantly increasing moisture content since the 

percolation ponds were placed into service in 

February 1984. These two zones occur from 145 

to 178 ft (44.2 to 54.3 m) and 274 to 332 ft (83.5 

to 101 m) bgs. 

2.1.5.4 Regional Aquifer. The general 

direction of regional groundwater movement 

underlying the INEEL is to the south and 

southwest. The average slope of the water table is 

about 4 ft/mi. In the northern part of the INEEL, 

near Birch Creek Valley, the water-table gradient 

is relatively low, sloping southward about 1 ft/mi 

(0.2 m/km) (Barraclough et al. 1967).

Groundwater in the Snake River Plain Aquifer 

generally occurs under unconfined conditions, but 

locally may be quasi-artesian or artesian (Nace et 

al. 1959). Layers of dense, massive basalt or 

sediments cause the quasi-artesian or artesian 

conditions as a result of their relatively low 

permeability. Nace et al. described quasi-artesian 

as the situation in which the groundwater level is 

first recognized in a borehole during drilling at a 

depth below the regional water table, and then the 

level rises significantly (5 to 50 ft) [2 to 20 m] to 

the level of the water table. This rise of the water 

level simulates artesian pressure but the conditions 

are not truly artesian. Nace et al. also noted that 

water levels in some wells in the Snake River 

Plain Aquifer respond to fluctuations in barometric 

pressure similar to wells in confined aquifers, 

indicating tight zones in the basalt may impede 

pressure equalization. True artesian or flowing 

artesian conditions in the Snake River Plain 

Aquifer were identified at Rupert, in parts of the 

Mud Lake Basin, and north of the American Falls 

Reservoir (Nace et al. 1959). Recharge to the 

aquifer is primarily by valley underflow from the 

mountains to the north and northeast of the plain 

and from infiltration of irrigation water. A small 

amount of recharge occurs directly from 

precipitation. Recharge to the aquifer within 

INEEL boundaries is primarily by underflow from 

the northeastern part of the plain and the Big Lost 

River (Bennett 1990). Significant amounts of 

recharge from the Big Lost River have caused 

water levels in some wells at the INEEL to rise as 

much as 6 ft (2 m) in a few months after high 

flows in the river (Barraclough, Lewis, and Jensen 

1982). Locally, the direction of groundwater flow 

is temporarily changed by recharge from the Big 

Lost River (Bennett 1990). 

The effective hydraulic conductivity of the 

basalt and interbedded sediments that compose the 

Snake River Plain Aquifer at and near the INEEL 

ranges from 1.0E−02 to 3.2E+04 ft/d (3.0E−03 to 

9.8E+03 m/d). This six-order-of-magnitude range 

was estimated from single-well aquifer tests in 

114 wells, and is mainly attributed to the physical 

characteristics and distribution of basalt flows and 

dikes (Anderson, Kuntz, and Davis 1999). 

According to Anderson, Kuntz, and Davis, the 

relative hydraulic conductivity distribution 

through a typical buried shield volcano in the 

ESRP suggests that the conductivity of a single 

lava field is greatest in near-vent volcanic deposits 

consisting of shelly pahoehoe and slab pahoehoe 

flows and bedded scoria, spatter, and ash. These 

near-vent deposits typically cover about 10 to 20% 

of the lava field. Hydraulic conductivity of basalt 

flows is least for thick, tube-fed pahoehoe flow 

and ponded flow inside the vent craters and 

topographic depressions. 

Anderson, Kuntz, and Davis (1999) compared 

the relationship between the thickness of the basalt 

and sediment layers in relation to the hydraulic 

conductivity. They concluded that stratigraphic 

intervals comprising thin basalt layers have more 

contacts, rubble zones, and cooling fractures. 

Hence, stratigraphic intervals have greater 

hydraulic conductivity than intervals of equal 

thickness comprising thick layers of basalt. 

Detailed analyses of the geologic control 

related to hydraulic conductivity, along with 

hydraulic testing results for the regional aquifer, 

are presented in reports by Walker (1960), 

Ackerman (1991), and Anderson, Kuntz, and 

Davis (1999). 

2.1.5.5 Infiltration. 

2.1.5.5.1 Rainfall—Natural recharge 

from precipitation also is available to support 

perched water bodies. Long-term average annual 
2-48

precipitation at the INEEL is 8.7 in. (22 cm), and 

pan evaporation is on the order of 43 in./yr 

(110 cm/yr). Nevertheless, water from snowmelt 

or heavy rains may infiltrate to a depth where it 

cannot be evaporated. Furthermore, many areas 

within INTEC are impervious and precipitation 

runs off to drainage ditches. The ditches are 

unlined and a significant fraction of infiltration is 

likely to occur along the ditches. 

A wide range in the rate of infiltration 

rates at the site is reported in the literature and 

site-specific studies. These values range from a 

low of 0.16 in./yr (0.41 cm/yr) to a high of 

4.9 in./yr (12 cm/yr). Most of the reported values 

in the literature are estimates based on the amount 

of precipitation and best guess estimates of the 

evapotranspiration rates for the area. 

One of the few actual field measurements 

conducted at the site is reported by Cecil et al. 

(1992). By measuring tritium and 36Cl profiles in 

the soil, test results yielded a range of 0.16 to 

4.3 in./yr (0.4 to 1.1 cm/yr) infiltration. 

A detailed investigation of water infiltration 

rates was conducted at the Central Facilities Area 

Landfill by Miller, Hammel, and Hall (1990). The 

CFA Landfill is covered by an earthen-based 

material from the surrounding area. Based on soil 

analysis, precipitation, and evapotranspiration 

estimates, a range of infiltration rates from 

1.0 in./yr (2.5 cm/yr) to 1.6 in./yr (4.1 cm/yr) was 

reported. This measured value is the same 

infiltration rate used in the WAG 3 modeling 

analysis (Rodriquez et al. 1997) for the Tank Farm 

area. 

The analysis of contaminant migration from 

the TFF was modeled using the measured value of 

1.6 in./yr (4.1 cm/yr) for the CFA Landfill. This is 

considered an appropriate value since the TFF area 

will be covered with an engineered barrier under 

the CERCLA program. Therefore, this infiltration 

rate was considered for the base-case modeling 

estimates. 

The infiltration rates at the INEEL are 

uncertain, especially for a given area such as the 

TFF. Therefore, the infiltration rates were treated 

as an uncertain parameter in the 

sensitivity/uncertainty analysis presented in 

Section 7. Three infiltration rates were considered 

in the uncertainty analysis: 12.4 cm/yr as the 

highest rate, 4.1 cm/yr as the base-case infiltration, 

and 1.1 cm/yr as the best-case infiltration. The low 

infiltration rate of 1.1 cm/yr is considered 

applicable to a well-engineered landfill cover. 

2.1.5.5.2 Man-Made Sources—

INTEC uses approximately 2.1 million gal 

(7.9 million L) of water per day. Two raw water 

wells and one potable water well supply water to 

INTEC. The water is used for process cooling, 

equipment cooling, steam production, process 

solutions, decontamination, fuel storage basin 

makeup, chemical laboratory use, regeneration of 

ion exchange units, fire protection, and human 

uses. Piping systems external to facility buildings 

are either buried or enclosed in utility tunnels. The 

INTEC water systems that were considered 

relevant to the water inventory study included the 

raw water, fire water, treated (softened) water, 

demineralized water, steam condensate, landscape 

water, potable water, service waste (industrial 

wastewater), and sanitary waste systems. The 

steam distribution system, HLLW process lines, 

and pipes used to distribute cooling water to the 

HLLW storage tanks in the Tank Farm were 

eliminated from leak testing during the water 

inventory study. The criteria used to eliminate 

these areas were low volume, secondary 

containment, distinct radiological or chemical 

signature, and close monitoring or visibility.

The primary water systems at INTEC include 

the raw water system, fire water system, water 

softeners (treated water system), and demineralizer 

(demineralized water system). The raw water 

system piping has an approximate length of 

6,250 ft (1,910 m) and an average flow of 

389 gal/min (1,470 L/m). Raw water is pumped 

from the aquifer from two production wells to the 

fire water storage tanks. The raw water feed tanks 

supply water to three distribution pumps. The 

firewater piping system has a length of 5 mi 

(8 km) with an average flow of 45 gal/min 

(170 L/min). This flow in the system is water used 

for watering lawns, safety showers, cooling of 

waste tanks and sump pump bearings, and flushing 

radiation monitor bowls. The treated (softened) 

water system has a length of 4,000 ft (1,000 m) 
2-49

and an average flow of 900 gal/min (3,000 L/min). 

Treated water is used for chemical process 

makeup and heat exchangers. The demineralized 

water system has a length of 4,200 ft (1,300 m) 

and an average flow of 1.5 gal/min (5.7 L/min). 

Demineralized water is used for process cooling, 

steam, and fuel storage basins. 

The steam condensate piping system has a 

length of 4,200 ft (1,300 m) and an average flow 

of 74 gal/min (280 L/min) between September and 

April. Primary steam use occurs between the 

months of September and March because of 

seasonal demands such as heating and freeze 

protection. Most of the steam is condensed and 

recycled or routed to the service waste system. 

Approximately 10% of the steam is either released 

to the atmosphere or discharged to the ground. 

The potable water system has a length of 2 mi 

(3 km) with an average flow of 61 gal/min 

(230 L/min). Potable water is supplied from two 

wells completed in the Snake River Plain Aquifer. 

Water is pumped to a storage tank. From there, 

three distribution pumps in building CPP-606 are 

used to supply potable water to INTEC facilities. 

The potable water system includes a chlorination 

system. 

The service waste piping system has a length 

of 2.4 mi (3.9 km) and an average flow of 

1,320 gal/min (5.00E+03 L/min). Raw water, 

treated water, demineralized water, and steam 

condensates are discharged to the service waste 

system. Waste streams that might be contaminated 

with radioactive materials are monitored before 

being discharged to the service waste system. If 

contamination is detected, the water is diverted to 

a holding tank for processing in the Process 

Equipment Waste Evaporator. Water in the service 

waste system is discharged to two percolation 

ponds located on the south end of the facility. 

The sanitary sewer piping system is 1.5 mi 

(2.4 km) long, and has an average flow rate of 

29 gal/min (110 L/min). Potable water and 

sanitary waste from INTEC facilities are 

discharged to the sanitary sewer system and 

gravity-drained to lift stations, where the waste is 

pumped to the INTEC Sewage Treatment Plant. 

The waste is then transferred to a series of 

treatment lagoons located at the northeast corner 

of INTEC. 

The seven landscape watering systems at 

INTEC are all located in the northern area and 

have historically covered a total area of 1.5 acres 

(0.61 ha). Approximately 20,000 gal/d 

(80,000 L/d) from the raw water, fire water, and 

potable water systems has been used to maintain 

lawns and landscaping at INTEC during the 

summer months. A total volume of 

2.35 million gal (8.90 million L) is used to water 

INTEC lawns each year. Consumptive use is 

calculated to require 784,000 gal (2,970,000 L) per 

growing season. Subtracting the consumption and 

evapotranspiration from the supplied water yields 

a net volume of 1.57 million gal (5.94 million L) 

available for infiltration and recharge of the 

perched water bodies. 

A summary of the estimated amount of water 

available to recharge the perched water bodies at 

INTEC is given in Table 2-8. Historically, 

approximately 720 million gal/yr (2.7 billion L/yr) 

are known to be recharging the perched water 

bodies at INTEC. This does not include an 

additional 38.4 million gal (145 million L) that 

could be leaking through the facility piping and 

going undetected by the current metering system. 

2.1.6 Geochemistry 

Site-specific geochemical modeling was 

not performed for this PA. However, assumptions 

based on basic geochemistry were used in the PA 

development. 

2.1.6.1 Sediment and Basalt Kd Values. 

Knowledge of the sorptive properties of 

contaminants is key to understanding contaminant 

movement at the TFF. The subsurface comprises 

surficial sediments, sedimentary interbeds, and 

basalt. There is limited site-specific adsorption 

information for contaminants in the subsurface 

environment at the INEEL (Del Debbio and 

Thomas 1989; Schmalz 1972). Distribution 

coefficients are available for cobalt, chromium, 

strontium, cesium, cadmium, mercury, and 

selenium for alluvium. For interbed sediments and 

basalt, distribution coefficients are available for 

cadmium, mercury, selenium, and strontium. 
2-50

Table 2-8. Estimated volume of water recharging the perched water bodies at INTEC. 

 Northern INTEC Facility-Wide 

 Source 

Volume 

(gal/yr) Percent 

Volume 

(gal/yr) Percent 

Service wastewater None 0 690,000,000 95.8 

Sewage treatment ponds 15,000,000 58 15,000,000 2.1 

Water system leaksa

 3,980,000 16 3,980,000 0.6 

Landscape irrigationa

 1,568,000 6 1,568,000 0.2 

Precipitation infiltration 3,800,000 15 8,000,000 1.1 

Steam condensate 1,300,000 5 1,700,000 0.2 

CPP-603 Basins None 0 49,275 0.1 

Total 25,648,000 100 720,297,275 100 

 

a. Estimate based on past leaks and irrigation practices. Actual loss from piping leaks is unknown. 

Limited availability of site-specific adsorption 

information for other radionuclides for sediments 

and basalts at the INEEL has resulted in the use of 

adsorption parameters measured for sediments and 

basalts from other sites. Table 2-9 summarizes the 

measured Kd values for the INEEL. 

Several investigators have published 

compendia of soil and sediment distribution 

coefficient data (Baes and Sharp 1983; Coughtrey, 

Jackson, and Thorne 1985; Sheppard and Thibault 

1990). The most thorough of these compendia is 

Sheppard and Thibault (1990), which contains a 

breakout of the distribution coefficients by major 

soil types (i.e., sand, loam, clay, and organic). Past 

compendia grouped all soil types together. 

Distribution coefficient data for basalts are not 

as readily available as data for soils and sediments. 

The main source of Kd values is basalts from the 

Nuclear Energy Agencies (NEA) sorption 

database (Ticknor and Ruegger 1989). These 

basalt sorption values are provided in Table 2-10. 

In addition, a survey of the literature for the 

WAG 3 modeling study (Rodriquez et al. 1997) is 

provided in Table 2-11. 

Sorption coefficient (Kd) values are considered 

sensitive parameters in predictive modeling of 

radionuclide transport in the subsurface. There are 

a limited number of actual measurements available 

for the radionuclides that are specific to the 

INEEL soils and basalt, as indicated by Table 2-9. 

Consequently, it was necessary to survey the 

literature for similar sites where actual 

radionuclide sorption studies have been conducted. 

Table 2-12 provides a list of Kd values for key 

radionuclides used in the predictive transport 

modeling. The rationale for the selection criteria is 

presented in Table 2-13. In the 

sensitivity/uncertainty analysis presented in 

Section 7, three separate combinations of the 

sorption coefficients were modeled to evaluate the 

sensitivity of this important transport parameter, 

1) worst-case transport, 2) conservative transport, 

and 3) realistic/best transport sorption coefficients. 

This method provided a means of comparing the 

bounding-case analysis (i.e., conservative 

transport sorption coefficients), presented in 

Sections 3 and 4, with the range of values 

presented in the literature. 

Column studies conducted on basalts at 

various moisture contents from saturation to 20% 

consistently resulted in essentially the same Kd

values (Porro, Newman, and Dunnivant 2000). 

This evidence indicates that Kd values for the 

saturated zone are applicable to the unsaturated 

zone. While the reliability of Kd values obtained 

from batch tests is still being debated, using 
2-51

Table 2-9. Summary of sorption coefficients measured at the INEEL. 

Kd Values (mL/g)a

Element Alluvium 

Interbed 

Sediment Basalt Reference 

Cd 4891-2864 10,115-8622 2319-785 Del Debbio and Thomas (1989) 

Co 56 Not measured Not measured Schmalz (1972) 

Cr 1.2 Not measured Not measured Schmalz (1972) 

Cs 950 Not measured Not measured Schmalz (1972) 

Hg 1921-236 673-72 87-9.2 Del Debbio and Thomas (1989) 

Se 63-5.8 17-4.9 3.4-0.29 Del Debbio and Thomas (1989) 

Sr 35-52 110-186 1.1-2.7 Del Debbio and Thomas (1989) 

24 Not measured Not measured Schmalz (1972) 

a. Kd values reported by Del Debbio and Thomas (1989) were measured using batch equilibrium sorption methods on crushed 

materials, which were passed through a 2-mm sieve. 

Table 2-10. Basalt Kd values from the NEA database. 

Element 

 NEA Basalt Kd Ranges 

(mL/g) 

 I 5.0E−02 

 Sr 2.0 to 1000 

 Tc 1.5E−01 to 5.5 
2-52

Table 2-11. Sorption coefficient ranges identified in WAG 3.a

Element Sediment Kd (Range) mL/g Basalt Kd (Range) mL/g 

C 5 (2-20)b — 

I 0.1 (0.02-5) — 

Sr 60 (35-186)c,d 6 (1-13)c,d

Tc 0d

 0d

 

a. Source: Rodriquez et al. 1997. 

b. Range from NEA (Ticknor and Ruegger 1989) and Sheppard and Thibault (1990). 

c. Source: Newman et al., in preparation, Evaluation of the Mobility of Am, Cs, Co, Pu, Sr, and U through INEEL Basalt and 

Interbed Materials: Summary Report of the INEL/Clemson University Laboratory Studies, ER-WAG7-82, INEEL internal 

report. 

d. Source: Del Debbio and Thomas 1989. 

Table 2-12. Overview of the radionuclide sorption coefficients (mL/g) used in transport modeling. 

Best Scenario 

Realistic 

Scenario 

Conservative 

Scenario 

Worst-Case 

Scenario 

C 20 20 5 2 

I 5 5 0.1 0.01 

Sr 24 24 18 12 

Interbed Sediment 

Sorption 

Coefficients 

(mL/g) 

Tc 0.1 0.1 0.01 0 

C 7.1 7.1 5 1.7 

I 1 1 0.1 0 

Sr 13 13 6 1 

Basalt Sorption 

Coefficients 

(mL/g) 

Tc 0.24 0.24 0.01 0 
2-53

Table 2-13. Selection rationale for the range of Kd values (mL/g) used in the sensitivity analysis. 

Radionuclide 

Range for Kd

values (mL/g) 

used in 

transport model Selection Rationalea

C (sediment) 2–20 Default: Ref. #1 loam soil Kd = 20 

Site Values: WAG 3 Ref. #2 and #3 Kd = 2–20 

Literature Confirmation: Sand Ref. #1 Kd = 5 

C (basalt) 1.7–7.1 Site Values: WAG 3 Kd = not reported 

Literature Confirmation: The lowest reported Ref. #4 Kd = 1.7 

I (sediment) 0.01–5 Default: Ref .#1 loam soil Kd = 5 

Site Values: WAG 3 Ref. #2 Kd = 5 

Literature Confirmation: No referenced value as low as Kd = 0.01 

I (basalt) 0–1 Site Values: WAG 3 Ref. #2 Kd = 0 

Literature Confirmation: Ref. #3 Kd = 0.05, Ref. #1 for sand Kd = 1 

Sr (sediment) 12–24 Ref. #1 loam soil Kd = 20 

Site Values: WAG 3 Ref.#2 Kd = 12–24 

Literature Confirmation: Ref. #5 reports a range of 35–52 alluvium 

and 110–186 for interbed sediments 

Sr (basalt) 1–13 Site Values: WAG 3 Ref. #2 Kd = 1–13 

Literature Confirmation: Ref. #5 reports a range of 1.1–2.7 for 

crushed basalt column tests, Ref. #4 Kd = 2–1000 

Tc (sediment) 0–0.1 Default: Ref. #1 loam soil Kd = 0.1 

Site Values: WAG 3 Ref. #2 Kd = 0 

Literature Confirmation: Ref. #3 lowest measured sand Kd = 0.02 

Tc (basalt) 0–0.24 Site Values: WAG 3 Ref. #2 Kd = 0 

Literature Confirmation: Ref. #5 measured Kd = 0.016 for crushed 

basalt column tests; Ref. #3 Kd = 0.15–5.5 

a. References: 

1. Sheppard and Thibault 1990. 

2. Rodriquez et al. 1997. 

3. NEA database (Ticknor and Ruegger 1989). 

4. Thibault, Sheppard, and Smith 1990. 

5. Del Debbio and Thomas 1989. 
2-54

saturated Kd values for unsaturated transport 

simulation is not a specific concern. 

The modeling study conducted for the 

Remedial Investigation/Baseline Risk Assessment

(RI/BRA) for INTEC (Rodriguez et al. 1997) used 

the sediment distribution coefficient for the 

alluvium and the interbed sediments, a zero 

distribution coefficient for the vadose zone basalts, 

and 1/25 of the sediment distribution coefficient 

for the aquifer basalts. The vadose zone basalts 

were assumed to have a zero distribution 

coefficient in the RI/BRA, based on the 

assumption that the water moves more quickly 

through vertical fractures in the vadose zone. This 

allows very little time for the contaminant to sorb 

to the basalt. In the aquifer, the RI/BRA modeling 

assumed that the water moves more slowly 

through the basalt fractures and matrix and there is 

sufficient time for the contaminants to sorb in the 

basalts. 

The RI/BRA modeling also based the aquifer 

basalt distribution coefficients on an evaluation of 

the ratio of the distribution coefficient for 

sediments to that for basalts from the Del Debbio 

and Thomas (1989) report. Here they found the 

ratio to the 1/25. The Del Debbio and Thomas 

(1989) report, however, provides conflicting 

information on the differences in sediment and 

basalt distribution coefficients. This report 

provides for higher distribution coefficients in the 

basalts (based on breakthrough curve studies) in 

comparison to the batch distribution coefficient 

study. Because of this, the distribution coefficients 

for basalts were treated as uncertain as well as the 

sediment distribution coefficients. 

Three separate combinations of sorption 

coefficients were evaluated in the uncertainty of 

the Kd values (see Section 7). The sorption 

coefficients listed for the conservative case were 

used in the bounding analysis presented in 

Sections 3 and 4. The remaining cases (i.e., worstcase and realistic/best) were used in the 

uncertainty analysis presented in Section 7. Listed 

below are guidelines and criteria used in the 

distribution coefficient selection process. 

2.1.6.1.1 Strontium—The basalt 

values were based on the range presented in 

Table 2-13, with the worst-case scenario being 

assigned the lowest reported values in the range 

and the best/realistic case being assigned the 

highest value in the reported range. The 

conservative values (i.e., bounding case) were 

assigned the mid-point of the reported basalt Kd

range.

The sediments (i.e., interbed sediments in the 

unsaturated zone) were assigned Kd values based 

upon the modeling conducted for WAG 3 

CERCLA groundwater modeling at INTEC 

(Rodriquez et al. 1997). This modeling showed 

that a Kd value of 12 for the sediment resulted in 

90Sr groundwater concentrations higher than 

measured. A sediment value of 24 also was 

investigated and showed a closer match of the 

modeled concentrations with the measured 

concentrations. Therefore, the best/realistic 

scenario was assigned a Kd value of 24 for the 

interbed sediments, while the worst-case scenario 

was assigned a value of 12. The conservative case 

was assigned the mid-point value of the range. 

2.1.6.1.2 Carbon—The basalt values 

were based on the Atomic Energy Commission 

range presented in Table 2-13. The worst-case 

scenario was assigned the lowest reported value, 

while the best/realistic scenario was assigned the 

high value for the range. 

The sediment Kd values were taken from the 

range reported in Table 2-13, with the worst-case 

scenario being assigned the low value of the range 

and the best/realistic scenario being assigned the 

high value of the range. The conservative case was 

assigned a value of 5 mL/g, which was closer to 

the low end of the reported range. 

As noted above, sand Kd values were selected 

for 14C for transport simulations in the basalts 

where site-specific basalt Kd values were not 

available (Table 2-9). The published sand Kd

values were obtained from Sheppard and Thibault 

(1990). It was assumed that the lower sand Kd

values are more representative of the sorptive 

behavior of 14C in basalt. This assumption is based 

on the coating of weathered mineral or chemical 

precipitates on sand grains and the irregular 

fracture surfaces that can react with the 

radionucludies. While the surface area to the void-
2-55

space ratio for fractured media is less than a 

porous media (thereby reducing the sorptive 

capability for a fractured media), the presence of 

abundant iron oxides in basalt fractures capable of 

sorbing radionuclides offset the surface area 

effects. The degree of this offset is uncertain and 

difficult to apply to natural systems. 

This approach allows a degree of retardation 

of 14C in the basalts. The lower sand Kd values, 

compared with the higher Kd values for finer-grain 

sediment, are conservative on a relative basis. 

However, there is an acknowledged degree of 

uncertainty because of the lack of direct Kd

measurements available for 14C in basalt. 

2.1.6.1.3 Technetium—The basalt 

value for the worst-case scenario was based on the 

value in Table 2-13 of zero. The best/realistic case 

was assigned a low Kd value of 0.24, which is in 

the low end of the range. This was done due to the 

fact that technetium has been found to be mobile 

at the INEEL Site. The conservative value was set 

to a very low value of 0.01, again due to reports 

that technetium is mobile at the INEEL.

The sediment Kd value for the worst-case was 

taken as zero, while the realistic case was assigned 

a value of 0.1 based on values in Table 2-13 for 

loam soil. The conservative value was assigned a 

value one order of magnitude less than the loam 

soil value. 

2.1.6.1.4 Iodine—The basalt value for 

the worst-case was assigned a value of zero, while 

the best/realistic case was assigned a value of 1.0 

based on a reasonable upper end for the values 

reported in Table 2-13. The sediment Kd value for 

the worst-case was assigned a value of 0.01, while 

the best/realistic case was assigned a value of 5.0. 

The conservative case was assigned a value of 0.1.

2.1.6.2 Grout Kd Values. The grout 

sorption coefficients Kd values were evaluated for 

the three groundwater pathway scenarios. A 

review of the Bradbury and Sarott (1995) study on 

“Sorption Databases for the Cementitious NearField of a LLW Repository for Performance 

Assessment” indicates that the Kd values presented 

in their study were based on the conservative end 

of the possible choices (i.e., lower Kd values). 

Therefore, a combination of grout Kd values were 

chosen based on Bradbury and Sarott (1995), 

Allard (1985), and other studies specific to the 

elements of interest.

Concrete Kd values for reducing conditions 

have been chosen. The concrete is expected to 

exhibit strong reducing conditions (Eh from –300 

to –500 mV) as do most concrete systems. In 

addition, the closure system will also consist of a 

mixture of concrete and fly ash, slag, or other 

substances to ensure reducing conditions in the 

grout. 

The concrete Kd values were chosen from 

Region II of Bradbury and Sarott (1995). In 

Region II, the pore water composition is 

dominated by portlandite (Ca(OH)2), which fixes 

the pH at 12.5. The portlandite will be removed 

slowly by groundwater flow but the quantities 

contained in the cement are so large that this phase 

buffers the system over very long periods of time. 

Bradbury and Sarott (1995) also discuss a 

Region III for concrete systems, where the 

removal of Ca(OH)2 has become significant and 

the pH falls continuously. The calcium-silicatehydrate (CSH) gel is no longer stable and begins 

to dissolve incongruently. However, investigation 

into the original paper on these concrete regions 

by Atkinson, Everitt, and Gappy (1989) reveals 

that concrete is not expected to reach this state 

until 100,000 yr. Since the PA compliance period 

is only 1000 yr, the concrete Kd values are not 

expected to change over this period of time and 

therefore are held constant. This assumption is 

supported by the concrete degradation analysis 

presented in Appendix E. Graphs of the hydraulic 

conductivity for the tank/grout show no increase 

over the 1000-yr time period. The concrete 

degradation sensitivity analysis also shows that the 

tank/grout hydraulic conductivity does not 

increase until 800 yr, for the least conservative 

case, with only a one order of magnitude change in 

the hydraulic conductivity. Thus, the amount of 

water available for loss of Ca(OH)2 is still 

minimal. For conservatism, the PA assumes that 

the tank fails at 500 yr; however, the degradation 

analysis indicates the tank/grout life is much 

longer. In addition, graphs of carbonation 

degradation and calcium hydroxide leaching do 
2-56

not show any significant changes during the 

1000-yr compliance period for loss of Ca(OH)2. 

The selection of concrete Kd values for the 

three groundwater pathway scenarios are 

presented in Table 2-14. The following discussion 

provides justification for each set of sorption 

coefficients for each element shown in Table 2-14. 

2.1.6.2.1 Strontium—Ewart, Terry, 

and Williams (1985) report strontium distribution 

ratios for hardened cement paste and concrete 

between 0.001 and 0.004 m3

/kg with very little 

dependency on concentration. Atkinson and 

Nickerson (1988) summarized the results from 

different types of tests and gave a best-estimate 

range for strontium of 0.003 to 0.006 m3

/kg. Due 

to the small difference in Kd values reported for 

strontium, the worst-case scenario was assigned 

the low value of 0.001 m3

/kg, and the value was 

scaled for the other scenarios to a high value of 

0.006 m3

/kg for the best/realistic scenario.

2.1.6.2.2 Technetium—Sorption data 

for technetium on cementitious materials are 

sparse. Under reducing conditions, technetium is 

present as hydrolyzed technetium (IV) species. In 

some recent work, using technetium (IV) at trace 

levels, distribution ratios of 5 m3/kg have been 

reported (Bayliss et al. 1991). Technetium may be 

expected to sorb strongly under reducing 

conditions at high pH and Bradbury and Sarott 

(1995) chose a conservative value (i.e., low value) 

of 1 m3

/kg for reducing conditions. Therefore, for 

the TFF PA the conservative value of 1 m3

/kg was 

chosen for the worst-case scenario, with the Kd

values being scaled up to the measured value of 

5 m3

/kg for the best/realistic Kd value.

2.1.6.2.3 Iodine—Iodine is assumed 

to be present as I-

 under oxidizing and reducing 

conditions (Bradbury and Sarott 1995). Many 

studies on the sorption of I-

 on cement paste at 

high pH indicate that its sorption is low but finite. 

Bradbury and Sarott (1995) selected a 

conservative value (i.e., lower value) of 

0.002 m3

/kg for I- in reducing concrete systems. 

Allard et al. (1985) recommended a Kd value of 

0.03 m3

/kg for I- in concrete systems. Therefore, 

for the TFF PA the conservative value of 

0.002 m3

/kg was chosen for the worst-case 

scenario, with the Kd value being scaled up to the 

recommend value of 0.03 m3

/kg by Allard et al. 

(1985) for the best/realistic case.

Table 2-14. Solubility limits for the TFF grout.

 Solubility Limit 

Nuclide 

Half-life 

(yr) 

Specific Activity

(Ci/g) Moles/L Atoms/L Ci/L 

241Am 4.32E+02 3.44E+00 3.70E−06 2.23E+18 3.06E−03 

243Am 7.38E+03 2.00E−01 3.70E−06 2.23E+18 1.79E−04 

237Np 2.14E+06 7.05E−04 1.00E−10 6.02E+13 1.67E−11 

238Pu 8.78E+01 1.71E+01 3.20E−11 1.93E+13 1.30E−07 

239Pu 2.41E+04 6.22E−02 3.20E−11 1.93E+13 4.74E−10 

240Pu 6.54E+03 2.28E−01 3.20E−11 1.93E+13 1.75E−09 

241Pu 1.44E+01 1.03E+02 3.20E−11 1.93E+13 7.93E−07 

242Pu 3.76E+05 3.93E−03 3.20E−11 1.93E+13 3.04E−11 

230Th 7.70E+04 2.02E−02 2.50E−10 1.51E+14 1.16E−09 

234U 2.45E+05 6.24E−03 2.50E−11 1.51E+13 3.64E−11 

235U 7.04E+08 2.16E−06 2.50E−11 1.51E+13 1.27E−14 

236U 3.42E+06 4.43E−04 2.50E−11 1.51E+13 2.61E−12 

238U 4.47E+09 3.36E−07 2.50E−11 1.51E+13 2.00E−15 
2-57

2.1.6.2.4 Carbon—Allard, Persson, 

and Torstenflt (1985) recommend a Kd value for 

carbon of 5 m3

/kg. Studies by Allard, Torstenflt, 

and Andersson (1981) and Bayliss et al. (1988) 

have investigated the sorption of 14C as 14CO3

2- in 

cement/concrete systems. Generally, very high 

sorption values (approximately 10 m3

/kg) have 

been reported. Therefore, for the TFF PA the value 

of 10 m3

/kg has been chosen for the best/realistic 

case and the value scaled down to 1 m3

/kg for the 

worst-case scenario.

2.1.6.3 Solubility Limits. Solubility limits 

for selected nuclides important to the groundwater 

pathway are shown in Table 2-14 and were taken 

from Allard, Persson, and Torstenflt (1985). 

(Solubility limits were not applied to the 

contaminants in the grout for this PA). Evaluation 

of the maximum solubility limits from Allard, 

Persson, and Torstenflt (1985) and a comparison 

to the pore-water concentrations in the tank grout 

system indicated little difference between the 

solubility limits and pore-water concentrations 

(i.e., a factor of 10 or less). Therefore, solubility 

limits were not applied in this PA.

2.1.7 Natural Resources 

2.1.7.1 Geologic Resources. Geologic 

resources at the INEEL are very limited in nature. 

INEEL mineral resources include sand, gravel, 

pumice, silt, clay, and aggregate. These resources 

are extracted at several quarries or pits at the 

INEEL and are used for road construction and 

maintenance, waste burial activities, and 

ornamental landscaping. INTEC uses mineral 

materials extracted from the TRA Gravel Pit, 1 mi 

(2 km) west of INTEC, and the Lincoln Boulevard 

Gravel Pit, approximately 7 mi (10 km) north of 

INTEC. The geologic history of the ESRP makes 

the potential for petroleum products at the INEEL 

very low. The potential for geothermal energy 

exists at the INEEL; however, a study conducted 

in 1979 identified no economic geothermal 

resources (Mitchell, Johnson, and Anderson 1980).

2.1.7.2 Water Resources. The Snake River 

Plain Aquifer, one of the largest and most 

productive groundwater resources in the United 

States, underlies the INEEL and adjacent 

properties. The aquifer is listed as a Class I aquifer 

and was designated by EPA as a sole-source 

aquifer in 1991 (EPA 1990). Groundwater from 

this aquifer supplies most of the water for the area 

surrounding the INEEL and essentially all 

drinking water consumed within the ESRP 

(INEEL 1999b). The water from the aquifer is 

used for agriculture, food processing, aquaculture, 

and domestic, rural, public, and livestock water 

supplies. In total, nearly 4.7 trillion gal 

(18 trillion L) of water are drawn from the aquifer 

annually, with the majority going to agriculture 

(DOE-ID 1998).

Irrigated agriculture provides a significant 

portion of the economic base for the people of 

southern Idaho and the Snake River Plain Aquifer 

plays a major role in meeting irrigation 

requirements. The aquifer provides ground water 

for irrigation of over one third of the three million 

irrigated acres of the ESRP. It is estimated that 

over 127,000 people depend on the aquifer for 

domestic and municipal water needs. Total 

domestic water consumption is approximately 

46,000 acre-ft/yr (5.7 million m3

/yr); ground water 

discharge from well pumpage equals 

approximately 1.92 million acre-ft 

(2.37 billion m3

/yr) (EPA 1990). 

In addition to providing water for INEEL Site 

operations and agriculture, the aquifer supplies 

water for other industries. Water discharged from 

springs in the Twin Falls-Hagerman area is used to 

commercially raise fish. The spring water flow of 

1,660 ft3

/s (47.0 m3/s) constitutes 76% of the water 

used for the commercial production of fish in 

Idaho. Most of the fish farms discharge water 

directly into the Snake River. The discharges from 

Hagerman Springs also significantly contribute to 

the flow of the Snake River downstream of Twin 

Falls, Idaho. 

Groundwater in the aquifer generally flows 

from the northeastern recharge areas to the 

southwestern discharge areas. Nearly 

6.5E+06 acre-ft (8.0E+09 m3

) of water is 

discharged by the aquifer annually. Most of the 

discharge occurs as spring flow between 

Hagerman and Twin Falls. About 2.1E+06 acre-ft 

(2.6E+09 m3

/yr) of irrigation water is pumped 

from the Snake River Plain Aquifer in a typical 
2-58

year. About half of this water reenters the ground 

as return flow to the aquifer (INEEL 1999b). 

The altitude of the regional groundwater 

surface underlying the INEEL ranges from about 

4,600 ft (1,400 m) in the north to about 4,400 ft 

(1,300 m) near the southwest boundary of the 

INEEL. The average hydraulic gradient slopes to 

the south and southwest on the INEEL at about 

4 ft/mi (0.8 m/km). Within the INEEL boundaries, 

the depth below the land surface to the regional 

groundwater table ranges from 200 ft (60 m) in the 

northeast to 900 ft (300 m) in the west-southwest 

(INEEL 1999b). 

The Snake River Plain Aquifer is the only 

source of water used at the INEEL. The combined 

groundwater withdrawal averages approximately 

7E+06 gal/d (3E+07 L/d) or 8,000 acre-ft/yr 

(1E+07 m3

/yr). Table 2-15 lists the INEEL 

production wells, the depth of the wells, the depth 

to water at the wells, and the annual volume of 

water withdrawn from the wells. All wells 

withdraw water from the main body of the Snake 

River Plain Aquifer. The water withdrawn from 

each well is used for potable water on the Site, for 

ground maintenance, and for necessary facility 

operations (INEEL 1999b). 

The underflow of the INEEL (i.e., the amount 

of water passing directly under the INEEL 

boundaries) is approximately 4.7E+11 gal/yr 

(1.8E+12 L/yr). The INEEL consumption is less 

than 1% of the INEEL underflow and less than 

0.1% of the total annual aquifer discharge (INEEL 

1999b). 

2.1.8 Natural Background Radiation 

Monitoring and assessment activities are 

conducted to characterize existing radiological 

conditions at the INEEL and the surrounding 

environment. Results of these activities show that 

exposures resulting from airborne radionuclide 

emissions are well within applicable standards and 

are a small fraction of the dose from background 

sources. 

DOE has compared radiation levels monitored 

on and near the INEEL with those monitored at 

distant locations to determine radiological 

conditions. Results from onsite and boundary 

community locations include contributions from 

background conditions and INEEL emissions. 

These data show that over the most recent 5-yr 

period for which results are available (1992 

through 1996), average radiation exposure levels 

for boundary locations were no different than 

those at distant stations. The average annual dose 

measured by the Environmental Science and 

Research Foundation Inc. during 1996 was 

123 mrem for distant locations and 124 mrem for 

boundary community locations (DOE-ID 1997b). 

The corresponding 5-yr averages were 127 mrem 

for the distant group and 125 mrem for the 

boundary group. These differences are well within 

the range of normal variation. 

The offsite population could receive a 

radiation dose as a result of radiological conditions 

directly attributable to INEEL operations. The 

dose associated with radiological emissions is 

assessed annually to demonstrate compliance with 

the National Emissions Standards for Hazardous 

Air Pollutants (40 CFR 61, Subpart H, 2002). The 

effective annual dose equivalent to the maximally 

exposed individual resulting from radionuclide 

emissions from INEEL facilities during 1995 and 

1996 has been estimated at 0.018 mrem and 

0.031 mrem, respectively (DOE-ID 1996, 1997c). 

These doses are well below both the EPA dose 

limit (10 mrem/yr) and the dose received from 

background sources (about 360 mrem/yr). 

The annual collective dose to the population 

surrounding the INEEL, based on 1990 U.S. 

Census Bureau data, was estimated at 0.3 person 

rem. This estimate is based on the air emissions 

from all facilities that were expected to become 

operational before June 1, 1995 (DOE 1995). The 

dose applies to a total population of about 120,000 

people, resulting in an average individual dose of 

less than 0.003 mrem. For comparison, this 

population receives an annual collective dose from 

background sources of about 43,000 person rem. 

In 1999, the population doses were revised in the 

High-Level Waste Environmental Impact 

Statement (DOE 2002); as a result, the dose was 

estimated at about 0.09 person rem per year. 
2-59

Table 2-15. INEEL production wells and annual volume pumped.a,b

Well Name 

Depth of Well 

(ft bgs) 

Depth to Water 

(ft bgs) 

Annual Volume 

Pumped 

(gal) 

ANP-01 360 208 2.561E+06 

ANP-02 340 211 1.433E+06 

ANP-08 309 218 3.908E+05 

Badging Facility Well 644 489 5.760E+04 

CFA-1 639 468 1.473E+07 

CFA-2 681 471 1.448E+05 

CPP-01 586 460 1.834E+08c

CPP-02 605 460 1.834E+08c

CPP-04 700 462 1.834E+08c

CPP-05 695 447 1.834E+08c

EBR-I 1075 596 4.491E+04 

EBR II-1 745 632 2.767E+06d

EBR II-2 753 630 2.767E+06d

FET-1 330 199 1.427E+06 

FET-2 455 200 5.067E+05 

Fire Station Well 516 420 1.057E+04 

NRF-1 535 363 2.594E+06 

NRF-2 529 362 9.368E+06 

NRF-3 546 363 9.802E+04 

NRF-4 597 363 1.649E+07 

Rifle Range Well 620 508 9.115E+04 

RWMC Production 685 568 4.824E+05 

SPERT-1 653 456 3.871E+05 

SPERT-2 1,217 463 3.450E+05 

TRA-01 600 453 3.595E+07 

TRA-03 602 456 2.074E+06 

TRA-04 965 463 9.006E+07 

a. Note: All wells are withdrawing water from the main body of the Snake River Plain Aquifer and are used as drinking water wells with the 

exception of Wells ANP-08, Fire Station Well, and NRF-4, which are production wells for facility operations. 

b. Source: INEEL 1999b. 

c. Total for Wells CPP-01, CPP-02, CPP-04, and CPP-05. 

d. Total for Wells EBR II-1 and EBR II-2. 
2-60

2.2 Principal Facility Design 

Features 

This PA has been developed to support the 

DOE-ID Tier 1 Closure Plan for the TFF. Thus, 

the PA analysis is intended to define and bound 

the parameters of a closure action. Much of the 

site-specific and detailed facility and waste 

characterization data needed to conduct a 

projected performance analysis of the TFF closure 

action are not available at this time. Therefore, 

several simplifying and bounding assumptions 

have been made for the TFF PA. This section 

presents an overview of the TFF design features 

and factors considered in the TFF PA. 

2.2.1 Water Infiltration 

Currently, the tanks and vaults at the TFF are 

covered with 10 ft (3 m) of soil to provide 

shielding from the liquid waste. The contaminated 

soils contained in the TFF from historical liquid 

releases (mainly from piping) are being evaluated 

under CERCLA. Therefore, at the present time, 

consideration of a cover and designs to limit water 

infiltration are not available. The CERCLA ROD 

for INTEC describes the proposed remedy for the 

TFF soils (DOE-ID 1999). 

The ROD for INTEC and the TFF soils states 

that a final remedial action selection decision 

concerning the Tank Farm soils release sites has 

been postponed and will be developed following 

additional site characterization, risk analysis, and 

remedial alternative evaluation. An interim action 

is selected at the Tank Farm until a final decision 

is made by the Agencies. The remedy selected for 

the Tank Farm Soils Interim Action is 

Alternative 3—Institutional Controls with Surface 

Water Control. The interim remedy for controlling 

surface water infiltration includes (a) surface 

water run-on diversion channels sized to 

accommodate a 1-in-25-yr, 24-hr storm event and 

(b) grading and surface sealing the Tank Farm 

soils or sufficiently covering the Tank Farm so as 

to divert 80% of the precipitation falling atop the 

Tank Farm soils area. The interim action is 

projected to last 8 yr or until a final risk 

management decision is made and implemented 

by the Agencies. 

Because of the uncertainty of the CERCLA 

action, the timing, and final remedial action, the 

infiltration rate was treated as uncertain. Three 

infiltration rates are analyzed in the uncertainty 

analysis presented in Section 7. The interim action 

was not considered in the PA calculations. The 8-

yr life of the covering would not have a 

significant controlling influence on infiltration or 

radiation dose. 

Therefore, the use of an engineered cover at 

the TFF was not assessed for the conservative 

case (i.e., coompliance case) groundwater 

analyses. The higher 4.1 cm/yr infiltration rate for 

earthen cover material was used for the 

compliance analyses. 

2.2.2 Disposal Unit Cover Integrity 

As stated in Section 2.2.1, future actions in 

the TFF to remediate contaminated soils will be 

conducted under CERCLA. As such, the future 

cover design is not known at this time. Therefore, 

a bounding analysis was conducted for the TFF 

PA. The overlying 10 ft (3 m) of cover soil was 

assumed to be the only cover available for the 

bounding TFF analyses. An infiltration rate of 

4.1 cm/yr for an earthen cover was used in the 

analysis presented in Sections 3 and 4. Since a 

conservative analysis is being conducted by 

assuming that no cover is in place, the erosion of 

the overlying 10 ft (3 m) of soil also was assumed 

to not take place. In this way, the conservative 

analysis does not become a “worst-case” analysis. 

Additional uncertainty analyses are presented in 

Section 7, where three infiltration rates are 

investigated. 

2.2.3 Structural Stability 

The structural stability of the vaults and tanks 

is provided by the closure concept of grouting all 

voids. The High-Level Waste Environmental 

Impact Statement (DOE 2002) considered several 

alternatives for the TFF tanks, including filling 

them with low-level contaminated grout or 

leaving a remaining void above the first grout lift. 

In this PA, the entire tank is assumed to be filled 

with grout. Therefore, no voids will be available 

in the TFF system and structural failure (i.e., 

collapse) is not considered. 
2-61 

An analysis was conducted for the TFF PA to 

evaluate degradation of the concrete vaults, tanks, 

and grouted systems (see Section 3 and 

Appendix E). The degradation analysis was 

conducted to ensure that assumptions were 

bounding, including degradation of the vault and 

outer grout systems at 100 yr, piping releases at 

closure, and tank degradation at 500 yr. However, 

the degradation analysis was conducted on the 

assumption that the entire void space in the vaults 

and tanks was filled with grout. Therefore, no 

structural collapse was predicted. Instead, the 

degradation analysis was conducted to facilitate 

an evaluation of the changes in hydraulic 

properties with time. 

2.2.4 Inadvertent Intruder Barrier 

The concrete vault structures and the closure 

concept of grouting the void spaces in the tanks 

provide some protection from intruders. However, 

the main component of intruder prevention for the 

TFF PA involves the depth of waste in relation to 

the ground surface. In most intruder scenarios, the 

intruder is assumed to excavate a basement for 

housing construction. The depth of the waste in 

the TFF tanks precludes the intruder from 

contacting the waste. The concrete system may 

preclude an intruder from drilling through the 

vault/tank structure. Opinions on this protective 

quality vary, however, depending on the type of 

drilling equipment employed. An evaluation of 

the intruder scenarios and the potential barrier to 

intrusion provided by the TFF closure concept are 

discussed in detail in Section 5. 

2.3 Waste Characteristics 

Physical characteristics, hazardous 

constituents, and radionuclide concentrations have 

been measured in Tanks WM-182, WM-183, and 

WM-188. The hazardous constituent 

concentrations are documented in the Idaho 

Hazardous Waste Management Act/Resource 

Conservation and Recovery Act Closure Plan for 

Idaho Nuclear Technology and Engineering 

Center Tanks WM-182 and WM-183 (DOE-ID 

2001a). The analytical results and the radionuclide 

inventory are presented below. The tanks, sand 

pads, and piping all contribute to the radionuclide 

inventory. An inventory for each is described in 

the following sections. 

2.3.1 Tank Inventory 

The single tank radionuclide inventory is 

based on the most recent analytical results from 

sampling of Tanks WM-182, WM-183, and WM188 and historical data regarding the contents of 

the eleven 300,000-gal tanks. The TFF 

radionuclide inventory includes ten 300,000-gal 

tanks (WM-190 was never used), two sand pads, 

piping, and the four 30,000-gal tanks. The 

concentrations for radionuclides lacking current 

analytical data have been estimated using the 

ORIGEN2 model (Croff 1980). The model used 

sodium-bearing waste as the nuclear fuel waste 

stream and radionuclide concentrations based on 

closure in 2016. The inventory is designed to be 

conservative based on data of liquid and solid tank 

contents and recent sampling data. The 

conservative approach was maintained by using 

the highest radionuclide concentrations in the 

radionuclide inventory calculations, rather than 

averaging the recently sampled tanks. 

To provide for a thorough sensitivity/ 

uncertainty analysis, four radionuclide inventories 

were prepared. The conservative inventory has 

been prepared as the compliance inventory for 

comparison to the radiation dose criteria 

established in DOE Order 435.1, “Radioactive 

Waste Management.” The other radionuclide 

inventories have been prepared to be included in 

the PA uncertainty analysis described in Section 7 

of this report. The scenarios include the worstcase, realistic, and best inventories. 

The conservative radionuclide inventory 

adjusted the radionuclide quantity and 

concentration to produce a very conservative yet 

more realistic scenario than used in the worst-case 

scenario. The worst-case scenario assumes the 

tank cleaning will not be effective (sodiumbearing waste remains unaltered at a depth of 

1 in.). The conservative scenario assumes the tank 

cleaning is marginally successful. The solid 

residual is reduced by 10% to 2,085 kg, and the 

liquid concentration reduced by half. The 

conservative scenario and the remaining scenarios 

use predicted values for 129I and 99Tc. 
2-62 

The concentrations of 129I and 99Tc were 

recognized as being higher than expected based 

on process knowledge and the fission yield of 

235U. Historically, it has been difficult to measure 

129I and 99Tc in TFF liquids and solids. Excellent 

radiochemical separations are difficult to achieve 

in a sample matrix with relatively high 

concentrations of 137Cs and 90Sr. Detection of 129I

and 99Tc are biased high by interference of these 

radionuclides.b

The conservative (compliance) inventory 

is significantly greater than the expected results of 

the planned tank cleaning. Recent cleaning of 

Tank WM-182 has shown a dramatic decrease in 

solid mass and liquid concentrations. 

Additionally, using WM-188 solid sample data is 

conservative because this tank contained the 

highest radionuclide content of any of the three 

recently sampled tanks. Based on past tank 

operating history and solids sample data, Tank 

WM-188 solids should represent the highest 

radionuclide concentrations remaining in tank 

solids (Tyson 2002). 

The assumption that 2,085 kg of solids in 

the compliance inventory is conservative because 

tank mockup testing and recent cleaning in Tank 

WM-182 have shown that tank solids can be 

flushed and pumped to much lower quantities. 

Only partial credit is taken for the liquid-waste 

concentration reduction that will occur during 

cleaning. The radionuclide concentration is 

assumed to be 50 % of the sodium-bearing waste 

sampled in Tank WM-188. 

The two additional inventories, the 

realistic and best scenarios included more efficient 

waste removal. The realistic scenario assumes a 

25% reduction in solid waste and an 80% 

reduction in liquid concentration. The best 

scenario assumes a a 50% reduction in solid waste 

and an 95% reduction in liquid concentration. 

 

b. Letter from M. C. Swenson, INEEL, to V. L. Jacobson, 

INEEL, 2002, “Response to NRC Questions on the 

Radiological Source Term for I-129 and TC-99 in Tank Farm 

Waste,” MCS-05-02, October 22. 

TFF Tanks WM-182, WM-183, and WM-188 

were sampled recently to determine the presence 

and quantity of radioactive and hazardous 

constituents (Portage Environmental 2002, 

Patterson 1999). These three tanks are the only 

Tank Farm tanks that have been sampled recently 

and directly. The samples were collected using a 

remote sampling device. Previously, samples were 

collected by jetting (steam jets) primarily liquid 

waste to holding tanks in the New Waste 

Calcining Facility (NWCF) and the Waste 

Calcining Facility. The samples were collected for 

operational data and not intended for 

characterizing the solid and liquid in the tanks. 

The recent sampling events were designed to 

collect solid and liquid data directly from the tank 

heels. The heels in this case refer to the condition 

when the tanks are emptied to 10,000 to 

15,000 gal of liquid and solid. The data from these 

tanks serve as an initial characterization of the 

waste stream and provide the most recent and 

defensible data of tank solids and liquids. 

The most recent data from the tanks were 

examined to determine which tank contained the 

greatest radionuclide concentrations. 

The analytical data obtained from Tanks 

WM-182, WM-183, and WM-188 reflect the 

quantities of radionuclides present when the tank 

samples were analyzed. Tank WM-188 

radionuclide concentrations were the highest of 

the three tanks and were used to calculate the 

bounding tank. 137Cs was used as an indicator for 

residual radionuclide concentrations in the tanks 

during preparation of this source term. 

Historically, 137Cs has been used in conjunction 

with ORIGEN2 to characterize waste for 

treatment in the NWCF. 137Cs was used as the 

indicator radionuclide because it is detected 

accurately and easily using gamma spectroscopy. 

Solid and liquid data are accurately produced by 

this method. 137Cs has been consistently 

monitored during fuel reprocessing operations, 

and its gamma energy (0.662 Mev) is high enough 

to provide good detection. 137Ba is the daughter of 

137Cs. 137Ba produces the gamma photon measured 

at (0.662Mev). By convention, the result is 

reported as 137Cs. 137Ba and 137Cs are in secular 

equilibrium.
2-63 

Historically, 137Cs has been used during the 

operations period of the NWCF to predict 

concentrations of other nuclides that could not be 

detected as easily or efficiently. The 137Cs data 

and the ORIGEN2 numerical code for predicting 

nuclide concentration in nuclear fuel (or in this 

case, nuclear fuel waste streams during and after 

reprocessing) has successfully been used in the 

past by INTEC Operations to predict 

concentrations of various nuclides in waste 

streams to be calcined (Staiger and Millet 2000). 

The radionuclide inventory for the tanks is 

based on the same premise that was used by 

INTEC Operations. The source term was 

established initially based on ORIGEN2 data, then 

adjusted to the 137Cs concentration in Tank 

WM-188. The mean concentration of 137Cs in 

Tank WM-188 was used in the calculations to 

adjust the source term generated by ORIGEN2. 

Radionuclides detected in WM-188 were input 

directly into the source term and the ORIGEN2 

values were removed. Many nuclides are not 

detected in the TFF waste stream because of the 

relatively high concentrations of radioactivity in 

the waste. The sample matrix makes it difficult to 

detect radionuclides with low energy emissions or 

those in relatively low concentrations. The utility 

of using ORIGEN2 data is increased because it 

provides a starting point to calculate a 

concentration. Total strontium is used for 90Sr 

because of the short half-lives of other strontium 

isotopes. The analytical data are presented in 

Appendix A. 

The mean concentration for each nuclide in 

the respective media (solid, liquid) was calculated 

and then decayed to 2016 using the following 

formula: 

. t / T1⁄2 A A e

0 693

0

− = (2-1) 

where 

A = activity after time t 

A0 = initial activity 

e = natural log 

t = decay time 

T1/2 = radioactive half-life of the 

nuclide. 

Activity levels for radionuclides that weren’t 

analyzed came from ORIGEN2 modeling results 

(Staiger and Millet 2000). The following 

description of ORIGEN2 was taken from the 

Radiation Safety Information Computational 

Center home page of the Oak Ridge National 

Laboratory, Oak Ridge, Tennessee (Radiation 

Safety Information Computational Center 2002): 

ORIGEN is a computer code system 

for calculating the buildup, decay, and 

processing of radioactive materials. 

ORIGEN2 is a revised version of 

ORIGEN and incorporates updates of 

the reactor models, cross sections, 

fission product yields, decay data, and 

decay photon data, as well as the 

source code. 

ORIGEN2 uses a matrix exponential 

method to solve a large system of 

coupled, linear, first-order ordinary 

differential equations with constant 

coefficients. 

ORIGEN2 has been variably 

dimensioned to allow the user to tailor 

the size of the executable module to 

the problem size and/or the available 

computer space. 

The radionuclide quantities in the ORIGEN2 

output were based on the burnup rates and relative 

percentages of fuel types present in the sodiumbearing waste. These numbers had been 

normalized to a 137Cs quantity of 1 Ci. The ratio of 

the 137Cs level in ORIGEN2 to the 137Cs level in 

the WM-188 samples provided a factor by which 

the ORIGEN2 radionuclide quantities could be 

multiplied to provide equivalent WM-188 

radionuclide quantities. The ratio between 137Cs in 

the ORIGEN2 numbers and the sample data for 

137Cs in Tank WM-188 is the ratio factor listed in 

Tables A-7 and A-8. 

Closure of the individual TFF tanks is 

currently planned to occur using the following 

general sequence (see Section 1.2.2): 
2-64 

1. Remove liquid waste in tanks 

2. Empty tank to heel with existing jets 

3. Flush tank and piping 

4. Empty to heel with existing jets 

5. Wash tank and remove waste 

6. Video and sample heel residuals 

7. Evaluate tank residuals against closure 

performance objectives 

8. Displace heel with grout 

9. Fill piping with grout 

10. Fill tank vault with grout 

11. Fill tank with grout. 

Steps 1 through 3 are considered part of the 

tank deactivation. Steps 4 through 7 are required 

to remove waste and decontaminate a given tank. 

These steps may be repeated, as necessary, until 

the decontamination sequence is no longer 

effective (i.e., additional decontamination is no 

longer economically practical). The last four steps 

will remove remaining tank residuals to the extent 

readily feasible, stabilize the residuals, and leave 

the wastes in a solid physical form. 

The assumptions given below are based on the 

closure concept presented above. The bounding 

tank inventory does not assume that dilution of the 

liquid or solid waste occurs during washing of the 

tanks. Credit is only taken for emptying the tanks 

to a 1-in. (3-cm) solid heel and a 1,318-gal 

(5,992-L) liquid heel. However, as presented later 

in the groundwater pathway and intruder analyses, 

credit for washing and removal of a portion of the 

liquid fraction of the waste is necessary to meet 

the performance objectives. 

The radionuclide inventory for closure of the 

300,000-gal tanks is based on the slurry heel 

volumes for solids and liquid stated in EDF-1920 

(Tyson 2002). The following assumptions were 

used in the development of the bounding source 

inventory (Staiger and Millet 2000): 

1. A waste consisting of the highest measured 

concentration of the radionuclides found in 

any of the solutions or solids currently 

stored in Tanks WM-182, WM-183, and 

WM-188 represents the concentration of 

residual species in the worst-case inventory. 

2. Estimates of non-measured residuals are 

based on the highest calculated amounts 

conforming to those estimated by D. R. 

Wenzel,c

 normalized to the highest 137Cs 

concentration. 

3. The tank internals are flushed with water to 

clean material from all internal surfaces. 

4. The acid concentration in the liquid heel 

after flushing is > pH 2. 

5. Residuals associated with the solids heel 

are neither dissolved nor diluted. 

6. Closure operations are successful in 

removing the liquid heel to a final volume 

of 1,318 gal (4,989 L of waste). 

7. The solids heel is thoroughly mixed, 

leaving the interstitial liquid with a 

chemistry equivalent to the liquid heel. 

8. The makeup of the solids heel is 27% solidinterstitial liquid and 73% free liquid. 

9. The solids heel is removed from the tank 

down to a level of 1 in. 

Based on ORIGEN2 and analytical sampling 

results, the bounding tank source inventory in 

2016 will be 24,102 Ci. Activity in the solids 

portion accounts for 92% (22,133 Ci) of the total 

curies. Only a small number of the total nuclides 

contribute significantly to the radionuclide 

inventory. 238Pu, 241Pu, 90Sr, 90Y, 137Cs, 137Ba, and 

151Sm are major contributors to the transuranic 

activity and fission product activity, respectively. 

The conservative tank inventory at the assumed 

time of facility closure (2016) is provided in 

Table 2-16. 

 

c. Letter from D. R. Wenzel, INEEL, to N. E. Russel, INEEL, 

1997, “Calculation of Radionuclide Inventories for SodiumBearing Wastes,” WEN-23-97, November 26. 
2-65

Table 2-16. Conservative single tank inventory. 

Nuclide 

Liquid 

Activity 

(Ci) 

Solid Activity 

(Ci) 

Total Activity 

(Ci) Nuclide 

Liquid 

Activity 

(Ci) 

Solid Activity 

(Ci) 

Total Activity 

(Ci) 

Ac-225 5.7E−09 3.4E−08 4.0E−08 Pb-212 3.0E−05 1.8E−04 2.1E−04 

Ac-227 8.6E−07 5.2E−06 6.0E−06 Pb-214 2.0E−07 1.2E−06 1.4E−06 

Ac-228 1.0E−11 6.2E−11 7.2E−11 Pd-107 2.5E−04 1.5E−03 1.7E−03 

Ag-108m 6.3E−09 3.8E−08 4.4E−08 Pm-146 1.5E−04 8.9E−04 1.0E−03 

Am-241 1.8E−01 4.4E−01 6.2E−01 Pm-147 8.6E−02 5.2E−01 6.0E−01 

Am-242 2.2E−04 1.3E−03 1.6E−03 Po-210 8.6E−08 5.2E−07 6.0E−07 

Am-242m 2.2E−04 1.3E−03 1.6E−03 Po-211 0.0E+00 0.0E+00 0.0E+00 

Am-243 3.2E−04 1.9E−03 2.2E−03 Po-212 1.8E−05 1.1E−04 1.3E−04 

At-217 5.7E−09 3.4E−08 4.0E−08 Po-213 5.7E−09 3.4E−08 4.0E−08 

Ba-137m 5.7E+02 3.4E+03 4.0E+03 Po-214 2.0E−07 1.2E−06 1.4E−06 

Be-10 4.6E−08 2.8E−07 3.2E−07 Po-215 8.6E−07 5.2E−06 6.0E−06 

Bi-210 8.6E−08 5.2E−07 6.0E−07 Po-216 3.0E−05 1.8E−04 2.1E−04 

Bi-210m 3.3E−21 2.0E−20 2.3E−20 Po-218 2.0E−07 1.2E−06 1.4E−06 

Bi-211 8.6E−07 5.2E−06 6.0E−06 Pr-144 9.1E−08 5.5E−07 6.4E−07 

Bi-212 2.9E−05 1.7E−04 2.0E−04 Pr-144m 1.1E−09 6.5E−09 7.6E−09 

Bi-213 5.7E−09 3.4E−08 4.0E−08 Pu-236 1.7E−06 1.0E−05 1.2E−05 

Bi-214 2.0E−07 1.2E−06 1.4E−06 Pu-238 2.8E+00 1.4E+01 1.7E+01 

C-14 4.9E−01 1.1E−05 4.9E−01 Pu-239 3.5E−01 9.0E−01 1.2E+00 

Cd-113m 2.7E−02 1.7E−01 1.9E−01 Pu-240 1.6E−01 9.6E−01 1.1E+00 

Ce-142 4.6E−07 2.8E−06 3.2E−06 Pu-241 2.1E+00 1.3E+01 1.5E+01 

Ce-144 9.1E−08 5.5E−07 6.4E−07 Pu-242 1.2E−04 7.2E−04 8.4E−04 

Cf-249 2.6E−13 1.5E−12 1.8E−12 Pu-244 1.0E−11 6.2E−11 7.2E−11 

Cf-250 1.1E−13 6.5E−13 7.6E−13 Ra-223 8.6E−07 5.2E−06 6.0E−06 

Cf-251 4.1E−15 2.4E−14 2.8E−14 Ra-224 3.0E−05 1.8E−04 2.1E−04 

Cm-242 1.8E−04 1.1E−03 1.3E−03 Ra-225 5.7E−09 3.4E−08 4.0E−08 

Cm-243 3.2E−04 1.9E−03 2.2E−03 Ra-226 2.0E−07 1.2E−06 1.4E−06 

Cm-244 1.6E−02 9.6E−02 1.1E−01 Ra-228 1.0E−11 6.2E−11 7.2E−11 

Cm-245 4.6E−06 2.8E−05 3.2E−05 Rb-87 4.5E−07 2.7E−06 3.1E−06 

Cm-246 3.0E−07 1.8E−06 2.1E−06 Rh-102 5.7E−07 3.4E−06 4.0E−06 

Cm-247 3.3E−13 2.0E−12 2.3E−12 Rh-106 1.8E−06 1.1E−05 1.3E−05 

Cm-248 3.6E−13 2.2E-12 2.5E−12 Rn-219 8.6E−07 5.2E−06 6.0E−06 

Co-60 7.0E−02 1.4E−01 2.1E−01 Rn-220 3.0E−05 1.8E−04 2.1E−04 

Cs-134 6.0E−03 4.7E−02 5.3E−02 Rn-222 2.0E−07 1.2E−06 1.4E−06 

Cs-135 1.4E−02 8.3E−02 9.6E−02 Se-79 6.9E−03 4.1E−02 4.8E−02 

Cs-137 5.7E+02 3.4E+03 4.0E+03 Sm-146 4.2E−09 2.5E−08 3.0E−08 

Eu-150 1.7E−07 1.0E−06 1.2E−06 Sm-147 1.1E−07 6.9E−07 8.0E−07 
Table 2-16. (continued). 

2-66 

Nuclide 

Liquid 

Activity 

(Ci) 

Solid Activity 

(Ci) 

Total Activity 

(Ci) Nuclide 

Liquid 

Activity 

(Ci) 

Solid Activity 

(Ci) 

Total Activity 

(Ci) 

Eu-152 2.0E−02 1.2E−01 1.4E−01 Sm-148 5.7E−13 3.4E−12 4.0E−12 

Eu-154 9.1E−01 3.0E−01 1.2E+00 Sm-149 5.2E−14 3.1E−13 3.6E−13 

Eu-155 1.1E−01 2.4E+00 2.6E+00 Sm-151 4.7E+00 2.8E+01 3.3E+01 

Fr-221 5.7E−09 3.4E−08 4.0E−08 Sn-121m 7.3E-4 4.5E−03 5.2E−03 

Fr-223 1.2E−08 7.2E−08 8.4E−08 Sn-126 6.3E−03 3.8E−02 4.4E−02 

Gd-152 2.2E−14 1.3E−13 1.6E−13 Sr-90 4.1E+02 7.6E+03 8.0E+03 

H-3 8.0E−02 4.8E−01 5.6E−01 Tc-98 3.9E−08 2.4E−07 2.8E−07 

Ho-166m 6.9E−07 4.1E−06 4.8E−06 Tc-99 1.5E−01 9.0E−01 1.0E+00 

I-129 3.7E−04 2.2E−03 2.6E−03 Te-123 5.7E−15 3.4E−14 4.0E−14 

In-115 1.5E−12 8.9E−12 1.0E−11 Te-125m 1.9E−03 1.1E−02 1.3E−02 

La-138 3.0E−12 1.8E−11 2.1E−11 Th-227 8.6E−07 5.2E−06 6.0E−06 

Nb-93m 2.9E−02 1.7E−01 2.0E−01 Th-228 3.0E−05 1.8E−04 2.1E−04 

Nb-94 1.7E−02 7.7E+00 7.7E+00 Th-229 5.7E−09 3.4E−08 4.0E−08 

Nd-144 2.5E−11 1.5E−10 1.7E−10 Th-230 1.4E−05 8.3E−05 9.6E−05 

Ni-59 1.3E−02 1.6E−01 1.7E−01 Th-231 3.2E−04 1.9E−03 2.2E−03 

Ni-63 4.3E−01 2.6E+00 3.0E+00 Th-232 1.1E−11 6.5E−11 7.6E−11 

Np-237 1.7E−03 5.9E−03 7.6E−03 Th-234 3.2E−04 1.9E−03 2.2E−03 

Np-238 1.1E−06 6.5E−06 7.6E−06 Tl-207 8.6E−07 5.2E−06 6.0E−06 

Np-239 3.2E−04 1.9E−03 2.2E−03 Tl-208 1.0E−05 6.2E−05 7.2E−05 

Np-240m 1.0E−11 6.2E−11 7.2E−11 Tl-209 1.3E−10 7.6E−10 8.8E−10 

Pa-233 4.5E−02 2.7E−01 3.1E−01 Tm-171 6.9E−14 4.1E−13 4.8E−13 

Pa-231 1.5E−06 8.9E−06 1.0E−05 U-232 2.9E−05 1.7E−04 2.0E−04 

Pa-234 4.1E−07 2.4E−06 2.8E−06 U-233 3.7E−06 2.2E−05 2.6E−05 

Pa-234m 3.2E−04 1.9E−03 2.2E−03 U-234 1.3E−02 7.6E−02 8.8E−02 

Pb-209 5.7E−09 3.4E−08 4.0E−08 U-235 6.0E−05 4.1E−04 4.7E−04 

Pb-210 8.6E−08 5.2E−07 6.0E−07 U-236 3.2E−05 3.1E−03 3.1E−03 

Pb-211 8.6E−07 5.2E−06 6.0E−06 U-237 5.3E−05 3.2E−04 3.7E−04 

Ru-106 1.8E−06 1.1E−05 1.3E−05 U-238 8.2E−05 2.4E−04 3.3E−04 

Sb-125 7.4E−03 4.5E−02 5.2E−02 U-240 1.0E−11 6.2E−11 7.2E−11 

Sb-126 8.6E−04 5.2E−03 6.0E−03 Y-90 4.1E+02 7.6E+03 8.0E+03 

Sb-126m 6.3E−03 3.8E−02 4.4E−02 Zr-93 3.3E−02 2.0E−01 2.3E−01 

Ru-106 1.8E−06 1.1E−05 1.3E−05 
2-67

The TFF also has four belowground 

30,000-gal stainless-steel tanks designated as 

Tanks WM-103, -104, -105, and -106. These 

tanks last stored HLW in 1974 and stored 

evaporator condensate until 1983. The tanks were 

taken out of service in 1983 and the tank inlets 

have been cut and capped (INEEL 1999a). The 

contents of these tanks were sampled for RCRA 

constituents in 1990 and then emptied. The 

sampling analysis provided limited radionuclide 

information, as followsd

: 

• The sample for Tank WM-103 was 35 mR/hr 

and had a 137Cs content of 398 ± 17 

disintegration per second per milliliter 

(d/s/mL) (11 ± 0.46 nCi/mL). 

• The sample for Tank WM-104 was 35 mR/hr 

and had a 137Cs content of 2,780 ± 150 d/s/mL 

(75 ± 4.0 nCi/mL) and a 154Eu content of 

4.5 ± 1 d/s/mL (0.12 ± 0.03 nCi/mL). 

• The sample for Tank WM-105 was 35 mR/hr 

and had a 137Cs content of 192 ± 9 d/s/mL 

(5.2 ± 0.24 nCi/mL). 

• The sample for Tank WM-106 was 35 mR/hr, 

had a 137Cs content of 1,280 ± 59 d/s/mL 

(35 ± 1.6 nCi/mL), and a 134Cs concentration 

of 10.8 ± 1.3 d/s/mL (0.29 ± 0.035 nCi/mL). 

Using the same analogy used in the 

development of the conservative tank inventory 

for the 300,000-gal tanks, an estimate of the total 

137Cs activity in a 30,000-gal tank was made using 

the data from Staiger and Millet (2000): 

• The 30,000-gal tanks have the same 

composition in terms of the contributions to 

activity for liquid and solids as the 

300,000-gal tanks. 

• The 30,000-gal tanks contain 10% of the 

residuals contained in the 300,000-gal tanks. 

 

c. E-mail communication between K. Poor, Portage 

Environmental, Idaho Falls, Idaho, and Duke Engineering, 

Idaho Falls, Idaho, April 11, 2000. 

• The 300,000-gal tanks have 4.9E+06 mL of 

liquids and 1.25E+06 mL of solids. Therefore, 

the 30,000-gal tanks contain 4.9E+05 mL of 

liquids and 1.25E+05 mL of solids, for a total 

volume of 6.2E+05 mL of residual 

contamination. 

• The maximum 137Cs content of a given 

30,000-gal tank was calculated based on the 

sample for WM-104, with a 137Cs 

concentration of 75 nCi/mL. 

These assumptions result in a maximum 137Cs 

tank content of 0.046 Ci for the 30,000-gal tanks. 

A comparison of the 137Cs activity calculated 

for the 30,000-gal tanks (i.e., 0.046 Ci) with the 

137Cs content of the 300,000-gal tanks from the 

inventory indicates that the contamination levels 

in the 30,000-gal tanks are insignificant by 

comparison. Therefore, the inventory for the 

300,000-gal tanks is considered to bound any 

additional contamination that may be released 

from the 30,000-gal tanks. 

2.3.2 Sand Pad Inventory 

The sand pads in Tanks WM-185 and 

WM-187 are contaminated from accidental 

releases into the vaults in 1962. A description of 

the leakage into the tanks is available in Latchum 

et al. (1962). Subsequent to these releases, liquid 

from sources such as precipitation and irrigation 

was periodically pumped out of the vaults. 

This section presents the background 

information and the approach, assumptions, and 

data used to estimate the activity of radionuclides 

contained in the sand pad beneath the tanks at the 

TFF. The results from the analysis are provided 

for each radionuclide in the year 2016, which is 

the expected closure date. These activities have 

been estimated by modeling diffusion of 

radionuclides from the liquid into the sand pad for 

the contaminating event in 1962. The analysis 

then models radioactive decay and periodic 

flushing of the sand pad to estimate activities in 

2016. 

2.3.2.1 Background Information. Surface 

water leaking into the tank vault has been a 
2-68 

common occurrence since the beginning of Tank 

Farm operations. The top of the vault and tank 

structure allow surface water to partially fill the 

vault. Before and subsequent to the 1962 releases, 

water from precipitation, spring runoff and 

irrigation has been pumped out of the tank vaults 

at least semiannually. 

The amount of water collected in the vaults is 

documented in the INEEL internal memorandum, 

“HLLW Tank Sump Transfer History.”e

 The 

report documents over 100 transfers from Tanks 

WM-185 and WM-187. Reports are not available 

before 1962 but water transfers from the vaults are 

known to have occurred. Water transfers continue 

from vault sumps. The residual inventory 

predicted for 2016 is based on 38 “flushing” 

events when water infiltrated to the vault from 

leaks in the tank/vault roof and was then jetted out 

of tank vaults. The 38 events correspond to the 

number of years from 1962 to 2000. The actual 

number of water transfers from the tank vaults and 

associated leaching of radionuclides from the sand 

pad likely exceeds 130 for each vault to date. 

Water was jetted from the tank vault sumps at 

least twice yearly and will continue until each 

tank is closed. 

Details of the sand pad analysis are shown in 

Appendix A. The results from the analysis are 

provided for each radionuclide at activities in 

2016, which is the expected closure date. These 

activities have been estimated by modeling 

diffusion of radionuclides from the liquid into the 

sand pad for the contaminating event in 1962. The 

analysis then models radioactive decay and 

periodic flushing of the sand pad to estimate 

activities in 2016. 

The curb and sand pad at the bottom of the 

concrete vault were designed to cushion the tank 

bottom. Additionally, the curb, which has drain 

holes, allows liquid to drain from the sand pad 

beneath the tanks to the sumps outside the curb. 

The sumps were periodically used to remove 

water that had accumulated in the vaults. Liquid 

 

e. Letter from K.J. Rebish, INEEL, to J.M. Roberts and B.H. 

O’Brien, INEEL, 1993, “HLLW Tank Sump History,” 

KJR-10-93, October 15. 

may have been present in the vault from sources 

such as precipitation, irrigation activities, and 

accidental spills. The curb is an octagonal shape, 

creating and area 51 ft between the sides of the 

octagon (or 52 ft including the 6-in. curb). 

Because the vault floor is conical and the tank 

bottom is flat, the thickness of the sand pad ranges 

from about 2 in. (5 cm) at the tank center to about 

6 in. (15 cm) at the curb. The 6-in. × 6-in. curb 

was installed 6 ft (2 m) from the outside edge of 

the vault floor. This curb encircles the sand pad 

and creates an area that is 51 ft (16 m) wide. 

Because the vault floor is conical and the tank 

bottom is flat, the thickness of the sand pad ranges 

from about 3 or 4 in. (8 or 10 cm) at the tank 

center to about 5 or 6 in. (13 or 15 cm) at the curb. 

2.3.2.2 Estimating Approach. The 

approach to estimate the activities of the 

radionuclides contained in the TFF sand pad in 

2016 is described in this section. The analysis 

represents a different approach to determine the 

mass of radionuclides in the sand pad than 

previous analyses (Poloski 2000). Previous 

analyses assumed radionuclides entered the entire 

sand pad and reached an equilibrium partitioning 

in the sand and the sand pad was subsequently 

flushed.

The approach presented in this report assumes 

radionuclides enter the sand pad through diffusion 

at the curb/sand pad interface. This assumption is 

reasonable since the contaminated liquid, which 

entered the vault in 1962, appears to have filled 

the concrete vault from the bottom toward the 

vault top (i.e., there was no horizontal pressure 

gradient driving the liquid through the sand pad). 

Once contained in the sand pad, radioactive decay 

and periodic (annual) flushing were modeled to 

determine the present-day activities. Flushing is 

assumed to occur 38 times and then the inventory 

is decayed over 16 years to estimate 2016 

activities of all radionuclides. 

There are numerous published reference 

materials that present solutions to diffusion 

modeling equations (e.g., Crank 1975; Choy and 

Reible 2000). The following equation describes 

one-dimensional diffusion from a semi-infinite 

media with uniform initial concentration and zero 
2-69 

concentration at the surface. Equation (2-2) is 

shown in Choy and Reible (2000). 

∈ [ ∞) ∂

∂

















= ∂

∂

z 0, z

C

R

D

t

C

2

A

2

fA

A A( eff ) (2-2) 

where 

CA = concentration of radionuclide A 

(M/L3

) 

t = time (T) 

DA(eff) = diffusion coefficient of 

radionuclide A, assumed to be 

10-5 cm2

/s, which is conservative 

for sand (L2

/T) 

RfA = retardation factor of radionuclide 

A in sand (unitless) 

z = distance (L). 

The boundary conditions associated with 

Equation (2-2) and its solution include the 

following: 

( ) , 0 0 CA z t z=0 = t > (2-3) 

( ) , 0 CA z t z→∞ = CA0 t > (2-4) 

( ) = ∈ [ ∞) = , 0, 0 0 C z t C z A t A (2-5) 

where 

CA0 = initial concentration of 

radionuclide A (M/L3

). 

The boundary and initial conditions 

[Equations (2-3) through (2-5)] for Equation (2-2) 

describe a semi-infinite system with a constant 

initial concentration of radionuclide A (CA0) and 

represent transport (diffusion) toward the surface, 

where the concentration is maintained at zero. The 

zero concentration boundary condition implies 

that as soon as the contaminant diffuses to the 

surface, it is immediately removed. For mass 

transport in the sand pad, this assumption is 

conservative (i.e., it does not underestimate mass 

transfer at the liquid/sand pad interface). 

The analytical solution to Equation (2-2), 

using boundary and initial conditions in Equations 

(2-3) through (2-5) for the surface flux, jA(t), in 

units of M/L2

T is offered in Choy and Reible 

(2000): 

( ) 0 0 0 >

π = = tt

D R

j t C A ( eff ) fA

A z A (2-6) 

Equation (2-6) determines the amount of a 

contaminant that diffuses through the liquid/sand 

pad interface. Records from the accident in 1962 

(Latchum et al. 1962) indicate that the time from 

the introduction of the liquid into the vault to the 

time when the liquid was pumped out of the vault 

was approximately 24 hr. Thus, for this analysis 

the diffusion of radionuclides from the liquid into 

the sand pad is modeled for 24 hr. Equation (2-6) 

is used to compute the flux through the 

liquid/sand pad interface, and the flux time history 

is integrated over 24 hr to determine the total 

mass of each radionuclide that enters the sand 

pad. 

After the accident, the mass of each 

radionuclide in the sand pad is assumed to be 

impacted by radioactive decay and flushing. 

These processes are modeled by assuming the 

mass of radionuclides in the sand pad undergoes 

radioactive decay for a period of time representing 

the length of time between flushing events. Then 

the system is flushed. During flushing, the sand 

pad is assumed to be saturated and the 

radionuclides are partitioned at equilibrium 

between the liquid and solid phase according to 

their Kd values. The flushing is assumed to 

remove all liquid and radionuclides that have 

partitioned into the liquid phase from the sand pad 

except for the residual liquid. Thus, the 

radionuclides remaining in the sand pad are 

contained in the residual liquid and sorbed onto 

the sand. Then the mass of radionuclides in the 

sand pad is again calculated, assuming radioactive 

decay occurs over a period of time that represents 

the time between flushing events. The cycle of 

modeling radioactive decay and flushing is 

repeated until present-day activities are computed. 
2-70 

In the absence of data, it was assumed that the 

flushing events occurred once a year for 38 yr 

(i.e., each year the seasonal effects of precipitation 

and irrigation are purged from the vaults). After 

present-day activities are computed, these results 

are decayed over 16 years to 2016, which 

corresponds to the time of TFF closure. 

To determine the mass of radionuclides in the 

sand pad following the initial event in 1962, the 

radioactive decay and flushing of the sand pad is 

modeled. A number of existing programs were 

examined, including the DUST-MS computer 

code (Sullivan 1993 and 1996). However, the 

repeated radioactive decay and flushing cycles 

made using these models, especially data entry, a 

burden. Consequently, a FORTRAN computer 

code was developed to model radioactive decay 

and then flushing of the sand pad based on 

partition coefficients and the volume of sand, void 

space, and residual saturation. The mathematical 

formulation of this approach is provided below; 

the FORTRAN source code is presented in 

Appendix A. 

Radioactive decay is modeled using a firstorder differential equation with a rate constant 

that represents the radionuclide’s half-life. For 

chain decay, the same first-order differential 

equation is applicable; however, an additional 

term accounts for the increase in the mass of a 

daughter product. The following illustrates the 

appropriate equations for a three-member decay 

chain (Thorne et al. 2000). 

The differential equations that define the rate 

change of atoms in each compartment (i.e., for 

each radionuclide in the decay chain) are as 

follows: 

1 1

1 k q dt

dq = − (2-7) 

1 1 2 2

2 k q k q dt

dq = − (2-8) 

2 2 3 3

3 k q k q dt

dq = − (2-9) 

where 

qi = atoms of radionuclide i (unitless) 

t = time (yr) 

ki = decay constant for radionuclide i (yr-1). 

Solutions to the above differential equations 

obtained by Laplace transforms are as follows: 

k t q (t ) q ( ) e 1 1 1 0 − = (2-10) 

( ) k t k t k t e e q ( )e k k

k q ( )

q (t ) 2 2 2 1

1 2

1 1

2 0 0 − − − − + − =

 (2-11) 

( )( )

( )( ) ( )( )

( ) k t k t k t

k t k t

k t

e e q e k k

k q

e k k k k

k k q e k k k kk k q

e k k k k

k k q q t

2 3 3

2 3

1

(0) (0)

(0) (0)

(0) ( )

3

3 2

2 2

1 3 2 3

1 2 1

1 2 3 2

1 2 1

2 1 3 1

1 2 1

3

− − −

− −

−

− + −

+ − − + − −

+ − − = (2-12) 

The activity (Ci) of each compartment may be 

determined from these equations by multiplying 

by the appropriate decay constant. The activity for 

individual radionuclides that were not part of a 

decay chain was determined according to 

Equation (2-10). 

After accounting for radioactive decay, this 

analysis models flushing of the liquid in the sand 

pad using the following equations. For any 

radionuclide A, the total mass of radionuclide A 

in the sand pad is either sorbed onto the solid 

phase (i.e., sand) or in the liquid phase. That is, 

M TA = M SA + M LA (2-13) 

where 

MTA = total mass of radionuclide A in 

the sand pad 

MSA = mass sorbed onto solid phase 

MLA = mass sorbed into liquid phase. 

The partitioning between the solid and liquid 

phases is described according to the following 

equation. 
2-71 

LA

SA

dA C

C K = (2-14) 

where 

KdA = partition coefficient of 

radionuclide A (mL/g) 

CSA = solid concentration (g A/g sand) 

CLA = liquid concentration (g A/mL 

liquid). 

The solid concentration and liquid 

concentration of radionuclide A are defined as 

S

SA

SA M

M

C = (2-15) 

where 

MSA = mass of radionuclide A in sand 

MS = mass of sand 

and 

L

LA

LA V

M C = (2-16) 

where 

MLA = mass of radionuclide A in liquid 

VL = volume of liquid in sand void 

space. 

Substituting Equations (2-15) and (2-16) into 

Equation (2-14) and simultaneously solving for 

the two unknown variables (i.e., MSA and MLA) in 

the resulting equation and in Equation (2-13), 

yields (2-17). 

















+

=

L

dA S

TA

LA

V

K M

M M

1

 (2-17) 

Substituting this equation into Equation (2-13) 

and solving for MSA yields 

TA

L

dA S

SA M

V

K M

M













































+

= −

1

1 1 (2-18) 

After flushing the sand pad, the amount of 

radionuclide A removed is (1−Θres)MLA, where Θres

is the residual saturation of the sand pad. The 

amount of radionuclide A remaining in the sand 

pad is MSA+(Θres)MLA. This remaining amount is 

then decayed. The radioactive decay and flushing 

is repeated annually 38 times to determine the 

present-day activity of radionuclide A. This 

activity is decayed 16 yr to calculate activities in 

2016. The cycle is repeated for each radionuclide. 

2.3.2.3 Data and Assumptions. The 

accidental spill in 1962 is modeled assuming the 

liquid is present in the vault 24 hr. The 24-hr time 

is consistent with the estimated time from the 

beginning of the spill to the time when the spilled 

liquid was removed by pumping the vault 

(Latchum et al. 1962). The area of the flux surface 

(i.e., the liquid/sand pad interface) is assumed to 

be an annular region having a radius equal to the 

tank radius and a thickness equal to the maximum 

sand pad height. That is, this area is 

approximately equal to the area of a rectangle 

with a length of 2πr, where r is the radius of the 

tank (25 ft), and a width of 6 in. (i.e., the 

maximum estimated height of the sand pad). 

The initial amount of each radionuclide in the 

tanks at the time of the accidental spills was 

evaluated with limited sampling of Tanks 

WM-185 and WM-187 on February 14, 1962. A 

sample also was obtained from each vault during 

the siphoning; however, limited analyses were 

conducted. Due to the limited number of 

radionuclides provided by the analysis, an 

alternative method was used to determine the 

radionuclide inventory. An ORIGEN2 simulation 

of a pseudo aluminum-clad fuel with an initial 

235U enrichment of 93% and a burnup of the 

processed fuel of 18%, Wenzel (1997) evaluated 

the expected radionuclide content of the tank. The 

fuel from Materials Test Reactor (MTR) Cycle 

No. 198 (Dykes 1963) was taken as typical for the 

fuel processed. Tanks WM-185 and WM-187 only 

contained aluminum waste prior to the incident in 
2-72 

1962 (Palmer et al. 1999). The reactor contained 

4,842 g 235U and had 684 MWd of operation over 

a 417-day period. For calculation purposes, 

inventories were normalized to the activity in a 

typical 200-g element. The 1972 analysis using 

the ORIGEN2 code is appropriate because the 

waste source was the same type of MTR fuel that 

was present in Tanks WM-185 and WM-187 in 

1962 when the siphoning event occurred. The 

ORIGEN2 analysis is the best estimate of the 

1962 inventory. The analysis did not account for 

the decay of radionuclides. The ORIGEN2 data 

were corrected to the concentration of 137Cs in 

Tank WM-185 one month before the incident. 

Data for Tanks WM-185 and WM-187 were 

shown in the record of the incident (Latchum et al. 

1962). Tank WM-185 data were used because of 

the slightly higher 137Cs concentration. The 137Cs 

concentration used was 1.71 Ci/L. Appendix A 

presents the initial (i.e., 1962) inventories of the 

liquid in Ci/mL for all radionuclides and 

associated half-lives. Appendix A also gives the 

data used for decay chain calculations and the 

partitioning coefficients used in the analysis. 

The effective diffusion coefficient for each 

radionuclide from the liquid to the sand pad is set 

conservatively high for a liquid diffusion 

coefficient (1.0E−05 cm2

/s). The sand pad void 

fraction and residual saturation are 0.34 and 0.30, 

respectively (typical values for sand) (Thorne et 

al. 2000). Using the dimensions of the tank and 

sand pad and a typical density of sand 

(1.75 g/cm3

), the total mass of sand (Ms) is about 

4.14E+07 g. Additionally, the volume of water in 

the sand pad (VL) can be computed (i.e., the 

product of the volume of sand and the void 

fraction). 

Using the input parameters described in this 

section, Equation (2-6) is used to compute the 

mass of each radionuclide entering the sand pad in 

the 24-hr period the liquid was in the vault. With 

this inventory, radioactive decay is modeled for 

one year and the sand pad is flushed. Flushing 

represents the periodic pumping of liquid that 

collected inside the vault, and is assumed to occur 

once every year for the 38 yr from 1962 to 2000. 

In flushing, the sand is assumed to be 

saturated and the amount of each radionuclide in 

the liquid and solid phases is assumed to be 

partitioned and in equilibrium according to 

Equation (2-14). Liquid is assumed to be removed 

from the void space in the sand until the degree of 

saturation changes from saturated conditions to 

the residual saturation. The amount of each 

radionuclide remaining in the sand and residual 

liquid is then calculated. This mass is decayed for 

another year, at which time the sand pad is 

assumed to be flushed again. The decay and 

flushing cycle is repeated 38 times to determine 

the activities of each radionuclide in 2000 and the 

inventory is decayed 16 yr to estimate the 

activities at the time of closure in 2016. 

These assumptions and the data used in the 

analysis are considered reasonably conservative. 

That is, available data that represent the modeled 

system are used; however, when data are not 

available, conservative assumptions are made. For 

example, the surface available for diffusive 

transport in this analysis is assumed to 

instantaneously remove all radionuclides that 

reach the surface (i.e., the liquid/sand pad 

interface). Thus, mass transfer is greater under 

these conditions than if a concentration gradient is 

assumed to exist in the sand pad. 

Using conservative assumptions and the 

available data, the analysis is expected to provide 

a reasonable and conservative estimate of the 

mass of each radionuclide contained in the sand 

pad. That is, the actual sand pad inventories are 

not expected to be larger than these estimates. 

Table 2-17 lists a summary of results from the 

diffusion, radioactive decay, and flushing 

modeling. 

2.3.3 Piping Inventory 

Analysis of residual metal contamination in 

process waste piping has recently been performed. 

The radionuclide inventory for piping has been 

based on the metal contamination data collected in 

2002. The calculations determined a total 

ancillary piping residual mass of 15.5 kg. The 

estimate of residual radionuclide inventory in 

piping is described in Appendix A. The estimate 

used for the compliance scenario is based on 

analytical data.
2-73

Table 2-17. Inventory for the sand pads at the time of closure (2016). 

Nuclide Ci Nuclide Ci Nuclide Ci Nuclide Ci 

225Ac 2.23E−08 154Eu 1.29E+00 214Po 1.13E−05 228Th 1.93E−05 

227Ac 1.40E−05 221Fr 2.23E−08 215Po 1.40E−05 229Th 2.23E−08 

228Ac 2.33E−01 223Fr 1.93E−07 216Po 2.03E−05 230Th 5.49E−04 

108Ag 2.86E−09 152Gd 3.23E−14 218Po 1.13E−05 231Th 8.41E−03 

108mAg 2.86E−09 3

H 2.47E−22 236Pu 4.23E−11 232Th 4.44E−12 

241Am 1.89E+00 166mHo 4.15E−07 238Pu 2.00E+00 234Th 8.99E−05 

242Am 1.64E−05 129I 1.08E−06 239Pu 1.57E+00 207Tl 1.40E−05 

242mAm 1.64E−05 115In 3.71E−12 240Pu 3.54E−01 208Tl 7.33E−06 

243Am 1.47E−04 138La 5.33E−11 241Pu 1.88E+00 209Tl 4.82E−10 

217At 2.23E−08 93mNb 1.98E−01 242Pu 5.69E−05 210Tl 1.13E−05 

137mBa 1.50E+03 94Nb 2.29E−02 243Pu 4.11E−16 232U 1.88E−05 

10Be 1.80E−07 63Ni 1.64E−10 244Pu 5.99E−12 233U 1.35E−06 

210Bi 5.74E−06 235Np 1.79E−24 223Ra 1.40E−05 234U 3.13E−01 

211Bi 1.40E−05 237Np 3.72E−04 224Ra 2.03E−05 235U 8.41E−03 

212Bi 2.03E−05 238Np 7.77E−08 225Ra 2.23E−08 236U 1.06E−02 

213Bi 2.23E−08 239Np 1.47E−04 226Ra 1.13E−05 237U 3.75E−05 

214Bi 1.13E−05 240Np 6.59E−15 228Ra 2.33E−01 238U 8.99E−05 

14C 3.90E−07 240mNp 5.99E−12 87Rb 7.69E−07 240U 5.99E−12 

113mCd 1.60E−02 231Pa 2.02E−05 219Rn 1.40E−05 90Y 2.26E+02 

249Cf 2.42E−16 233Pa 3.72E−04 220Rn 2.03E−05 93Zr 2.11E−01 

250Cf 1.37E−17 234Pa 1.44E−07 222Rn 1.13E−05 

251Cf 5.95E−19 234mPa 8.99E−05 126Sb 1.65E−02 

252Cf 2.79E−25 209Pb 2.23E−08 126mSb 1.65E−02 

242Cm 1.35E−05 210Pb 5.74E−06 79Se 2.00E−02 

243Cm 7.91E−07 211Pb 1.40E−05 146Sm 3.77E−10 

244Cm 2.21E−04 212Pb 2.03E−05 147Sm 2.53E−06 

245Cm 5.11E−08 214Pb 1.13E−05 151Sm 3.76E+01 

246Cm 1.14E−09 107Pd 4.21E−04 121mSn 2.06E−03 

247Cm 4.11E−16 146Pm 1.71E−05 126Sn 1.65E−02 

248Cm 1.26E−16 147Pm 3.23E−03 90Sr 2.26E+02 

60Co 1.60E−03 210Po 5.74E−06 98Tc 6.18E−20 

135Cs 5.99E−03 211Po 3.83E−08 99Tc 2.02E−12 

137Cs 1.50E+03 212Po 1.30E−05 123Te 4.93E−16 

152Eu 9.28E−03 213Po 2.19E−08 227Th 1.40E−05 
2-74

2.3.4 Residual Radionuclide Inventory 

in Piping 

The Sampling and Analysis Plan for the PostDecontamination Characterization of the Process 

Waste Lines from INTEC Tank Farm Facility 

Tanks WM-182 and WM-183 (Portage 

Environmental 2001) was prepared to define the 

data collection steps for residual contamination in 

piping. The process waste lines in the TFF have 

carried acidic waste in solution and have routinely 

been flushed after waste transfer with either acid 

or acid plus a water flush. During closure of the 

tank systems, the piping was triple rinsed with 

water to remove loose residual waste. 

Sections of horizontal and vertical process 

waste line have been removed from Tank 

WM-182. Samples from the decontaminated 

process waste lines were collected; these data are 

used to represent the effectiveness of triple rinsing 

all of the lines remaining in the WM-182 and 

WM-183 tank systems. The piping was removed 

from the system and 2- to 6-in.-long sections of 

the pipe were removed from each end using a 

wheel pipe cutter. Rinsate samples from the pipes 

were collected and analyzed for metals. The 

piping was filled with water (sealed at one end 

with an inert cap), allowed to equilibrate for a 

minimum of 30 min, decanted, and analyzed for 

total metals. 

The analytical results for the metals are shown 

in Appendix A. The concentration of metals in 

liquid (mg/L) sodium-bearing waste is greater 

than the concentration of radionuclides (mg/L). 

Therefore, the residual metal concentration in 

piping is a conservative estimate of residual 

radionuclide concentration. The maximum 

concentration of each metal in the piping rinsate 

samples was summed with the remaining 23 data 

points. This yielded a concentration of 

2,922 μg/L. The rinsate was collected from 18-in. 

lengths of the 2.5-in. outside diameter pipe. 

Several assumptions were made to ensure the 

estimate of the piping residuals is conservative: 

• The mass of metals was assumed to be equal 

to 1 ft of piping. Therefore, the starting value 

was 1/3 greater than indicated by the analysis. 

• The sample volumes were less than 1 L (less 

than 300 mL). The data were not adjusted 

downward to correspond with the actual 

sample volumes. 

• A safety factor of 500 was applied to the data. 

This safety factor was used to provide a 

conservative estimate and to provide for the 

possibility of greater concentrations being 

found in other piping. 

Below are the equations and results that form 

a basis for this approach. The comparability of 

metals data to radionuclide data on a mass per unit 

volume was established by using the specific 

activity for a few radionuclides that are significant 

contributors to the inventory. Equation (2-19) 

shows the calculation for determining grams of 

137Cs. 

Csg = Ci / SpA (2-19) 

where 

Csg = 

137Cs grams 

Ci = curies of 137Cs in the liquid 

inventory (1142.5 Ci) 

SpA = 

137Cs specific activity (87 Ci/g). 

Equation (2-20) shows the calculation to 

determine the mass per unit volume. There is 

4,989 L of liquid in each tank. 

g / L = 13.13g / 4989L (2-20) 

This equation yields a concentration of 

3 mg/L. The same equation can be used to 

determine the concentration in mg/L of the 

remaining radionuclides. The mass of metals in 

the tank is found to be greater than the mass of 

radionuclides. The total concentration of 

radionuclides is approximately 73 mg/L, while the 

mass of metals is near 15 g/L. 

It is not assumed that the chemical properties 

of radionuclides and metals are the same; 

therefore, the conservative assumptions shown 

above were used to ensure the decontamination 
2-75 

factor for metals was not overstated and the fixed 

contamination would be included in the estimate. 

The following equations show the estimated mass 

of radionuclides in the piping residual based on 

the conservative inventory. 

C g L

i ∑ max = 2922μ / (2-21) 

where 

Cmaxi = maximum concentration of 

each metal in the rinsate 

samples. 

2,922 μg is assumed to correspond with the 

mass of metals in 18 in. of process waste pipe. 

Some waste piping is up to 3 in. outside diameter. 

To arrive at a conservative estimate, the mass is 

assumed to correspond to 1 ft of piping. 10,600 

linear ft of process waste piping is located in the 

TFF. 

mp = m × ftl × k (2-22) 

where 

mp = total mass in piping

m = mass per linear ft (2,922 μg) 

ftl = total linear ft of pipe (10,600) 

k = safety factor of 500. 

This equation yields 15.5 kg of sodiumbearing waste solid in the residual piping. The 

sodium-bearing waste in the conservative 

inventory is then apportioned to mass. Table 2-18 

shows the total Ci by radionuclide for the piping. 

Previous estimates were based on work 

contained in INTEC Tank Farm Facility Process 

Waste Line Source Term Estimate (Wilcox 2000). 

Wilcox estimated the residual contamination 

contained in the pipes based on a 0.002-in. film 

thickness coating the internal surface of process 

waste piping, not including vessel off-gas lines. 

Wilcox conducted conversations with INTEC 

Operations personnel in which they estimated the 

piping residue thickness within the C2 valve box 

piping to be about 2 mil (0.002 in.) on the bottom 

portion of the piping. For conservative reasons, 

the entire internal piping surface was given a 

2-mil thickness. The inventory that is predicted 

for the piping is approximately one-fourth of the 

estimate in Wilcox. The inventories of both 

estimates are similar except for the added 

conservative assumption to include a 2-mil 

thickness of the internal pipe surface. 
2-76

Table 2-18. Radionuclide inventory for piping at 2016. 

Nuclide Ci Nuclide Ci Nuclide Ci Nuclide Ci 

225Ac 2.6E−10 137Cs 2.6E+01 210Po 3.8E−09 148Sm 2.6E−14 

227Ac 3.8E−08 150Eu 7.7E−09 212Po 8.2E−07 149Sm 2.3E−15 

228Ac 4.6E−13 152Eu 9.0E−04 213Po 2.6E−10 151Sm 2.1E−01 

108mAg 2.8E−10 154Eu 2.2E−03 214Po 9.0E−09 121mSn 3.3E−05 

241Am 3.3E−03 155Eu 1.8E−02 215Po 3.8E−08 126Sn 2.8E−04 

242Am 1.0E−05 221Fr 2.6E−10 216Po 1.3E−06 90Sr 5.6E+01 

242mAm 1.0E−05 223Fr 5.4E−10 218Po 9.0E−09 98Tc 1.8E−09 

243Am 1.4E−05 152Gd 1.0E−15 144Pr 4.1E−09 99Tc 7.0E−02 

217At 2.6E−10 3

H 3.6E−03 144mPr 4.9E−11 123Te 2.6E−16 

137mBa 2.6E+01 166mHo 3.1E−08 236Pu 7.7E−08 125mTe 8.4E−05 

10Be 2.0E−09 129I 3.6E−05 238Pu 1.0E−01 227Th 3.8E−08 

210Bi 3.8E−09 115In 6.6E−14 239Pu 6.7E−03 228Th 1.3E−06 

210mB 1.5E−22 138La 1.3E−13 240Pu 7.2E−03 229Th 2.6E−10 

211Bi 3.8E−08 93mNb 1.3E−03 241Pu 9.5E−02 230Th 6.1E−07 

212Bi 1.3E−06 94Nb 5.7E−02 242Pu 5.4E−06 231Th 1.4E−05 

213Bi 2.6E−10 144Nd 1.1E−12 244Pu 4.6E−13 232Th 4.9E−13 

214Bi 9.0E−09 59Ni 1.2E−03 223Ra 3.8E−08 234Th 1.4E−05 

14C 8.2E−08 63Ni 1.9E−02 224Ra 1.3E−06 207Tl 3.8E−08 

113mCd 1.2E−03 237Np 4.4E−05 225Ra 2.6E−10 208Tl 4.6E−07 

142Ce 2.0E−08 238Np 4.9E−08 226Ra 9.0E−09 209Tl 5.6E−12 

144Ce 4.1E−09 239Np 1.4E−05 228Ra 4.6E−13 171Tm 3.1E−15 

249Cf 1.2E−14 240mNp 4.6E−13 87Rb 2.0E−08 232U 1.3E−06 

250Cf 4.9E−15 233Pa 2.0E−03 102Rh 2.6E−08 233U 1.7E−07 

251Cf 1.8E−16 231Pa 6.6E−08 106Rh 8.2E−08 234U 5.6E−04 

242Cm 8.2E−06 234Pa 1.8E−08 219Rn 3.8E−08 235U 3.1E−06 

243Cm 1.4E−05 234mPa 1.4E−05 220Rn 1.3E−06 236U 2.3E−05 

244Cm 7.2E−04 209Pb 2.6E−10 222Rn 9.0E−09 237U 2.4E−06 

245Cm 2.0E−07 210Pb 3.8E−09 106Ru 8.2E−08 238U 1.8E−06 

246Cm 1.3E−08 211Pb 3.8E−08 125Sb 3.3E−04 240U 4.6E−13 

247Cm 1.5E−14 212Pb 1.3E−06 126Sb 3.8E−05 90Y 5.6E+01 

248Cm 1.6E−14 214Pb 9.0E−09 126mSb 2.8E−04 93Zr 1.5E−03 

60Co 1.0E−03 107Pd 1.1E−05 79Se 3.1E−04 

134Cs 3.5E−04 146Pm 6.6E−06 146Sm 1.9E−10 

135Cs 6.1E−04 147Pm 3.8E−03 147Sm 5.1E−09 
3-1 

3. ANALYSIS OF PERFORMANCE 

The methods used to analyze the long-term 

performance of the TFF are described in this 

section. Figure 3-1 illustrates the steps taken in the 

analysis of performance with the final point 

comparing the dose to the performance objectives. 

Five main models were selected and used in 

the analysis of performance. These models are 

1. Groundwater flow/transport models (see 

Section 3.3.2) 

2. Groundwater release model (see 

Section 3.3.3) 

3. Radon flux model (see Section 3.3.4) 

4. Atmospheric transport (see Section 3.3.5) 

5. Intruder (see Section 5). 

Source term development is discussed in 

Section 3.1. This section includes a description of 

the inventories assumed, potential mechanisms of 

contaminant release from the facility, and potential 

mechanisms responsible for loss of integrity of the 

engineered TFF barriers. 

Following the source term discussion, 

potential receptors are identified in Section 3.2 by 

recognizing the time periods of concern in this PA, 

the potentially significant pathways to human 

exposure, and exposure scenarios. In this section, 

the TFF radionuclide inventory for the 

groundwater pathway is screened to eliminate 

radionuclides that, under conservative conditions, 

are insignificant with respect to potential human 

exposures. 

The conceptual models developed and the 

computational approach used to assess the 

performance of the TFF also are described in 

Section 3.2. The conceptual models are derived 

from technical information presented in Section 2. 

These models embody a number of simplifying 

assumptions to facilitate the computational 

analysis required to assess the long-term 

performance of the TFF. The dose calculations are 

presented in Section 3.3.5. 

3.1 Source Term 

This section discusses factors affecting the 

rate at which radionuclides are released from the 

disposal facility. Source term considerations are 

typically a large source of uncertainty. The 

uncertainty includes prediction of future 

conditions of the tanks, vaults, and piping 

contamination after the closure sequence is 

completed. The uncertainty is compounded by 

issues related to the release mechanisms from the 

waste form. Once releases from the waste form 

occur, the effectiveness and longevity of the 

concrete/grout system must be considered. Each of 

these topics is discussed in the following sections. 

3.1.1 Radionuclide Inventory 

The inventory used for the analysis of 

compliance (i.e., conservative case inventory) for 

a given 300,000-gal tank presented in Section 2.3 

was applied to each individual tank. The 

development of the conservative case is 

documented in Appendix A and is provided in 

Section 2.3. The inventory is provided for the year 

2016, which is the assumed year of completion of 

closure activities at the TFF. Two of the tanks 

contain contaminated sand pads located beneath 

the tanks (Tanks WM-185 and WM-187). The 

sand pad activities also are discussed in 

Section 2.3. In addition, the sand pad 

contamination was evaluated for the effect of 

rinsing from water infiltration into the vaults and 

drainage that occurred from 1962 to 2000. This 

analysis is provided in Appendix A. 

Contamination in the piping also was considered. 

The piping inventory calculations also are 

provided in Appendix A. 

The estimated residual inventory for the 

INTEC TFF tanks contains several decay chains. 

Decay chains are important because radionuclides 

with low inventories at the time of closure may 

increase over the time of interest because of the 

decay of their parent radionuclide. The inventory 

can be divided between radionuclides that are not 

part of a decay chain, short decay chains (two to 

three radionuclides), and large decay chains. The 
3-2 

Figure 3-1. Flow chart for the PA analyses. 
3-3 

The decay and ingrowth of short decay chain radionuclides are controlled by the following equations. 

These equations are based on a three-box model defined as 

93mNb 93Zr 125mTe 125Sb 

137mBa 137Cs 

147Sm 147Pm 

152Eu 152Gd 

106Rh 106Ru 

108Ag 108mAg 

109mAg 109Cd 

126Sn 126mSb 126Sb 

144Ce 144mPr 144Pr 

90Sr 90Y 110mAg 110Ag 

q1 q2 q3
3-4 

following decay chains were identified in the TFF 

tank inventories: 

The differential equations that define the rate 

change of atoms in each compartment (i.e., for 

each radionuclide in the decay chain) are as 

follows 

1 1

1 k q dt

dq = − (3-1) 

1 1 2 2

2 k q k q dt

dq = − (3-2) 

2 2 3 3

3 k q k q dt

dq = − (3-3) 

where 

q = atoms of radionuclide of interest 

t = time (yr) 

k = decay constant for radionuclide of 

interest (yr-1). 

Solutions to the above differential equations 

were obtained by Laplace transforms as follows: 

k t q t q e 1 ( ) (0) 1 1

− = (3-4) 

( ) k t k t k t e e q e

k k

k q q t 2 1 2 (0) (0) ( ) 2

1 2

1 1

2

− − − − + − = (3-5) 

( )( ) ( )( )

( )( ) ( )( ) k t k t k t k t

k t k t

e e q ( )e k k

k q ( ) e k k k k

k k q ( )

e k k k k

k k q ( ) e k k k k

k k q ( ) q (t)

3 2 3 3

1 2

0 0 0

0 0

3

3 2

2 2

1 3 2 3

1 2 1

1 2 3 2

1 2 1

2 1 3 1

1 2 1

3

− − − −

− −

− + − + − − +

− − + − − =

 (3-6) 

The activity of each compartment may be 

determined from these equations by simply 

multiplying by the appropriate decay constant. The 

activity for individual radionuclides that were not 

part of a decay chain was determined according to 

Equation (3-4). 

Four major decay chains were identified in the 

TFF tank residual inventories: 252Cf, 249Cf, 251Cf, 

and 250Cf. These decay chains are shown in 

Figures 3-2 through 3-5. Radionuclides in italics 

are short-lived progeny that are considered to be in 

equilibrium with their parent radionuclide. The 

radionuclide activities as a function of time for the 

analyses were determined using the RadDecay 

computer code (Grove Engineering 1990). 

3.1.2 Engineered Barriers Degradation 

and Failure 

The following sections address degradation of 

the cover and the TFF closure system (i.e., vault, 

tank, grout, and piping). The degradation analyses 

were conducted to provide support for the PA 

analyses that assume degradation step changes of 

100 yr for the vault and 500 yr for the tanks and 

piping. 

3.1.2.1 Cover Degradation. As stated in 

Section 2.2.1, future actions in the TFF to 

remediate contaminated soils will be conducted 

under CERCLA. As such, the future cover design 

is not known at this time. Therefore, a bounding 

analysis was conducted for the TFF PA (i.e., 

compliance analysis based on the conservative 

case parameters) and no additional cover was 

assumed to be available. The overlying 10 ft (3 m) 

of cover soil was assumed to be the only cover 

available for the TFF analyses. The erosion of the 

overlying 10 ft (3 m) of soil also was assumed to 

not take place. In this way, the analysis 

conservative case does not become a “worst-case” 

analysis. Since the TFF lies within an area of 

relatively flat terrain and a future cover will be 

placed over the facility, the assumption of no 

erosion of the existing 10 ft (3 m) of soil cover is 

considered appropriate.

3.1.2.2 Degradation of the Vault, Grout, 

Tank, and Piping. The degradation analysis of 

the TFF closure system (i.e., vaults, tanks, grout, 

and piping) is described in detail in Appendix E. 

The following discussion provides an overview of 

the degradation analysis and the assumptions 

incorporated into the TFF PA analyses.

Potential degradation mechanisms for the 

concrete, grout, piping and tanks, and the 

preferential water pathways were assessed for the 

TFF PA. The potential degradation mechanisms 
3-5 

Figure 3-2. 252Cf radionuclide decay chain. 

252Cf 

248Cm 

244Pu 

240U 

240mNp 

240Pu 

228Ra 

236U 

232Th 

224Ra 

228Ac 

220Rn 

216Po 

212Pb 

212Po 

212Bi 

Stable 208Pb 

244Cm 

240Np 

232U 236Pu 

208Tl 

228Th

* Radionuclides in italics are short-lived 

progeny considered to be in equilibrium 

with the parent radionuclide. 
3-6 

Figure 3-3. 249Cf radionuclide decay chain. 

249Cf 

245Cm 

241Pu 

241Am 

237Np 

233Pa 

225Ra 

233U 

229Th 

217

At 

221Fr 

225Ac 

213Bi 

213Po 

209Pb 

Stable 209Bi 

237

U 

209Tl 

* Radionuclides in italics are short-lived 

progeny considered to be in equilibrium 

with the parent radionuclide. 

249Cf 

245Cm 

241Pu 

241Am 

237Np 

233Pa 

225Ra 

233U 

229Th 

217

At 

221Fr 

225Ac 

213Bi 

213Po 

209Pb 

Stable 209Bi 

237

U 

209Tl 

* Radionuclides in italics are short-lived 

progeny considered to be in equilibrium 

with the parent radionuclide. 
3-7 

Figure 3-4. 251Cf radionuclide decay chain. 

251Cf 

247Cm 

243Pu 

243Am 

239Np 

239Pu 

231Pa 

235U 

231Th 

223Ra 

227

Th 

227Ac 

219Rn 

215Po 

211Pb 

211Po 

211Bi 

Stable 207Pb 

235Np 

215At 

223Fr 

207

Tl 

243Cm 

* Radionuclides in italics are short-lived 

progeny considered to be in equilibrium 

with the parent radionuclide. 

251Cf 

247Cm 

243Pu 

243Am 

239Np 

239Pu 

231Pa 

235U 

231Th 

223Ra 

227

Th 

227Ac 

219Rn 

215Po 

211Pb 

211Po 

211Bi 

Stable 207Pb 

235Np 

215At 

223Fr 

207

Tl 

243Cm 

* Radionuclides in italics are short-lived 

progeny considered to be in equilibrium 

with the parent radionuclide. 
3-8 

Figure 3-5. 250Cf radionuclide decay chain. 

250Cf 

246Cm 

242Pu 

238U 

234Th 

234mPa 

226Ra 

234U 

230Th 

214Pb 

218Po 

222Rn 

214Bi 

214Po 

210Pb 

210Po 

210Bi 

Stable 206Pb 

238Np 

234Pa 

218At 

210Tl 

206Tl 

238Pu 

242Cm 

242mAm 

242Am 

238Pu 

* Radionuclides in italics are short-lived 

progeny considered to be in equilibrium 

with the parent radionuclide. 

250Cf 

246Cm 

242Pu 

238U 

234Th 

234mPa 

226Ra 

234U 

230Th 

214Pb 

218Po 

222Rn 

214Bi 

214Po 

210Pb 

210Po 

210Bi 

Stable 206Pb 

238Np 

234Pa 

218At 

210Tl 

206Tl 

238Pu 

242Cm 

242mAm 

242Am 

238Pu 

* Radionuclides in italics are short-lived 

progeny considered to be in equilibrium 

with the parent radionuclide. 
3-9

and factors that can affect the permeability include 

cracks and voids, sulfate and magnesium attack, 

calcium hydroxide leaching, alkali-aggregate 

reaction, carbonation, acid attack, and corrosion of 

the tank, pipes, and reinforcement. This section 

provides a summary of these degradation 

mechanisms. A complete discussion of the 

approach used to model these mechanisms is 

presented in Appendix E. 

Because of the limited availability of sitespecific data (e.g., properties of the concrete and 

grout, chemistry of the soil moisture and water 

entering the vault), a number of assumptions were 

made and simple models used. The assumptions 

rely on using values for the conservative case 

model input parameters that are viewed as 

reasonably conservative (i.e., expected values). To 

compare the sensitivity of the model results to the 

expected values, a sensitivity analysis was 

performed and is presented in Appendix E. The 

sensitivity analysis evaluates the models using 

input parameters with the maximum protection 

and minimum protection parameter values. 

The analysis modeled sulfate and magnesium 

attack, carbonation, and calcium hydroxide 

leaching. Reinforcement corrosion of the outer 

vault and localized corrosion that results in 

complete corrosion of the tank also were modeled. 

Based on the expected conditions at the TFF 

following closure, effects of acid attack, alkaliaggregate reaction, and corrosion of the pipes on 

the concrete and grout degradation are assumed to 

be insignificant compared to the three modeled 

chemical attacks. This approach is discussed in 

Appendix E. 

In the closure of the TFF tanks, the tanks, 

vaults, and piping leading to the tanks will be 

filled with grout (INEEL 1999a). For modeling 

purposes, each of the 11 tanks and vaults and the 

piping external to the vaults is conceptualized as a 

representative tank, vault, or piping having 

average dimensions. Thus, because only a 

representative system is modeled, the same 

permeabilities are computed for each of the 11 

tanks, vaults, and piping systems. Considering the 

limited data and the lack of information regarding 

variabilities between the tanks, it is reasonable to 

model a single tank, vault, and piping that 

represents each of the 11 tanks, vaults, and piping 

systems. 

Figure 3-6 shows the system conceptualization 

and the grout degradation sequence of the 

representative vault, tank, grout, and piping. This 

sequence was used as the basis to develop the 

approach and mathematical models for the 

degradation analysis. 

Figure 3-6a shows the conceptualized 

individual tank/vault unit. The concrete and grout 

are not cracked, all voids are filled, and the tank 

has no penetrations. 

Figure 3-6b displays a tank/vault unit 

following the grout pour. A few shrinkage cracks 

are present in the outer vault and grout (Odler and 

Jawed 1991). A small number of localized 

penetrations have occurred in the tank wall. These 

penetrations could occur from pitting corrosion, 

galvanic corrosion related to pipe entrances, or 

stress corrosion cracking. In order to take a 

conservative approach to degradation, this figure 

is used as the initial condition assumed in the 

failure calculations. Multiple studies indicate that 

none of these initial failures are likely to be 

present in any of the vault/tank units now. 

In Figure 3-6c, the steel reinforcement in the 

outer concrete vault is assumed to corrode. As the 

steel corrodes, it expands, leading to disruption of 

the concrete. The gravel texture in the figures is 

intended to represent a rubble condition where the 

concrete has lost its structural properties. Even 

after structural failure, the concrete still acts as a 

chemical barrier that conditions water and air 

reaching the outer grout layer. This condition 

occurs at about 500 yr. 

Figure 3-6d depicts the vault/tank unit after 

the grout outside the tank has failed from chemical 

attack. Chemical attack must penetrate the outer 

vault rubble before initiating attack on the grout. 

No steel reinforcement is present in the outer 

grout. Thus, steel corrosion is eliminated as a 

failure mechanism. This condition occurs at about 

5,000 yr. 

Figure 3-6e illustrates the final stage where the 

tank has corroded substantially and all concrete 
3-10

Figure 3-6. Conceptualization of the degradation sequence for the representative tank, grout, vault, and 

piping system at the TFF. 

and grout have turned to rubble. At this point, the 

vault/tank unit has lost all hydraulic control of 

contaminant leaching. The chemically-reducing 

grout formulation may slow the migration of many 

radionuclides, but no credit is taken for this effect 

in the degradation modeling. This condition occurs 

at about 40,000 yr. 

A representative vertical cross section of the 

tank, vault, and grout system and its dimensions 

are shown in Figure 3-7. The system 

conceptualization consists of four zones: 

• The tank and the grout inside the tank (Zone a) 

• The grout inside the vault but outside the tank 

(Zone b) 

• The outer concrete vault walls, roof, and floor 

(Zone c) 

• The piping (Zone d). 
3-11

Figure 3-7. Representative tank, vault, and grout showing the dimensions for the degradation analysis. 

Conceptually, because these zones contain 

potential water pathways arising from initial voids 

and cracks, each zone has a nonzero initial 

permeability (Seitz and Walton 1993). Over time, 

various mechanisms that increase the permeability 

cause the tank to corrode and the concrete and 

grout to degrade. The degradation mechanisms 

and processes considered in this analysis that can 

affect the system permeability include cracks and 

voids, sulfate and magnesium attack, calcium 

hydroxide leaching, alkali-aggregate reaction, 

carbonation, acid attack, and corrosion of the tank, 

pipes, and reinforcement. 

The likely degradation processes of the 

representative system are conceptualized as 

occurring at the outside of the vaults and moving 

inward to the center of the system (i.e., from 

Zone c to Zone b to Zone a). For Zones a, b, and c, 

this approach is analogous to peeling an onion. 

Once the outer concrete vault (Zone c) completely 

turns to rubble, the next layer (Zone b) begins to 

degrade. After the grout in Zone b turns to rubble, 

the grout in Zone a begins to degrade. 

Additionally, the permeability of Zone a is not 

only affected by the degradation of the grout 

inside the tank but also by general and localized 

corrosion of the tank. Because the piping is 

separate from Zones a, b, and c, Zone d is modeled 

independent of the other three zones. However, the 

grout in Zone d is subject to the same chemical 

attack as Zones a, b, and c. 

During the degradation period for each zone, 

the permeability increases from its initial 

permeability. Conceptually, the increasing 

permeability reflects an increase in the pathways 

available for water flow. Following complete 

degradation of the concrete vault and grout in 

Zones a through d, including complete corrosion 

of the tank and piping, all zones are assumed to 

have the same permeability (i.e., the permeability 

of rubble). 

All zones of the modeled system were 

assigned reasonable values for an initial 

permeability (Breysse and Gerard 1997; Walton 

and Seitz 1991). Water can flow through cement 

and grout matrix, shrinkage cracks, and voids. For 

example, although the autogenous healing in 
3-12

cracks is a well-known physical phenomenon 

(Edvardsen 1999; Hearn, Detwiler, and Sframeli 

1994), the concrete permeability is significantly 

affected by the number of cracks and the crack 

width (Wang, Jansen, and Shah 1997) and 

shrinkage cracks (Hearn 1999). The presence of 

cracks and voids is conservatively assumed in this 

analysis, even though the grouting operation is 

designed to ensure (a) good pours without voids 

around and within pipes, and (b) a minimum of 

shrinkage or thermal stress cracking. 

The reinforced concrete vaults either were 

poured in place or constructed using pillars and 

panels. Because of the difficulty of verifying 

initial conditions and predicting future conditions, 

very little performance credit is taken for the 

vaults. The initial permeability of the vaults 

(Zone c) is taken as 10-9 m/s. The fresh grout 

inside and outside of the tank is assumed to have 

an initial permeability of 10-13 m/s (Iriya et al. 

1992). 

In order to flow through the inside of the tank, 

water must pass through a flaw in the tank roof or 

side, flow through the grout inside the tank, then 

exit through a second hole in the tank. 

Permeability of the grouted tank (Zone a) is 

computed in this analysis as a lumped 

permeability that takes into account the status of 

the tank and the inner grout. Thus, the initial 

permeability for the grouted tank in Zone a is 

assumed to be 10-16 m/s. 

The models used in the degradation analysis of 

Zones a, b, and c for sulfate and magnesium 

attack, calcium hydroxide leaching, and 

carbonation are presented in detail in Appendix E. 

The results are shown in Figure 3-8. The 

figure indicates that the percent of concrete or 

grout lost from sulfate and magnesium attack 

occurs much faster than carbonation and calcium 

hydroxide leaching, with Zone c completely 

degrading after about 2,000 yr. Figure 3-9 shows 

the percent of grout lost for Zone d (piping) and 

reveals that the piping (assuming no credit for the 

pipe) completely degrades after 500 yr from 

sulfate and magnesium attack. The overall 

degradation for each zone is shown in Figure 3-10. 

This figure shows corrosion of the reinforcement 

steel fails the vault (Zone c) at about 500 yr. Note 

that the degradation of Zone c is faster by 

reinforcement corrosion than by sulfate attack in 

Figure 3-8. However, the time for failure of Zones 

a and b is the same in Figures 3-8 and 3-10. The 

reason for this consistency is the model assumes 

that although reinforcement corrosion can fail 

Zone c, chemical attack must still penetrate the 

cracked concrete vault before the grout in Zone b 

is affected. 

The hydraulic conductivities for Zones a, b, c, 

and d are displayed in Figures 3-11 and 3-12. 

These figures show the effects of chemical attack 

on all four zones, reinforcement corrosion on Zone 

c, and corrosion of the tank on Zone a. The results 

also are consistent with the percent of loss 

provided in Figures 3-8 through 3-10. 

As discussed in Appendix E, the permeability 

of Zone a is modeled assuming localized corrosion 

produces holes in the tank walls. The number of 

holes and the permeability of the material 

surrounding the holes (i.e., the grout in Zones a 

and b) determines the permeability of Zone a. The 

model used to predict the permeability of Zone a is 

based on the approach from Walton et al. (1997). 

Using this model, the permeability in Zone a 

for each chemical attack is computed. For the 

maximum degradation shown in Figure 3-8, the 

permeability in Zone a begins to increase because 

of the increased number of holes in the tank at 

about 2,000 yr. After Zone b turns to rubble at 

5,000 yr, the grout inside the tank degrades and 

the Zone a permeability begins to increase at a 

faster rate until about 10,000 yr. At about 

10,000 yr, the grout inside the tank has turned to 

rubble. The remaining tank walls continue to 

corrode at a decreased rate until the entire tank has 

corroded at about 40,000 yr, when the 

permeability of Zone a equals the permeability of 

rubble. This time corresponds to the time when 

localized corrosion causes enough holes in the 

tank walls to completely fail the tank. This time 

also is consistent with the time the tanks are 

expected to fail from general corrosion (as 

estimated using data for corrosion rates measured 

from the coupons retrieved from the TFF tanks) 

(Palmer et al. 1999). 
3-13

Figure 3-8. Base-case results for the percentage of 

concrete and grout degraded from sulfate and 

magnesium attack, carbonation, and calcium 

hydroxide leaching for Zones a (tank and grout), b 

(grout), and c (vault). 

Figure 3-9. Base-case results for percent grout loss 

for the maximum degradation from sulfate and 

magnesium attack, carbonation, and calcium 

hydroxide leaching for Zone d (piping). 

Figure 3-10. Base-case results for the maximum 

percent of concrete and grout degraded from 

reinforcement corrosion, sulfate and magnesium 

attack, carbonation, and calcium hydroxide 

leaching for Zones a (tank and grout), b (grout), 

and c (vault). 

Figure 3-11. Base-case results of the hydraulic 

conductivity of Zone a (tank and grout), Zone b 

(grout inside vault and outside tank), and Zone c 

(vault). 

 c

 b

 a

 c

b,a

 

 

 c 

 

 

 b,a
3-14

Figure 3-12. Base-case results of the hydraulic 

conductivity of Zone d (piping). 

The model results show reinforcement 

corrosion, which cracks the concrete, causes the 

outer vault to turn to rubble in an estimated 500 yr. 

Once the outer vault completely degrades, the 

grout between the vault and the tank turns to 

rubble at approximately 5,000 yr. The tank and 

grout contained in the tank completely degrade 

and turn to rubble at approximately 40,000 yr. The 

grout associated with the piping turns to rubble 

after approximately 500 yr. The predominant 

chemical attack on the grout is from sulfate and 

magnesium. Sensitivity analysis indicates the 

times for the four zones to turn to rubble can vary 

greatly from tens of years to tens of thousands of 

years or beyond (see Appendix E). 

In the above discussion, the grouted pipe and 

encasement system was assumed to degrade by 

sulfate and magnesium attack, carbonation, and 

calcium hydroxide leaching. The pipe was 

assumed to corrode instantaneously and have no 

affect on (a) the permeability of the grouted pipe 

and encasement system and (b) contaminant 

releases. To provide a more realistic representation 

of contaminant transport from the grouted pipe 

and encasement system into the vadose zone, a 

more detailed alternative analysis was conducted 

to include the effects of the pipe on contaminant 

releases. The analysis assumed that the pipe is 

3-in. schedule 40 stainless steel, which (according 

to ANSI B36.10 [1970]) has a wall thickness of 

0.216 in. Corrosion rates reported by Palmer et al. 

(1999) for coupons placed into the TFF stainlesssteel tanks were applied to the pipe degradation. 

This is reasonable since the assumption implies 

the water contacting the pipe has the same 

characteristics as the water contacting the tank. 

For consistency in calculations, the rate at which 

localized corrosion in the tanks results in the tank 

wall holes also occurs at the same rate in the pipes 

(i.e., holes in the tank and pipe walls are assumed 

to appear with the same frequency). 

Using the corrosion rates in Palmer et al. 

(1999), the pipe will completely corrode from 

general corrosion in about 40,000 yr. Additionally, 

the model for localized corrosion used in the 

degradation analysis predicts the first hole in the 

tank (which is assumed to occur at the same rate as 

in the pipes) appears at approximately 10,000 yr. 

The current PA analysis used the following 

step degradation times (i.e., complete failure) for 

the TFF closure system: (1) outer vault/grout fail 

at 100 yr post-closure, (2) tank and grout in the 

tank fails at 500 yr post-closure, and (3) piping 

fails at 500 yr post-closure. These assumptions are 

conservative based on the best estimates provided 

from the degradation analysis and, due to the 

uncertainty in the degradation results, are 

considered appropriate for the PA analyses. 

3.2 Pathways and Scenarios 

The following sections describe the time 

periods of concern and pathways of human 

exposure to TFF constituents potentially released 

after closure. The information provided in this 

section was subsequently used to develop models 

for evaluating potential doses. 

3.2.1 Time Periods of Concern 

For the purposes of assessing the performance 

of the TFF closure facility, two time periods were 

addressed. The time periods include the 

institutional control period and the postinstitutional control period. Each period of interest 

is described in the following sections. 

3.2.1.1 Institutional Control Period. The 

institutional control period is the 100-yr time 

interval specified in DOE Manual 435.1-1 (2001) 

following closure of the TFF. The assumed closure 

date of the TFF is currently 2016. Periodic 

maintenance and monitoring activities are 

conducted during the institutional control period. 

The TFF site is assumed to be stabilized and no 
3-15

longer operational during this period but will 

remain part of the INEEL and, therefore, fenced 

and patrolled to eliminate the possibility of 

inadvertent intruders. Institutional control is 

assumed to last 100 yr (i.e., until 2116) for the 

purposes of this PA.

3.2.1.2 Post-Institutional Control Period. 

The final time period of concern is when the 

facility is no longer maintained by the INEEL and 

could be accessed by the public. For the purpose 

of this PA, the total duration of this period is 

1,000 yr from the time of closure (i.e., 3016). 

Projections of conditions and activities during this 

period are uncertain and difficult to assess. In 

addition, peak doses may not occur from TFF 

releases until several thousand years after closure 

because of the presence of long-lived 

radionuclides. The potential peak doses from these 

long-lived radionuclides are in the sensitivity 

analysis. However, for compliance purposes, the 

performance period is 1,000 yr after facility 

closure. Reasonable human activities are assumed 

to occur over this period of time based on the 

existing population patterns for the INEEL region.

3.2.2 Transport Pathways 

This section identifies potential pathways of 

human exposure to radionuclides released from the 

TFF and justifies eliminating some of these 

pathways from further consideration. The results 

of this section were used to develop exposure 

scenarios for off-site members of the public. 

3.2.2.1 Pathway Identification. 

Radionuclides released from the TFF to the 

geosphere have the potential of reaching humans 

through numerous pathways. Most conceivable 

pathways for a buried LLW source are indicated in 

Figure 3-13, and each pathway is briefly defined. 

The pathways identified in this figure are for 

facilities undisturbed from the standpoint of 

human intrusion. Pathways pertinent to intruder 

exposures are addressed separately in Section 5. 

3.2.2.2 Pathway Screening. The pathways 

listed in Figure 3-13 are generic in nature, and the 

significance of each pathway must be evaluated on 

a site-specific basis to develop an exposure model. 

Many pathways may be removed from 

consideration for particular sites because of a 

negligible contribution to human exposure.

For the TFF, leaching and transport of 

radionuclides to the saturated zone [pathway (1)] 

is the predominant means that radionuclides may 

be subsequently transported in the environment. 

Thus, this pathway must be addressed in 

developing an exposure model and is addressed in 

the groundwater model of this PA. Other pathways 

that may contribute to human exposure are those 

tied to groundwater concentrations of 

contaminants. Irrigation with contaminated 

groundwater may lead to contamination of 

agricultural crops and animals [pathways (3), (21), 

(23), (25) and (26)]. Ingestion of contaminated 

groundwater by terrestrial animals [pathway (19)] 

may lead to human exposure, and can be tied to 

groundwater contamination. Human exposure may 

occur as a result of direct human ingestion of 

contaminated groundwater [pathway (41)] as a 

result of consumption of contaminated food 

supplies [pathways (44) and (45)]. These were all 

considered for the TFF PA exposure scenarios. 

Discharge of contaminated groundwater to 

surface water [pathway (6)] may result in 

contamination of the aquatic ecosystem including 

the water body itself, sediment, and aquatic plants 

and animals [pathways (13), (14), (15), (34), (35), 

(36), (37), (38), and (39)]. The groundwater at the 

INEEL does not intersect the Big Lost River in the 

vicinity of the INEEL. The groundwater at the 

INEEL does interact with surface water as springs 

that flow into the Snake River. Major areas of 

springs and seepages from the aquifer occur in the 

vicinity of the American Falls Reservoir 

(southwest of Pocatello), approximately 44 mi 

(71 km) from the TFF, and the Thousand Springs 

area (near Twin Falls) between Milner Dam and 

King Hill, approximately 100 mi (200 km) from 

the TFF (Garabedian 1986). These spring 

locations are at such great distances from the TFF 

that contributions to the human dose from surface 

water at these locations would be insignificant 

because of groundwater dilution. Therefore, these 

pathways are not considered further in the TFF 

PA. In addition, exposure pathways associated 

with the ingestion of surface water, swimming, 

aquatic animals, or associated plant/animal uptake 

of radionuclides from surface water [pathways 
3-16

Figure 3-13. Potential transport and exposure pathways from a near-surface facility (pathways selected for 

analysis are identified by italics). 
3-17

(1) Leaching—migration of radionuclides from the waste form 

by a combination of dissolution, diffusion, and advection. 

(2) Gaseous Diffusion—upward migration of gaseous 

radionuclides from the waste form by diffusion through the 

caps and cover soils to the atmosphere. 

(3) Irrigation—contamination of cover soil by radionuclides 

that have reached groundwater, which is subsequently used 

for irrigation. 

(4) Deposition—contamination of surface water by 

radionuclides that have reached the atmosphere; represents 

deposition of particulate-associated radionuclides or gaseous 

species partitioning at the air-water interface. 

(5) Volatilization—partitioning of volatile radionuclides species 

present in surface water into air above the water body. 

(6) Discharge—discharge of radionuclides present in 

groundwater into surface water. 

(7) Recharge—movement of radionuclides into the groundwater 

from contaminated surface water. 

(8) Irrigation—contamination of cover soil by radionuclides 

that have reached surface water, which is subsequently 

being used for irrigation. 

(9) Washload—contamination of surface water by soil 

containing radionuclides as a result of erosion by rain or 

irrigation water. 

(10) Deposition—contamination of cover soil by radionuclides 

that have reached the atmosphere and have become 

associated with airborne particulate matter. 

(11) Resuspension—Resuspension of soil-associated 

radionuclides as a result of wind erosion. 

(12) Biointrusion—contamination of cover soil by soilassociated radionuclides that are brought to the surface from 

the vicinity of the waste form by burrowing animals, such as 

rodents or ants, or by intruding plant roots. 

(13) Deposition—deposition of radionuclides in surface water 

that have partitioned onto suspended sediment. 

(14) Resuspension—resuspension of particulate-borne 

radionuclides in the sediment of surface water as a result of 

hydrodynamic forces at the sediment-water interface. 

(15) Immersion—contamination of aquatic plants by 

radionuclides in surface water attributable to the immersion 

of the plants in the contaminated water. 

(16) Immersion—human exposure to radionuclides as a result of 

immersion in contaminated surface water. 

(17) Ingestion—human exposure to radionuclides as a result of 

ingestion and inhalation of radionuclides present in surface 

water. 

(18) Ingestion—contamination of terrestrial animals from their 

ingestion of radionuclides in surface water. 

(19) Ingestion—contamination of terrestrial animals from their 

ingestion of radionuclides in groundwater. 

(20) Irrigation—contamination of terrestrial plants as a result of 

irrigation with surface water containing radionuclides. 

(21) Irrigation—contamination of terrestrial plants as a result of 

irrigation with groundwater. 

(22) Decomposition—contamination of cover soil as a result of 

decomposition of terrestrial plants in the soil. 

(23) Root uptake—contamination of terrestrial plants by uptake 

through roots of soil water containing radionuclides. 

(24) Deposition—deposition of airborne radionuclides onto 

terrestrial plant surfaces. 

(25) Ingestion—ingestion of radionuclides by grazing animals as 

a result of contaminated soil ingestion. 

(26) Ingestion—ingestion of radionuclide-containing vegetation 

by terrestrial animals. 

(27) Decomposition—contamination of cover soil as a result of 

decomposition of terrestrial animals in the soil. 

(28) Washoff—contamination of surface soil as a result of 

washoff of externally-contaminated terrestrial animals. 

(29) Resuspension—resuspension of surficial radionuclides on 

terrestrial animals to the atmosphere. 

(30) Resuspension—resuspension of surficial radionuclides on 

terrestrial plants to the atmosphere. 

(31) Inhalation—contamination of terrestrial animals as a result 

of inhalation of radionuclides in the atmosphere. 

(32) Deposition—surface contamination of terrestrial animals via 

deposition of particulate-borne radionuclides in the 

atmosphere. 

(33) Ingestion—contamination of terrestrial animals as a result of 

their ingestion of aquatic animals. 

(34) Decomposition—contamination of surface water sediment 

as a result of decomposition of aquatic plants in the 

sediment. 

(35) Decomposition—contamination of surface water sediment 

as a result of decomposition of aquatic animals in the 

sediment. 

(36) Surface contact—surface contamination of aquatic animals 

as a result of contact with contaminated sediment. 

(37) Root uptake—contamination of aquatic plants via 

radionuclide uptake through roots. 

(38) Immersion—contamination of aquatic animals as a result of 

immersion in surface water containing radionuclides. 

(39) Ingestion—contamination of aquatic animals as a result of 

their ingestion of aquatic plants containing radionuclides. 

(40) Ingestion—human exposure to radionuclides as a result of 

ingestion of contaminated aquatic flora. 

(41) Ingestion—human exposure to radionuclides as a result of 

ingestion of contaminated groundwater. 

(42) Inhalation—human exposure to radionuclides as a result of 

inhalation of airborne radionuclides. 

(43) Immersion—human exposure to radionuclides as a result of 

immersion in contaminated air. 

(44) Ingestion—human exposure to radionuclides as a result of 

ingestion of contaminated terrestrial animals and animal 

products (such as milk). 

(45) Ingestion—human exposure to radionuclides as a result of 

ingestion of contaminated terrestrial plants. 

(46) Ingestion—human exposure to radionuclides as a result of 

ingestion of contaminated aquatic animals. 

(47) Washoff—contamination of surface soil below vegetation 

due to rain-induced surface washoff. 
3-18

(16), (17), (18), (33), (40), and (46)] also were not 

considered for further analysis. 

Pathway (7) considers contamination of 

groundwater due to recharge by surface water. 

This pathway is not considered significant for the 

TFF since migration of radionuclides to the nearby 

surface water body (i.e., Big Lost River) is not 

considered possible. The waste will be in a 

grouted form buried beneath the ground surface 

such that the interaction of the waste with surface 

water is not physically possible. Pathways (8) and 

(20), representing contamination of cover soil and 

terrestrial plants as a result of irrigation with 

contaminated surface water, also not considered. 

They were withdrawn because of a lack of a 

physical mechanism to move the radionuclides 

from the source to the surface water. 

Contamination of surface water from erosion of 

contaminated soil [pathway (9)] was not 

considered significant because movement of the 

contaminants from their present location below the 

ground surface to the surface is not considered 

feasible [i.e., no erosion of 10-ft (3-m) cover soil]. 

In addition, pathways (22) and (27), representing 

the pathways to surface soil via decomposition of 

terrestrial plants and animals, were not considered 

significant relative to the exposure resulting from 

direct consumption of these potentially 

contaminated products. 

Pathways that result in human exposure 

directly or indirectly as a result of atmospheric 

dispersion and deposition [pathways (4), (5), (10), 

(11), (24), (28), (29), (30), (31), (32), and (47)] are 

not included in exposure scenarios for the TFF PA 

for the following reasons: 

• The only potentially volatile radioactive 

components of the TFF are 3

H and 14C. 

Calculations providing an upper bound on 

doses received from volatilization of these 

radionuclides from the TFF facility after 

disposal are described in Section 3.3.5. 

• Pathways leading to exposure from 

resuspended contaminated soil [pathway (11)] 

are not considered since there is no 

mechanism for releasing particulates from the 

TFF (i.e., grouted waste form). 

• Other atmospheric pathways are indirect in 

nature (e.g., the contaminants must first be 

suspended or volatilized from one medium, 

then redeposited in another). These indirect 

pathways are not believed to be more 

significant than the direct pathways (2, 42, and 

43), and thus, are not addressed in this PA. 

Therefore, pathways (2), (42), and (43) are 

included in the dose analysis for atmospheric 

releases of 3

H and 14C. Please note that the 

requirements for radon are specified as a flux term 

in the performance objectives and are dealt with 

separately in this PA. 

Finally, contamination of cover soil over the 

TFF as a result of biointrusion of burrowing 

animals or plant roots [pathway (12)] also is not 

considered a major transport pathway. It is 

acknowledged that biointrusion is a potentially 

significant pathway of contamination of cover soil 

over a LLW facility (McKenzie et al. 1983). 

However, studies at the INEEL have indicated that 

burrowing animals and plant roots for species of 

the area do not exceed 7ft (2 m) in depth (Arthur 

1982, Reynolds and Fraley 1989; Markham 1987; 

Arthur, Grant, and Markham 1983; Reynolds and 

Laundre 1988). Therefore, given the 10 ft (3 m) of 

cover over the vaults, this pathway is not 

considered in this PA. Data on burrowing animals 

and plant roots are provided in Section 2.1.3. 

In summary, of the original 47 pathways 

identified in Figure 3-13, the following are 

considered to be of possible consequence to offsite members of the public and are considered 

further in this PA: 

• Pathways related to contaminated groundwater 

believed to be of potential consequence. This 

includes leaching of the waste form, resulting 

in contamination of groundwater and 

contamination of agricultural crops and 

animals as a result of irrigation with 

contaminated groundwater [pathways (1), (3), 

(19), (21), (23), (25), (26), (41), (44), and 

(45)]. 

• The release of the volatile radionuclides 3

H 

and 14C [pathways (2), (42), and (43)]. 
3-19

Radon releases are considered separately from 

these volatile radionuclides because the 

performance objective is defined in terms of flux 

rate for radon. 

3.2.3 Exposure Scenarios 

Three exposure scenarios were considered in 

the TFF PA: 1) drinking water dose from 

groundwater, 2) all-pathways dose from 

groundwater, and 3) air dispersion pathways. In 

addition, the intruder pathway and radon flux 

analyses also were considered and are described in 

Sections 3.3.4 and 5. 

3.2.3.1 Protection of Groundwater. As 

discussed in Section 1.4.2, the groundwater 

protection standard chosen for evaluation in this 

PA is a drinking water dose of 4 mrem/yr. The 

nearest location from the disposal site for off-site 

members of the public depends on the time period 

after disposal. During the period of active 

institutional control for the first 100 yr after 

facility closure, off-site members of the public are 

assumed to be located no closer to the disposal site 

than the present boundary of the INEEL. 

However, after active institutional control ceases, 

off-site members of the public could be located as 

close as 100 m (300 ft)a

 from the TFF. However, 

for the groundwater pathway, the member of the 

public is assumed to have a well located in the 

groundwater at the point of maximum 

concentration, which for the analyses presented in 

this PA is located 600 m (2,000 ft) from the 

facility.

As discussed in Section 3.2.2, the primary 

mechanism for transport of radionuclides from the 

TFF is expected to be leaching of radionuclides to 

the groundwater and subsequent human 

consumption. Thus, in the dose analysis for 

groundwater protection, an off-site member of the 

public is assumed to use water from a well for 

domestic purposes. The well is assumed to be 

located where the maximum concentrations of 

 

a. Inputs to the modeling software must be in metric units. 

Therefore, in these sections, measurements are given in metric 

units, followed by U.S. customary units. 

radionuclides in groundwater are predicted to 

occur. 

3.2.3.2 All-Pathways Exposure Scenario. 

As stated previously, the primary mechanism for 

transport of radionuclides from the TFF is 

expected to be leaching to the groundwater and 

subsequent human consumption. Thus, in the allpathways dose analysis, an off-site member of the 

public is assumed to use water from a well for 

domestic purposes. The well is assumed to be 

located where the maximum concentrations of 

radionuclides in groundwater are predicted to 

occur.

The following exposure pathways involving 

the use of contaminated well water are assumed to 

occur: 

• Direct ingestion of contaminated water 

• Ingestion of milk and meat from dairy and 

beef cattle that drink contaminated water 

• Ingestion of vegetables grown in garden soil 

irrigated with contaminated water 

• Ingestion of milk and meat from dairy and 

beef cattle that eat fodder from pasture 

irrigated with contaminated water. 

Additional exposure pathways for off-site 

members of the public could involve releases of 

radionuclides into the air (i.e., volatile 

radionuclides). Exposures from the air pathway 

also are considered in this PA. 

3.2.3.3 Air Dispersion Scenario. Volatile 

radionuclides 3

H and 14C were evaluated in the air 

dispersion scenario. The depth to the waste and the 

waste physical characteristics (i.e., grouted waste) 

limits the analysis to volatile radionuclides in the 

waste form. Additional details on the air 

dispersion scenario are provided in Section 3.3.5.

3.2.4 Groundwater Transport 

Radionuclide Screening 

The TFF inventory includes several 

radionuclides that will not be significant in terms 

of projected receptor doses from the groundwater 
3-20

pathway. Therefore, screening analysis methods 

were investigated for use in the TFF PA to reduce 

the number of radionuclides considered in the 

groundwater analysis. A more detailed analysis of 

releases from the disposal facility and transport in 

groundwater is required for the radionuclides 

selected by the screening procedure. 

The groundwater screening analysis 

considered that active institutional control will be 

maintained over the disposal site for 100 yr after 

facility closure and, furthermore, that the 

performance of the facility will be monitored to 

detect and/or prevent significant releases of 

radionuclides to the environment throughout the 

period of active institutional control. This 

consideration eliminates from concern any 

radionuclide with a half-life less than about 5 yr 

because the inventory in the waste at 100 yr after 

facility closure will be reduced to innocuous levels 

by radioactive decay. However, radionuclides with 

half-lives less than 5 yr cannot be neglected if the 

radionuclide appears in a decay chain. This is 

because its activity may increase with time due to 

decay of a parent radionuclide, unless the parent is 

also short-lived. 

The use of the 5-yr half-life screening criteria 

results in the elimination of the following 

radionuclides from further consideration: 102Rh, 

119mSn, 134Cs, 150Eu, 153Gd, 155Eu, 55Fe, and 171Tm. 

In addition, the following radionuclides in short 

decay chains were eliminated from further analysis 

since the parent and progeny each have half-lives 

less than 5 yr: 106Ru/106Rh, 109Cd/109mAg, 

110mAg/110Ag, 125Sb/125mTe, and 144Ce/144mPr/144Pr. 

Additional radionuclides were screened from 

further consideration since their half-lives indicate 

that they are either stable or have such long halflives that their specific activity, and thus 

contribution, to dose would be insignificant. These 

radionuclides include 142Ce (5E+16 yr), 149Sm 

(1E+15 yr), 144Nd (2.4E+15 yr), and 148Sm 

(2E+14 yr). 

The screening procedure used for the TFF 

source inventory was based first on the 

concentrations of radionuclides in the waste porewater that would give an annual effective dose 

EDE of 4 mrem/yr from consumption of 70 oz/d 

(2 L/d) [i.e., 190 gal/yr (730 L/yr)] of 

contaminated water. The 4 mrem/yr standard was 

used for screening each nuclide because this 

portion of the screening process does not consider 

transport from the waste form to the groundwater. 

Prior groundwater transport analyses provided in 

Appendix F show that the screening analysis based 

on an individual drinking the pore water in the 

waste is very conservative because it does not 

consider transport through the unsaturated zone 

and subsequent dilution. Therefore, the entire limit 

was applied to each radionuclide because of the 

conservative nature of the screening analysis based 

on waste pore-water concentrations to limit the 

number of radionuclides considered in the 

analysis. Thus, for radionuclide i, the limiting 

concentration in the waste pore-water in units of 

μCi/L is given by 

730 ( )

4

, L yr DCF mrem Ci

mrem

iw limit C i ⋅ μ = (3-9) 

where 

Ciw,limit = limiting concentration in 

drinking water for radionuclide i 

(μCi/L) 

730 L/yr = drinking water ingestion rate 

DCFi = ingestion dose conversion factor 

for radionuclide I (mrem/μCi). 

Given the assumption of a 100-yr active 

institutional control period, the screening 

procedure can be based on the concentrations of 

radionuclides in the waste at 100 yr after disposal. 

In addition, radionuclide concentrations in the 

waste at 500 and 1,000 yr after closure were 

calculated to evaluate progeny ingrowth from 

decay chains. 

The next step in the screening procedure is to 

convert the assumed radionuclide concentration in 

the waste-form to concentrations in the waste 

pore-water. These concentrations are clearly 

maximum (conservative) values, as they represent 

the pore-solution concentration in the waste form 

and do not consider the transport time out of the 

waste-form or the dilution due to transport in the 

unsaturated zone. Given the concentrations in the 
3-21

disposal facility at 100, 500, and 1,000 yr after 

disposal, the maximum concentrations in 

groundwater obtained from the screening model 

are given by 

d s

iv iw,max(t ) K

C (t ) C

θ + ρ = (3-10) 

where 

Ciw,max(t) = maximum radionuclide 

concentration in groundwater 

for radionuclide i at time t μCi 

Civ(t) = concentration in the disposal 

facility for radionuclide i at time 

t μCi 

0.001 = conversion factor to convert the 

waste concentration from 

μCi/m3

 to μCi/L of the solid 

waste 

θ = porosity of the waste form 

(unitless) 

Kd = equilibrium solid/solution 

distribution coefficient (L/kg) 

ρs = bulk density of the waste form 

(kg/L). 

The concentration in the waste forms, Civ, 

were based on the worst-case inventories for the 

tanks and sand pad presented in Appendix A. The 

conservative case presented in Section 2.3 was not 

used for the screening analysis. This was done to 

ensure that all radionuclides of importance would 

remain after screening. The volume of the waste in 

a tank was based on 55.5 m3

, while the sand pad 

volume was 23.4 m3

. 

The denominator of Equation (3-10) accounts 

for the waste concentration being a total value. 

The denominator of the equation partitions the 

total activity of each radionuclide from the solid 

portion of the waste-form to the pore-water of the 

waste-form. A simple division by the Kd is no 

applicable in this case, since the inventories are 

the totals in the waste (i.e., both liquid fraction and 

solid fractions) before the waste has been grouted. 

Therefore, after grouting the total inventory must 

be partitioned into the grout pore-water. 

The worst-case inventory (see Appendix A) 

and worst-case equilibrium solid/solution 

distribution coefficients were used in the screening 

analysis to ensure a complete list of radionuclides. 

The porosity of the grout was assumed to be 

0.2; the sand pad had an assumed porosity of 0.34. 

The grout porosity will change over time because 

of degradation to approximately 0.4. However, 

this change does not result in different 

radionuclides being chosen from the screening 

analysis. The bulk density of grout was assumed to 

be 2.1E+06 g/m3

, while the sand pad was assumed 

to be at a density of 1.76E+06 g/m3

. 

Given Ciw,limit and Ciw,max , at 100, 500, and 

1000 yr, a radionuclide is selected for further 

analysis only if the following inequality is 

satisfied: 

Ciw,max(t) > Ciw,limit (3-11) 

The screening procedure is sufficiently 

conservative so that any radionuclides eliminated 

from further consideration would not contribute 

more than a small fraction of the dose limit of 

4 mrem/yr (EDE) from contamination of 

groundwater, considering that the waste porewater concentration were used. 

The development of pore-water concentrations 

in the tank grout required the evaluation of 

distribution coefficients (i.e., Kd values) specific to 

grout. These grout distribution coefficients were 

taken from studies conducted by Allard, Persson, 

and Torstenflt (1985) and Bradbury and Sarott 

(1995) (see Section 2.1.6). The development of 

pore-water concentrations in the sand pad required 

the evaluation of Kd values specific to sand. These 

coefficients were taken from Sheppard and 

Thibault (1990) (see Section 2.1.6). The selected 

distribution coefficients for grout and sand are 

provided in Table 3-1. 

The worst-case TFF inventories, presented in 

Appendix A, were screened according to the above 

calculations. The screened inventory for the tank is 

provided in Table 3-2. The same process was used 
3-22

Table 3-1. Distribution coefficients for grout and sand. 

Nuclide 

Grouta

(m3/kg) 

Sanda

(m3/kg) Nuclide 

Grouta

(m3/kg) 

Sanda

(m3/kg) 

10Be 1.00E−03 2.50E−01 94Nb 5.00E−01 1.60E−01 

14C 1.00E−02 5.00E−03 63Ni 1.00E−01 4.00E−01 

113mCd 1.00E−01 8.00E−02 107Pd 1.00E−01 5.50E−02 

60Co 1.00E−01 6.00E−02 146Pm 0.00E+00 0.00E+00 

135Cs 2.00E−03 2.80E−01 87Rb 0.00E+00 5.50E−02 

150Eu 5.00E+00 0.00E+00 79Se 1.00E−04 1.50E−01 

154Eu 5.00E+00 1.90E+00 146Sm 5.00E+00 2.45E−01 

3

H 0.00E+00 0.00E+00 151Sm 5.00E+00 2.45E−01 

166mHo 5.00E+00 2.50E−01 121mSn 1.00E+00 1.30E−01 

129I 2.00E−03 1.00E−03 98Tc 1.00E+00 1.00E−04 

115In 1.00E−01 0.00E+00 99Tc 1.00E+00 1.00E−04 

138La 5.00E+00 0.00E+00 123Te 0.00E+00 1.25E−01 

Short Decay Chains 

108Ag 1.00E−03 9.00E−02 126Sb 0.00E+00 4.50E−02 

108mAg 1.00E−03 9.00E−02 126mSb 0.00E+00 4.50E−02 

137mBa 1.00E−03 4.00E−04 147Sm 5.00E+00 2.45E−01 

137Cs 2.00E−03 2.80E−01 126Sn 1.00E+00 1.30E−01 

152Eu 5.00E+00 1.90E+00 90Sr 1.00E−03 1.50E−02 

152Gd 0.00E+00 0.00E+00 90Y 1.00E−03 1.70E−01 

93mNb 5.00E−01 1.60E−01 93Zr 5.00E+00 6.00E−01 

147Pm 0.00E+00 0.00E+00 

252Cf Decay Chain 

228Ac 5.00E+00 4.50E−01 240Pu 5.00E+00 5.50E−01 

212Bi 5.00E−01 1.00E−01 244Pu 5.00E+00 5.50E−01 

252Cf 0.00E+00 0.00E+00 224Ra 5.00E−02 5.00E−01 

244Cm 5.00E+00 4.00E+00 228Ra 5.00E−02 5.00E−01 

248Cm 5.00E+00 4.00E+00 220Rn 0.00E+00 0.00E+00 

240Np 5.00E+00 5.00E−03 228Th 5.00E+00 3.20E+00 

240mNp 5.00E+00 5.00E−03 232Th 5.00E+00 3.20E+00 

212Pb 5.00E−01 2.70E−01 208Tl 0.00E+00 0.00E+00 

212Po 5.00E−01 1.50E−01 232U 2.00E+00 3.50E−02 

216Po 5.00E−01 1.50E−01 236U 2.00E+00 3.50E−02 

236Pu 5.00E+00 5.50E−01 240U 2.00E+00 3.50E−02 

249Cf Decay Chain 

225Ac 5.00E+00 4.50E−01 209Pb 5.00E−01 2.70E−01 

241Am 5.00E+00 1.90E+00 213Po 5.00E−01 1.50E−01 

217At 0.00E+00 0.00E+00 241Pu 5.00E+00 5.50E−01 
Table 3-1. (continued). 

3-23

Nuclide 

Grouta

(m3/kg) 

Sanda

(m3/kg) Nuclide 

Grouta

(m3/kg) 

Sanda

(m3/kg) 

213Bi 5.00E−01 1.00E−01 225Ra 5.00E−02 5.00E−01 

249Cf 0.00E+00 0.00E+00 229Th 5.00E+00 3.20E+00 

245Cm 5.00E+00 4.00E+00 209Tl 0.00E+00 0.00E+00 

221Fr 0.00E+00 0.00E+00 233U 2.00E+00 3.50E−02 

237Np 5.00E+00 5.00E−03 237U 2.00E+00 3.50E−02 

233Pa 5.00E+00 5.50E−01 

251Cf Decay Chain 

227Ac 5.00E+00 4.50E−01 211Po 5.00E−01 1.50E−01 

243Am 5.00E+00 1.90E+00 215Po 5.00E−01 1.50E−01 

211Bi 5.00E−01 1.00E−01 239Pu 5.00E+00 5.50E−01 

251Cf 0.00E+00 0.00E+00 243Pu 5.00E+00 5.50E−01 

243Cm 5.00E+00 4.00E+00 223Ra 5.00E−02 5.00E−01 

247Cm 5.00E+00 4.00E+00 219Rn 0.00E+00 0.00E+00 

223Fr 0.00E+00 0.00E+00 227Th 5.00E+00 3.20E+00 

235Np 5.00E+00 5.00E−03 231Th 5.00E+00 3.20E+00 

239Np 5.00E+00 5.00E−03 207Tl 0.00E+00 0.00E+00 

231Pa 5.00E+00 5.50E−01 235U 2.00E+00 3.50E−02 

211Pb 5.00E−01 2.70E−01 

250Cf Decay Chain 

242Am 5.00E+00 1.90E+00 210Po 5.00E−01 1.50E−01 

242mAm 5.00E+00 1.90E+00 214Po 5.00E−01 1.50E−01 

210Bi 5.00E−01 1.00E−01 218Po 5.00E−01 1.50E−01 

214Bi 5.00E−01 1.00E−01 238Pu 5.00E+00 5.50E−01 

250Cf 0.00E+00 0.00E+00 242Pu 5.00E+00 5.50E−01 

242Cm 5.00E+00 4.00E+00 226Ra 5.00E−02 5.00E−01 

246Cm 5.00E+00 4.00E+00 222Rn 0.00E+00 0.00E+00 

238Np 5.00E+00 5.00E−03 230Th 5.00E+00 3.20E+00 

234Pa 5.00E+00 5.50E−01 234Th 5.00E+00 3.20E+00 

234mPa 5.00E+00 5.50E−01 210Tl 0.00E+00 0.00E+00 

210Pb 5.00E−01 2.70E−01 234U 2.00E+00 3.50E−02 

214Pb 5.00E−01 2.70E−01 238U 2.00E+00 3.50E−02 

a. Distribution coefficients assumed to be zero for all radionuclides not listed in the references. 
3-24

Table 3-2. Tank activity groundwater screening for the drinking water pathway. 

Nuclide 

Half-life 

(yr) 

Kd

(m3

/kg) 

Ingestion

DCF 

(mrem/μCi)

Ciw, limit

Allowable 

Concentration

(μCi/L) 

100 yr 

Ciw,max 

(μCi/L) 

500 yr 

Ciw,max 

(μCi/L) 

1000 yr 

Ciw,max 

(μCi/L) 

225Ac 2.74E−02 5.00E+00 1.11E+02 4.94E−05 6.24E−10 3.42E−09 8.29E−09 

227Ac 2.18E+01 5.00E+00 1.41E+04 3.89E−07 2.25E−08 3.07E−08 4.05E−08 

228Ac 6.99E−04 5.00E+00 2.16E+00 2.54E−03 1.83E−13 2.98E−13 4.41E−13 

108Ag 4.51E−06 1.00E−03 0.00E+00 NAa 2.24E−08 2.53E−09 1.65E−10 

108mAg 1.27E+02 1.00E−03 7.62E+00 7.19E−04 2.41E−07 2.72E−08 1.77E−09 

241Am 4.32E+02 5.00E+00 3.64E+03 1.51E−06 2.09E−03 1.11E−03 4.96E−04 

242Am 1.83E−03 5.00E+00 1.41E+00 3.89E−03 2.04E−06 3.29E−07 3.36E−08 

242mAm 1.52E+02 5.00E+00 3.52E+03 1.56E−06 2.05E−06 3.30E−07 3.38E−08 

243Am 7.38E+03 5.00E+00 3.62E+03 1.51E−06 4.59E−06 4.42E−06 4.22E−06 

217At 1.02E−09 0.00E+00 0.00E+00 NA 3.37E−05 1.85E−04 4.48E−04 

137mBa 4.86E−06 1.00E−03 0.00E+00 NA 3.60E+03 3.68E−01 3.77E−06 

10Be 1.60E+06 1.00E−03 4.66E+00 1.18E−03 3.03E−06 3.03E−06 3.03E−06 

210Bi 1.37E−02 5.00E−01 6.40E+00 8.56E−04 1.05E−07 1.14E−06 3.74E−06 

211Bi 4.05E−06 5.00E−01 0.00E+00 NA 2.25E−07 3.07E−07 4.05E−07 

212Bi 1.15E−04 5.00E−01 1.06E+00 5.17E−03 1.62E−06 3.45E−08 2.85E−10 

213Bi 8.68E−05 5.00E−01 7.22E−01 7.59E−03 6.24E−09 3.42E−08 8.29E−08 

214Bi 3.78E−05 5.00E−01 2.83E−01 1.94E−02 1.47E−07 1.27E−06 3.95E−06 

14C 5.73E+03 1.00E−02 2.09E+00 2.62E−03 8.05E−01 7.67E−01 7.22E−01 

113mCd 1.37E+01 1.00E−01 1.61E+02 3.40E−05 1.26E−04 2.06E−13 2.13E−24 

249Cf 3.51E+02 0.00E+00 4.74E+03 1.16E−06 1.65E−10 7.46E−11 2.78E−11 

250Cf 1.31E+01 0.00E+00 2.13E+03 2.57E−06 4.24E−13 2.64E−22 8.20E−34 

251Cf 9.00E+02 0.00E+00 4.85E+03 1.13E−06 2.94E−12 2.16E−12 1.47E−12 

252Cf 2.64E+00 0.00E+00 1.08E+03 5.07E−06 3.52E−40 8.30E−86 7.66E−143 

242Cm 4.47E−01 5.00E+00 1.15E+02 4.76E−05 1.69E−06 2.73E−07 2.79E−08 

243Cm 2.85E+01 5.00E+00 2.51E+03 2.18E−06 4.07E−07 2.42E−11 1.27E−16 

244Cm 1.81E+01 5.00E+00 2.02E+03 2.71E−06 5.04E−06 1.13E−12 5.52E−21 

245Cm 8.50E+03 5.00E+00 3.74E+03 1.47E−06 6.56E−08 6.35E−08 6.10E−08 

246Cm 4.75E+03 5.00E+00 3.70E+03 1.48E−06 4.24E−09 4.00E−09 3.71E−09 

247Cm 1.56E+07 5.00E+00 3.42E+03 1.60E−06 4.80E−15 4.80E−15 4.80E−15 

248Cm 3.39E+05 5.00E+00 1.36E+04 4.03E−07 5.21E−15 5.21E−15 5.20E−15 

60Co 5.27E+00 1.00E−01 2.69E+01 2.04E−04 4.79E−08 6.94E−31 1.96E−59 

135Cs 2.30E+06 2.00E−03 7.07E+00 7.75E−04 4.74E−01 4.74E−01 4.74E−01 

137Cs 3.02E+01 2.00E−03 5.00E+01 1.10E−04 1.99E+03 2.03E−01 2.08E−06 
Table 3-2. (continued). 

3-25

Nuclide 

Half-life 

(yr) 

Kd

(m3

/kg) 

Ingestion

DCF 

(mrem/μCi)

Ciw, limit

Allowable 

Concentration

(μCi/L) 

100 yr 

Ciw,max 

(μCi/L) 

500 yr 

Ciw,max 

(μCi/L) 

1000 yr 

Ciw,max 

(μCi/L) 

152Eu 1.36E+01 5.00E+00 6.48E+00 8.46E−04 1.77E−06 2.48E−15 2.13E−26 

154Eu 8.80E+00 5.00E+00 9.55E+00 5.74E−04 1.36E−06 2.85E−20 2.26E−37 

221Fr 9.13E−06 0.00E+00 0.00E+00 NA 3.37E−05 1.85E−04 4.48E−04 

223Fr 4.14E−05 0.00E+00 8.62E+00 6.36E−04 1.68E−05 2.29E−05 3.02E−05 

152Gd 1.10E+14 0.00E+00 1.61E+02 3.40E−05 1.80E−11 1.80E−11 1.80E−11 

3

H 1.23E+01 0.00E+00 6.40E−02 8.56E−02 2.21E−01 3.48E−11 1.94E−23 

166mHo 1.20E+03 5.00E+00 8.07E+00 6.79E−04 9.37E−09 7.44E−09 5.57E−09 

129I 1.57E+07 2.00E−03 2.76E+02 1.99E−05 2.77E−02 2.77E−02 2.77E−02 

115In 4.60E+15 1.00E−01 1.58E+02 3.47E−05 1.07E−12 1.07E−12 1.07E−12 

138La 1.12E+11 5.00E+00 5.88E+00 9.32E−04 4.30E−14 4.30E−14 4.30E−14 

93mNb 1.46E+01 5.00E−01 5.22E−01 1.05E−02 4.79E−03 4.80E−03 4.79E−03 

94Nb 2.03E+04 5.00E−01 7.14E+00 7.67E−04 1.42E−01 1.40E−01 1.38E−01 

59Ni 7.50E+04 1.00E−01 2.10E−01 2.61E−02 1.60E−04 1.60E−04 1.60E−04 

63Ni 1.00E+02 1.00E−01 5.77E−01 9.50E−03 1.57E−01 9.85E−03 3.09E−04 

235Np 1.08E+00 5.00E+00 2.43E−01 2.26E−02 2.91E−61 2.69E−172 0.00E+00 

237Np 2.14E+06 5.00E+00 4.44E+03 1.23E−06 1.67E−05 1.69E−05 1.70E−05 

238Np 5.80E−03 5.00E+00 4.00E+00 1.37E−03 9.75E−09 1.57E−09 1.61E−10 

239Np 6.45E−03 5.00E+00 3.26E+00 1.68E−03 4.59E−06 4.42E−06 4.22E−06 

240Np 1.24E−04 5.00E+00 2.37E−01 2.31E−02 1.64E−16 1.64E−16 1.64E−16 

240mNp 1.41E−05 5.00E+00 0.00E+00 NA 1.49E−13 1.49E−13 1.49E−13 

231Pa 3.28E+04 5.00E+00 1.06E+04 5.17E−07 2.35E−08 3.14E−08 4.12E−08 

233Pa 7.39E−02 5.00E+00 3.63E+00 1.51E−03 1.67E−05 1.69E−05 1.70E−05 

234Pa 7.64E−04 5.00E+00 2.16E+00 2.54E−03 1.16E−09 1.16E−09 1.16E−09 

234mPa 2.22E−06 5.00E+00 0.00E+00 NA 7.25E−07 7.25E−07 7.25E−07 

209Pb 3.71E−04 5.00E−01 2.13E−01 2.57E−02 6.24E−09 3.42E−08 8.29E−08 

210Pb 2.23E+01 5.00E−01 5.37E+03 1.02E−06 1.05E−07 1.14E−06 3.74E−06 

211Pb 6.86E−05 5.00E−01 5.25E−01 1.04E−02 2.25E−07 3.07E−07 4.05E−07 

212Pb 1.21E−03 5.00E−01 4.55E+01 1.20E−04 1.62E−06 3.45E−08 2.85E−10 

214Pb 5.10E−05 5.00E−01 6.25E−01 8.77E−03 1.47E−07 1.27E−06 3.95E−06 

107Pd 6.50E+06 1.00E−01 1.49E−01 3.68E−02 1.78E−04 1.78E−04 1.78E−04 

146Pm 5.53E+00 0.00E+00 3.67E+00 1.49E−03 4.23E−07 7.46E−29 4.79E−56 

147Pm 2.62E+00 0.00E+00 1.05E+00 5.22E−03 2.25E−10 2.83E−56 1.20E−113 

210Po 3.79E−01 5.00E−01 1.90E+03 2.88E−06 1.05E−07 1.14E−06 3.74E−06 

211Po 1.64E−08 5.00E−01 0.00E+00 NA 6.14E−10 8.39E−10 1.11E−09 
Table 3-2. (continued). 

3-26

Nuclide 

Half-life 

(yr) 

Kd

(m3

/kg) 

Ingestion

DCF 

(mrem/μCi)

Ciw, limit

Allowable 

Concentration

(μCi/L) 

100 yr 

Ciw,max 

(μCi/L) 

500 yr 

Ciw,max 

(μCi/L) 

1000 yr 

Ciw,max 

(μCi/L) 

212Po 9.44E−15 5.00E−01 0.00E+00 NA 1.04E−06 2.21E−08 1.82E−10 

213Po 1.33E−13 5.00E−01 0.00E+00 NA 6.10E−09 3.34E−08 8.11E−08 

214Po 2.02E−12 5.00E−01 0.00E+00 NA 1.47E−07 1.27E−06 3.95E−06 

215Po 2.47E−11 5.00E−01 0.00E+00 NA 2.25E−07 3.07E−07 4.05E−07 

216Po 4.63E−09 5.00E−01 0.00E+00 NA 1.62E−06 3.45E−08 2.85E−10 

218Po 5.80E−06 5.00E−01 0.00E+00 NA 1.47E−07 1.27E−06 3.95E−06 

236Pu 2.85E+00 5.00E+00 1.17E+03 4.68E−06 6.86E−19 3.99E−61 6.42E−114 

238Pu 8.78E+01 5.00E+00 3.20E+03 1.71E−06 1.62E−02 6.87E−04 1.33E−05 

239Pu 2.41E+04 5.00E+00 3.54E+03 1.55E−06 2.82E−03 2.79E−03 2.75E−03 

240Pu 6.54E+03 5.00E+00 3.54E+03 1.55E−06 2.29E−03 2.20E−03 2.08E−03 

241Pu 1.44E+01 5.00E+00 6.85E+01 8.00E−05 2.49E−04 6.36E−08 6.11E−08 

242Pu 3.76E+05 5.00E+00 3.36E+03 1.63E−06 1.73E−06 1.73E−06 1.73E−06 

243Pu 5.65E−04 5.00E+00 3.34E−01 1.64E−02 4.80E−15 4.80E−15 4.80E−15 

244Pu 8.26E+07 5.00E+00 3.32E+03 1.65E−06 1.49E−13 1.49E−13 1.49E−13 

223Ra 3.13E−02 5.00E−02 6.59E+02 8.31E−06 2.25E−06 3.07E−06 4.05E−06 

224Ra 1.00E−02 5.00E−02 3.66E+02 1.50E−05 1.62E−05 3.45E−07 2.84E−09 

225Ra 4.05E−02 5.00E−02 3.85E+02 1.42E−05 6.23E−08 3.41E−07 8.27E−07 

226Ra 1.60E+03 5.00E−02 1.32E+03 4.15E−06 1.47E−06 1.27E−05 3.94E−05 

228Ra 5.75E+00 5.00E−02 1.44E+03 3.81E−06 1.83E−11 2.97E−11 4.41E−11 

87Rb 4.73E+10 0.00E+00 4.92E+00 1.11E−03 3.48E−04 3.48E−04 3.48E−04 

219Rn 1.25E−07 0.00E+00 0.00E+00 NA 1.21E−03 1.66E−03 2.19E−03 

220Rn 1.76E−06 0.00E+00 0.00E+00 NA 8.77E−03 1.87E−04 1.54E−06 

222Rn 1.05E−02 0.00E+00 0.00E+00 NA 7.95E−04 6.84E−03 2.13E−02 

126Sb 3.40E−02 0.00E+00 1.07E+01 5.12E−04 6.87E−01 6.85E−01 6.83E−01 

126mSb 3.62E−05 0.00E+00 9.36E−02 5.85E−02 4.91E+00 4.89E+00 4.88E+00 

79Se 6.50E+04 1.00E−04 8.70E+00 6.30E−04 2.57E+00 2.56E+00 2.55E+00 

146Sm 7.00E+07 5.00E+00 2.04E+02 2.69E−05 6.12E−11 6.12E−11 6.12E−11 

147Sm 6.90E+09 5.00E+00 1.85E+02 2.96E−05 1.65E−09 1.65E−09 1.65E−09 

151Sm 9.00E+01 5.00E+00 3.89E−01 1.41E−02 3.14E−02 1.44E−03 3.07E−05 

121mSn 7.60E+01 1.00E+00 1.55E+00 3.54E−03 2.49E−04 6.49E−06 6.80E−08 

126Sn 1.00E+05 1.00E+00 1.95E+01 2.81E−04 4.54E−04 4.53E−04 4.52E−04 

90Sr 2.86E+01 1.00E−03 1.42E+02 3.86E−05 6.25E+03 3.85E−01 2.10E−06 

98Tc 4.20E+06 1.00E−03 4.88E+00 1.12E−03 2.61E−06 2.61E−06 2.61E−06 

99Tc 2.13E+05 1.00E−03 1.46E+00 3.75E−03 8.60E+01 8.59E+01 8.58E+01 
Table 3-2. (continued). 

3-27

Nuclide 

Half-life 

(yr) 

Kd

(m3

/kg) 

Ingestion

DCF 

(mrem/μCi)

Ciw, limit

Allowable 

Concentration

(μCi/L) 

100 yr 

Ciw,max 

(μCi/L) 

500 yr 

Ciw,max 

(μCi/L) 

1000 yr 

Ciw,max 

(μCi/L) 

123Te 1.00E+13 0.00E+00 4.18E+00 1.31E−03 4.47E−12 4.47E−12 4.47E−12 

227Th 5.12E−02 5.00E+00 3.81E+01 1.44E−04 2.22E−08 3.03E−08 4.00E−08 

228Th 1.91E+00 5.00E+00 3.96E+02 1.38E−05 1.62E−07 3.45E−09 2.85E−11 

229Th 7.34E+03 5.00E+00 3.53E+03 1.55E−06 6.24E−10 3.42E−09 8.29E−09 

230Th 7.70E+04 5.00E+00 5.48E+02 1.00E−05 3.65E−07 1.06E−06 1.92E−06 

231Th 2.91E−03 5.00E+00 1.35E+00 4.06E−03 9.60E−07 9.61E−07 9.62E−07 

232Th 1.41E+10 5.00E+00 2.73E+03 2.01E−06 1.86E−13 3.00E−13 4.44E−13 

234Th 6.60E−02 5.00E+00 1.37E+01 4.00E−04 7.25E−07 7.25E−07 7.25E−07 

207Tl 9.07E−06 0.00E+00 0.00E+00 NA 1.21E−03 1.66E−03 2.18E−03 

208Tl 5.81E−06 0.00E+00 0.00E+00 NA 3.15E−03 6.70E−05 5.53E−07 

209Tl 4.18E−06 0.00E+00 0.00E+00 NA 7.28E−07 3.99E−06 9.67E−06 

210Tl 2.47E−06 0.00E+00 0 NA 0.00E+00 0.00E+00 0.00E+00 

232U 7.20E+01 2.00E+00 1.31E+03 4.18E−06 3.95E−07 8.41E−09 6.82E−11 

233U 1.59E+05 2.00E+00 2.89E+02 1.90E−05 1.53E−07 2.26E−07 3.18E−07 

234U 2.45E+05 2.00E+00 2.83E+02 1.94E−05 4.71E−04 4.85E−04 4.85E−04 

235U 7.04E+08 2.00E+00 2.66E+02 2.06E−05 2.40E−06 2.40E−06 2.41E−06 

236U 3.42E+06 2.00E+00 2.69E+02 2.04E−05 1.45E−05 1.45E−05 1.46E−05 

237U 1.85E−02 2.00E+00 3.17E+00 1.73E−03 1.53E−08 3.90E−12 3.74E−12 

238U 4.47E+09 2.00E+00 2.55E+02 2.15E−05 1.81E−06 1.81E−06 1.81E−06 

240U 1.61E−03 2.00E+00 4.44E+00 1.23E−03 3.72E−13 3.72E−13 3.72E−13 

90Y 7.32E−03 1.00E−03 1.08E+01 5.07E−04 6.25E+03 3.85E−01 2.10E−06 

93Zr 1.53E+06 5.00E+00 1.66E+00 3.30E−03 4.80E−04 4.80E−04 4.80E−04 
3-28

for the sand pads; these results are provided in 

Table 3-3. The radionuclides selected for analysis 

are given in Table 3-4. 

The next step in the screening process 

involved evaluation of the release of the 

radionuclides from the waste-form and the 

resulting groundwater concentrations. Releases 

and groundwater concentrations were previously 

analyzed. The results are presented in Appendix F. 

Based on these analyses, only 129I, 99Tc, 90Sr, and 

14C were found to result in appreciable 

groundwater concentrations and significant doses. 

The Idaho High-Level Waste & Facilities 

Disposition Final Environmental Impact Statement 

(DOE 2002) also conducted screening and 

groundwater transport analyses for the TFF. Their 

screening and analyses found that only 99Tc and 

129I resulted in significant doses from the 

groundwater pathway. 

Therefore, base on previous modeling of the 

TFF presented in Appendix F, four radionuclides 

(i.e., 129I, 99Tc, 90Sr, and 14C) were determined to 

result in appreciable groundwater concentrations 

and doses. Therefore, the groundwater analysis 

was focused on the four radionuclides. 

The releases calculated with DUST-MS 

(Sullivan 1993, 1998) discussed in Section 3.3, 

show that several radionuclides are never released 

from the waste form during the 1,000-yr 

compliance period. This is because distribution 

coefficients (Kd values) are used in the DUST-MS 

release model, whereas the screening method does 

not take into account the transport from the tank 

and sand pad. Therefore, radionuclides not 

released in significant quantities during the 

1,000-yr compliance period were not modeled in 

the PORFLOW (Runchal 1997; ACRi 2000) 

groundwater model. However, all radionuclides 

selected in the groundwater screening were run 

using the long-term groundwater model 

GWSCREEN (Rood 1998) in previous modeling 

analyses for the TFF. Results from the previous 

analyses conducted for all radionuclides are 

presented in Appendix F: input parameters used in 

those analyses are provided in tables in this 

section. The modeling conducted in this report 

focused on the four radionuclides found to be the 

major contributors to the groundwater pathway 

doses in the previous modeling exercise (i.e., 90Sr, 

99Tc, 129I, and 14C). 

3.3 Models and Assumptions 

3.3.1 Source Term Model 

For the TFF concrete vaults, release rates of 

radionuclides were estimated using the Disposal 

Unit Source Term-Multiple Species (DUST-MS) 

computer code (Sullivan 1993, 1996). Onedimensional DUST-MS transport simulations were 

conducted for radionuclide sources in the grouted 

tank, piping, and the sand pad beneath the tank. 

Infiltration was assumed to contact and 

transport radionuclides after concrete degradation 

and tank corrosion. The concrete vault and grout 

between the vault wall and the tank were assumed 

to completely degrade after 100 yr, at which time 

infiltrating water contacted radionuclides in the 

sand pad and transported the radionuclides through 

the sand pad and the degraded vault floor to the 

basalt. At 500 yr, the stainless-steel tanks were 

assumed to have totally corroded and the grout 

inside the tank was assumed to have completely 

degraded. At this time, infiltrating water 

contacting radionuclides contained in the grout 

inside the tank was assumed to transport 

radionuclides through the degraded grout in the 

tank, sand pad, and degraded vault floor to the 

basalt. Piping releases were assumed to occur in 

the same manner as the tanks. The DUST-MS 

computer code modeled these two releases as 

separate and independent events. 

The two waste forms for these DUST-MS 

simulations were the radionuclides in the grouted 

tank and piping and the radionuclides in the sand 

pad. The DUST-MS waste form release rate model 

selected for these simulations was a surface rinse 

model. The surface rinse model accounts for 

partitioning between the infiltrating water and the 

radionuclides in the waste form. Additionally, 

partitioning and retardation were modeled in 

DUST-MS for radionuclide transport occurring in 

the grouted tank, sand pad, and vault floor. 

Radionuclide source locations specified in the 

DUST-MS computer code for the two waste forms 
3-29

Table 3-3. Sand pad activity groundwater screening for the drinking water pathway. 

Nuclide 

Half-life 

(yr) 

Kd

(m3/kg) 

Ingestion 

DCF 

(mrem/μCi) 

Ciw, limit

Allowable 

Concentration

(μCi/L) 

100 yr 

Ciw,max 

(μCi/L) 

500 yr 

Ciw,max 

(μCi/L) 

1000 yr 

Ciw,max 

(μCi/L) 225Ac 2.74E−02 4.50E−01 1.11E+02 4.94E−05 2.67E−10 8.00E−10 1.84E−09 227Ac 2.18E+01 4.50E−01 1.41E+04 3.89E−07 2.42E−07 7.72E−07 1.43E−06 228Ac 6.99E−04 4.50E−01 2.16E+00 2.54E−03 1.02E−08 1.96E−12 3.91E−12 108Ag 4.51E−06 9.00E−02 0.00E+00 NA 5.77E−12 6.50Ev13 4.24E−14 108mAg 1.27E+02 9.00E−02 7.62E+00 7.19E−04 6.20E−11 6.99E−12 4.56E−13 241Am 4.32E+02 1.90E+00 3.64E+03 1.51E−06 2.96E−03 1.56E−03 6.99E−04 242Am 1.83E−03 1.90E+00 1.41E+00 3.89E−03 1.84E−08 2.96Ev09 3.03E−10 242mAm 1.52E+02 1.90E+00 3.52E+03 1.56E−06 1.85E−08 2.98E−09 3.05E−10 243Am 7.38E+03 1.90E+00 3.62E+03 1.51E−06 2.58E−07 2.49E−07 2.37E−07 217At 1.02E−09 0 0.00E+00 NA 6.22E−07 1.87E−06 4.29E−06 137mBa 4.86E−06 4.00E−04 0.00E+00 NA 8.13E+02 8.30E−02 8.51E−07 10Be 1.60E+06 2.50E−01 4.66E+00 1.18E−03 2.43E−09 2.43Ev09 2.43E−09 210Bi 1.37E−02 1.00E−01 6.40E+00 8.56E−04 1.02E−06 7.91E−06 2.33E−05 211Bi 4.05E−06 1.00E−01 0.00E+00 NA 1.09E−06 3.47E−06 6.41E−06 212Bi 1.15E−04 1.00E−01 1.06E+00 5.17E−03 3.17E−07 5.28E−09 6.03E−11 213Bi 8.68E−05 1.00E−01 7.22E−01 7.59E−03 1.20E−09 3.59E−09 8.26E−09 214Bi 3.78E−05 1.00E−01 2.83E−01 1.94E−02 1.35E−06 8.67E−06 2.45E−05 14C 5.73E+03 5.00E−03 2.09E+00 2.62E−03 2.51E−07 2.39E−07 2.25E−07 113mCd 1.37E+01 8.00E−02 1.61E+02 3.40E−05 4.29E−06 7.00E−15 7.26E−26 249Cf 3.51E+02 0.00E+00 4.74E+03 1.16E−06 3.48E−15 1.58E−15 5.87E−16 250Cf 1.31E+01 0.00E+00 2.13E+03 2.57E−06 1.19E−18 7.43E−28 2.31E−39 251Cf 9.00E+02 0.00E+00 4.85E+03 1.13E−06 9.64E−18 7.08E−18 4.82E−18 252Cf 2.64E+00 0.00E+00 1.08E+03 5.07E−06 1.91E−35 4.50E−81 4.15E−138 242Cm 4.47E−01 4.00E+00 1.15E+02 4.76E−05 7.24E−09 1.17E−09 1.19E−10 243Cm 2.85E+01 4.00E+00 2.51E+03 2.18E−06 5.86E−11 3.49E−15 1.83E−20 244Cm 1.81E+01 4.00E+00 2.02E+03 2.71E−06 4.05E−09 9.10E−16 4.44E−24 245Cm 8.50E+03 4.00E+00 3.74E+03 1.47E−06 4.28E−11 4.14E−11 3.97E−11 246Cm 4.75E+03 4.00E+00 3.70E+03 1.48E−06 9.50E−13 8.96E−13 8.33E−13 247Cm 1.56E+07 4.00E+00 3.42E+03 1.60E−06 3.47E−19 3.47E−19 3.47E−19 248Cm 3.39E+05 4.00E+00 1.36E+04 4.03E−07 1.06E−19 1.06E−19 1.06E−19 60Co 5.27E+00 6.00E−02 2.69E+01 2.04E−04 1.75E−10 2.53E−33 7.13E−62 135Cs 2.30E+06 2.80E−01 7.07E+00 7.75E−04 7.21E−05 7.21E−05 7.21E−05 137Cs 3.02E+01 2.80E−01 5.00E+01 1.10E−04 1.82E+00 1.86E−04 1.90E−09 152Eu 1.36E+01 1.90E+00 6.48E+00 8.46E−04 1.01E−07 1.41E−16 1.21E−27 154Eu 8.80E+00 1.90E+00 9.55E+00 5.74E−04 8.69E−07 1.81E−20 1.44E−37 221Fr 9.13E−06 0 0.00E+00 NA 6.22E−07 1.87E−06 4.29E−06 223Fr 4.14E−05 0 8.62E+00 6.36E−04 7.81E−06 2.49E−05 4.60E−05 152Gd 1.10E+14 0.00E+00 1.61E+02 3.40E−05 5.70E−13 5.70E−13 5.70E−13 3

H 1.23E+01 0.00E+00 6.40E−02 8.56E−02 1.53E−23 2.41E−33 1.34E−45 166mHo 1.20E+03 2.50E−01 8.07E+00 6.79E−04 5.28E−09 4.19E−09 3.14E−09 129I 1.57E+07 1.00E−03 2.76E+02 1.99E−05 3.05E−06 3.05E−06 3.05E−06 
Table 3-3. (continued). 

3-30

Nuclide 

Half-life 

(yr) 

Kd

(m3

/kg) 

Ingestion 

DCF 

(mrem/μCi) 

Ciw, limit

Allowable 

Concentration

(μCi/L) 

100 yr 

Ciw,max 

(μCi/L) 

500 yr 

Ciw,max 

(μCi/L) 

1000 yr 

Ciw,max 

(μCi/L) 115In 4.60E+15 0.00E+00 1.58E+02 3.47E−05 6.49E−11 6.49E−11 6.49E−11 138La 1.12E+11 0.00E+00 5.88E+00 9.32E−04 9.33E−10 9.33E−10 9.33E−10 93mNb 1.46E+01 1.60E−01 5.22E−01 1.05E−02 4.45E−03 4.45E−03 4.45E−03 94Nb 2.03E+04 1.60E−01 7.14E+00 7.67E−04 4.81E−04 4.75E−04 4.67E−04 63Ni 1.00E+02 4.00E−01 5.77E−01 9.50E−03 6.92E−13 4.34E−14 1.36E−15 235Np 1.08E+00 5.00E−03 2.43E−01 2.26E−02 2.03E−52 1.88E−163 3.03E−302 237Np 2.14E+06 5.00E−03 4.44E+03 1.23E−06 2.79E−04 3.83E−04 4.46E−04 238Np 5.80E−03 5.00E−03 4.00E+00 1.37E−03 3.21E−08 5.19E−09 5.30E−10 239Np 6.45E−03 5.00E−03 3.26E+00 1.68E−03 9.44E−05 9.09E−05 8.68E−05 240Np 1.24E−04 5.00E−03 2.37E−01 2.31E−02 4.28E−15 4.28E−15 4.28E−15 240mNp 1.41E−05 5.00E−03 0.00E+00 NA 3.89E−12 3.89E−12 3.89E−12 231Pa 3.28E+04 5.50E−01 1.06E+04 5.17E−07 2.33E−07 6.66E−07 1.20E−06 233Pa 7.39E−02 5.50E−01 3.63E+00 1.51E−03 2.64E−06 3.61E−06 4.21E−06 234Pa 7.64E−04 5.50E−01 2.16E+00 2.54E−03 8.82E−10 8.82E−10 8.82E−10 234mPa 2.22E−06 5.50E−01 0.00E+00 NA 5.51E−07 5.51E−07 5.51E−07 209Pb 3.71E−04 2.70E−01 2.13E−01 2.57E−02 4.44E−10 1.33E−09 3.06E−09 210Pb 2.23E+01 2.70E−01 5.37E+03 1.02E−06 3.77E−07 2.93E−06 8.65E−06 211Pb 6.86E−05 2.70E−01 5.25E−01 1.04E−02 4.04E−07 1.29E−06 2.38E−06 212Pb 1.21E−03 2.70E−01 4.55E+01 1.20E−04 1.17E−07 1.96E−09 2.24E−11 214Pb 5.10E−05 2.70E−01 6.25E−01 8.77E−03 5.00E−07 3.22E−06 9.10E−06 107Pd 6.50E+06 5.50E−02 1.49E−01 3.68E−02 2.57E−05 2.57E−05 2.57E−05 146Pm 5.53E+00 0.00E+00 3.67E+00 1.49E−03 1.09E−09 1.92E−31 1.23E−58 147Pm 2.62E+00 0.00E+00 1.05E+00 5.22E−03 1.89E−13 2.39E−59 1.01E−116 210Po 3.79E−01 1.50E−01 1.90E+03 2.88E−06 6.75E−07 5.27E−06 1.55E−05 211Po 1.64E−08 1.50E−01 0.00E+00 NA 1.98E−09 6.32E−09 1.17E−08 212Po 9.44E−15 1.50E−01 0.00E+00 NA 1.35E−07 2.26E−09 2.58E−11 213Po 1.33E−13 1.50E−01 0.00E+00 NA 7.82E−10 2.35E−09 5.39E−09 214Po 2.02E−12 1.50E−01 0.00E+00 NA 9.00E−07 5.79E−06 1.64E−05 215Po 2.47E−11 1.50E−01 0.00E+00 NA 7.26E−07 2.31E−06 4.28E−06 216Po 4.63E−09 1.50E−01 0.00E+00 NA 2.11E−07 3.52E−09 4.02E−11 218Po 5.80E−06 1.50E−01 0.00E+00 NA 9.00E−07 5.79E−06 1.64E−05 236Pu 2.85E+00 5.50E−01 1.17E+03 4.68E−06 7.17E−24 4.18E−66 6.71E−119 238Pu 8.78E+01 5.50E−01 3.20E+03 1.71E−06 5.56E−03 2.36E−04 4.55E−06 239Pu 2.41E+04 5.50E−01 3.54E+03 1.55E−06 9.63E−03 9.52E−03 9.38E−03 240Pu 6.54E+03 5.50E−01 3.54E+03 1.55E−06 2.15E−03 2.06E−03 1.95E−03 241Pu 1.44E+01 5.50E−01 6.85E+01 8.00E−05 9.34E−05 3.02E−10 2.89E−10 242Pu 3.76E+05 5.50E−01 3.36E+03 1.63E−06 3.49E−07 3.48E−07 3.48E−07 243Pu 5.65E−04 5.50E−01 3.34E−01 1.64E−02 2.52E−18 2.52E−18 2.52E−18 244Pu 8.26E+07 5.50E−01 3.32E+03 1.65E−06 3.67E−14 3.67E−14 3.67E−14 223Ra 3.13E−02 5.00E−01 6.59E+02 8.31E−06 2.18E−07 6.95E−07 1.28E−06 224Ra 3.62E+00 5.00E−01 3.66E+02 1.50E−05 6.34E−08 1.06E−09 1.21E−11 
Table 3-3. (continued). 

3-31

Nuclide 

Half-life 

(yr) 

Kd

(m3

/kg) 

Ingestion 

DCF 

(mrem/μCi) 

Ciw, limit

Allowable 

Concentration

(μCi/L) 

100 yr 

Ciw,max 

(μCi/L) 

500 yr 

Ciw,max 

(μCi/L) 

1000 yr 

Ciw,max 

(μCi/L) 225Ra 4.05E−02 5.00E−01 3.85E+02 1.42E−05 2.40E−10 7.20E−10 1.65E−09 226Ra 1.60E+03 5.00E−01 1.32E+03 4.15E−06 2.70E−07 1.74E−06 4.92E−06 228Ra 5.75E+00 5.00E−01 1.44E+03 3.81E−06 9.14E−09 1.76E−12 3.52E−12 87Rb 4.73E+10 5.50E−02 4.92E+00 1.11E−03 4.70E−08 4.70E−08 4.70E−08 

219Rn 1.25E−07 0 0.00E+00 NA 5.65E−04 1.80E−03 3.33E−03 220Rn 1.76E−06 0 0.00E+00 NA 1.64E−04 2.74E−06 3.13E−08 222Rn 1.05E−02 0 0.00E+00 NA 7.01E−04 4.51E−03 1.27E−02 126Sb 3.40E−02 4.50E−02 1.07E+01 5.12E−04 1.72E−04 1.72E−04 1.71E−04 126mSb 3.62E−05 4.50E−02 9.36E−02 5.85E−02 1.23E−03 1.23E−03 1.22E−03 79Se 6.50E+04 1.50E−01 8.70E+00 6.30E−04 4.49E−04 4.47E−04 4.44E−04 146Sm 7.00E+07 2.45E−01 2.04E+02 2.69E−05 5.19E−12 5.18E−12 5.18E−12 147Sm 6.90E+09 2.45E−01 1.85E+02 2.96E−05 3.48E−08 3.48E−08 3.48E−08 151Sm 9.00E+01 2.45E−01 3.89E−01 1.41E−02 2.40E−01 1.10E−02 2.34E−04 121mSn 7.60E+01 1.30E−01 1.55E+00 3.54E−03 2.14E−05 5.58E−07 5.84E−09 126Sn 1.00E+05 1.30E−01 1.95E+01 2.81E−04 4.26E−04 4.25E−04 4.24E−04 90Sr 2.86E+01 1.50E−02 1.42E+02 3.86E−05 4.44E+00 2.74E−04 1.50E−09 98Tc 4.20E+06 1.00E−04 4.88E+00 1.12E−03 7.12E−19 7.12E−19 7.12E−19 99Tc 2.13E+05 1.00E−04 1.46E+00 3.75E−03 2.33E−11 2.32E−11 2.32E−11 123Te 1.00E+13 1.25E−01 4.18E+00 1.31E−03 1.33E−17 1.33E−17 1.33E−17 227Th 5.12E−02 3.20E+00 3.81E+01 1.44E−04 3.36E−08 1.07E−07 1.98E−07 228Th 1.91E+00 3.20E+00 3.96E+02 1.38E−05 9.90E−09 1.65E−10 1.89E−12 229Th 7.34E+03 3.20E+00 3.53E+03 1.55E−06 3.75E−11 1.13E−10 2.59E−10 230Th 7.70E+04 3.20E+00 5.48E+02 1.00E−05 8.75E−07 2.06E−06 3.53E−06 231Th 2.91E−03 3.20E+00 1.35E+00 4.06E−03 8.87E−06 8.87E−06 8.87E−06 232Th 1.41E+10 3.20E+00 2.73E+03 2.01E−06 5.97E−14 2.80E−13 5.55E−13 234Th 6.60E−02 3.20E+00 1.37E+01 4.00E−04 9.48E−08 9.48E−08 9.48E−08 207Tl 9.07E−06 0 0.00E+00 NA 5.64E−04 1.80E−03 3.32E−03 208Tl 5.81E−06 0 0.00E+00 NA 5.90E−05 9.85E−07 1.13E−08 209Tl 4.18E−06 0 0.00E+00 NA 1.34E−08 4.03E−08 9.26E−08 210Tl 2.47E−06 0 0 NA 0.00E+00 0.00E+00 0.00E+00 232U 7.20E+01 3.50E−02 1.31E+03 4.18E−06 6.87E−07 1.46E−08 1.19E−10 233U 1.59E+05 3.50E−02 2.89E+02 1.90E−05 1.46E−07 2.32E−07 3.66E−07 234U 2.45E+05 3.50E−02 2.83E+02 1.94E−05 3.00E−02 3.00E−02 3.00E−02 235U 7.04E+08 3.50E−02 2.66E+02 2.06E−05 8.07E−04 8.07E−04 8.07E−04 236U 3.42E+06 3.50E−02 2.69E+02 2.04E−05 1.01E−03 1.01E−03 1.01E−03 237U 1.85E−02 3.50E−02 3.17E+00 1.73E−03 3.58E−08 1.16E−13 1.11E−13 238U 4.47E+09 3.50E−02 2.55E+02 2.15E−05 8.62E−06 8.62E−06 8.62E−06 240U 1.61E−03 3.50E−02 4.44E+00 1.23E−03 5.74E−13 5.74E−13 5.74E−13 90Y 7.32E−03 1.70E−01 1.08E+01 5.07E−04 3.97E−01 2.44E−05 1.34E−10 93Zr 1.53E+06 6.00E−01 1.66E+00 3.30E−03 1.19E−03 1.19E−03 1.19E−03 

a. Allowable concentration not calculated since ingestion DCF not reported for these nuclides in EPA 1988. 
3-32

Table 3-4. Radionuclides remaining from the initial screening based on waste pore-water concentrations. 

Nuclide

Half-life 

(yr) Nuclide

Half-life 

(yr) Nuclide

Half-life 

(yr) 

241Am 4.32E+02 94Nb 2.03E+04 224Ra 1.00E−02 

242mAm 1.52E+02 63Ni 1.00E+02 226Rab

 1.60E+03 

243Am 7.38E+03 237Np 2.14E+06 79Se 6.50E+04 

14C 5.73E+03 210Pb 2.23E+01 151Sm 9.00E+01 

113mCd 1.37E+01 210Po 3.79E−01 126Sn+Da

 1.00E+05 

244Cm 1.81E+01 238Pu 8.78E+01 90Sr+Da

 2.86E+01 

135Cs 2.30E+06 239Pu 2.41E+04 99Tc 2.13E+05 

137Cs+Da

 3.02E+01 240Pu 6.54E+03 234U 2.45E+05 

3

H 1.23E+01 241Pu 1.41E+01 236U 3.42E+06 

129I 1.57E+07 242Pu 3.76E+05 

a. +D indicate there is a daughter to consider in the groundwater and dose analysis. 

b. 226Ra is included for the groundwater analysis because it will provide activity to 210Pb during transport and decay. 

were at the center of the 1-ft layer of grout in the 

tank (i.e., the bottom 1 ft of the grouted tank was 

conservatively assumed to contain all of the tank 

contamination) and at the center of the sand pad. 

Radionuclide inventories used to determine 

the DUST-MS release rates for the tank and sand 

pad are provided in Section 2.3. Justification for 

the times required for vault degradation and tank 

corrosion (i.e., 100 and 500 yr, respectively) is 

summarized in Section 3.1.2.2 and presented 

completely in Appendix E. Additional DUST-MS 

input parameters are discussed subsequently in this 

section. 

The release rates computed in the DUST-MS 

simulations were estimated at the location of the 

vault floor/basalt interface. These release rates 

were used in the PORFLOW unsaturated/saturated 

zone transport simulations discussed in 

Section 3.2. The following sections present a more 

detailed discussion of the DUST-MS model, the 

modeling approach, the input parameters, and 

verification of DUST-MS model results. 

3.3.1.1 DUST-MS Model Description. 

DUST-MS considers four major processes: fluid 

flow, container degradation, waste form leaching, 

and contaminant transport. The DUST-MS model 

permits selection of a unique failure time and 

waste form type for each container. To simulate 

different waste forms, DUST-MS has four 

different models to estimate release rates: rinse 

with partitioning, diffusion, uniform degradation, 

and solubility-limited release. For these four 

processes, DUST-MS computes release rates with 

an analytical model or a finite difference model. 

After calculating waste form releases, the 

movement of the contaminants through subsurface 

system materials is determined using a onedimensional finite difference procedure with 

material- and contaminant-specific distribution 

coefficients, diffusion coefficients, dispersion 

coefficients, and bulk densities. The governing 

transport equation simulates the distribution and 

movement of contaminants from advection, 

dispersion, diffusion, radioactive production and 

decay, sorption, and sources and sinks external to 

the containers. The DUST-MS code has received 

extensive testing and verification. DUST-MS code 

predictions compare favorably to known analytical 

solutions as well as other code predictions.

3.3.1.2 Conceptual Model. The conceptual 

model for computing release rates of radionuclides 

out of the TFF vault is illustrated in Figure 3-14. 

Infiltrating water from the ground surface contacts 

the waste inside the vaults. As shown in 
3-33

Figure 3-14. DUST-MS conceptual model for release from the TFF vaults and tanks. 

Figure 3-14, radionuclides are assumed to be 

located in the grouted tank heel and in the sand 

pad beneath the tank. The release of radionuclides 

from the grouted tank heel and from the sand pad 

is modeled in DUST-MS assuming surface 

rinsing. The vault is assumed to remain intact for 

100 yr. The tanks are assumed to remain intact for 

500 yr. Upon vault and tank failure, radionuclides 

are released from the two sources. 

The DUST-MS code calculates release rates 

out of the vault and tanks into the unsaturated 

zone. Release rates are computed for transport 

through the grout inside the tank (0.15 m [5.9 in.]), 

the sand pad (0.15 m [5.9 in.]), and the degraded 

concrete (0.76 m [30. in.]). Retardation is assumed 

to occur in the waste release model (surface 

rinsing), in the grout inside the tank, in the sand 

pad, and in the degraded concrete. 

3.3.1.3 Model Input Parameters. Input 

parameter values that describe source geometry, 

infiltration rates, failure times, dispersion 

coefficients, and computational options used in all 

DUST-MS simulations are included in Table 3-5. 

Vadose zone transport properties (infiltration and 

moisture content) also are listed in Table 3-5. 

Table 3-6 lists the modeled radionuclides; 

half-lives; partition coefficients (Kd values) for the 

sand pad, liquid tank waste, solid tank waste, and 

concrete vault; and sand pad liquid tank waste and 

solid tank waste initial inventories. The 

distribution coefficients for grout were taken from 

studies conducted by Allard, Persson, and 

Torstenflt (1985) and Bradbury and Sarott (1995). 

The distribution coefficients for sand were taken 

from Sheppard and Thibault (1990) (see 

Appendix A). 

3.3.1.4 Verification of DUST-MS Model 

Results. The PORFLOW computer code was 

initially used to estimate tank and sand pad release 

rates. However, because of the numericallyintensive calculations associated with grid 

discretization and modeling the entire source term 

and unsaturated/saturated model domain, the 

DUST-MS computer code was used. The DUSTMS computer code minimized the computational 

time required in PORFLOW simulations to
3-34

Table 3-5. Parameter values for source term analysis using the DUST-MS model. 

Parameter Value Comments 

Number of finite difference nodes 200 Reasonable number 

Spatial discretization (cm) 1.3 Reasonable number 

Geometry of grouted tank heel Cylindrical Based on tank geometry 

Radius of tank heel (cm) 762 Based on tank geometry 

Failure time of vault (yr) 100 Based on degradation analysis 

Failure time of tank (yr) 500 Based on degradation analysis 

Geometry of underlying sand Cylindrical Based on tank geometry 

Radius of underlying sand (cm) 762 Based on tank geometry 

Dispersion coefficient (cm) 1.2 Reasonable value 

Thickness of grouted heel (m) 0.30 Reasonable value 

Transport distance in tank (m) 0.15 One-half total thickness 

Thickness of sand pad (m) 0.15 Reasonable value 

Transport distance in sand pad (m) 0.075 One-half total thickness 

Thickness of vault floor (m) 0.76 Average value 

Boundary condition top No flux Reasonable condition 

Boundary condition bottom Zero concentration Reasonable condition 

Initial time interval (yr) 1.0 Reasonable step 

Fractional change in time interval (yr) 0.001 Reasonable step 

Time step maximum (yr) 1000.0 Reasonable step 

Number of time steps 10,000 Reasonable value 

Maximum simulation time (yr) 1.0E+06 Value computing peaks 

Darcy velocity (cm/s) 4.0E−07 Based on vadose zone flow modeling 

Grouted tank moisture content (by volume) 0.242 Average based on vadose zone modeling 

Sand pad moisture content (by volume) 0.05 Average based on vadose zone modeling 

Vault concrete moisture content (by 

volume) 

0.282 Average based on vadose zone modeling 

Grout density (g/cm3) 2.12 Reasonable value 

Sand pad density (g/cm3

) 1.75 Reasonable value 

Concrete vault density (g/cm3) 2.12 Reasonable value 

Location of release rate output Bottom of vault Source location for vadose and saturated zone modeling 

Table 3-6. Non-decay chain DUST-MS input parameters. 

Nuclide 

Half-life 

(yr) 

Kd

Sand Pad 

(m3

/kg) 

Kd

Grout/Concrete 

(m3

/kg) 

14C 5.73E+03 5.00E−03 5.00E+00 

90Sr 2.86E+01 1.50E−02 3.00E−03 

99Tc 2.13E+05 1.00E−04 2.50E+00 
3-35

calculate release rates from the tank and the sand 

pad. 

To verify DUST-MS release rates, results 

from PORFLOW simulations for selected 

radionuclides were compared to DUST-MS 

results. Release rates, computed with DUST-MS 

using Table 3-5 input parameters and with 

PORFLOW yielded similar predictions and 

justified the use of the DUST-MS computer code 

with these input parameters to estimate release 

rates from the grout tank and the sand pad. These 

results are discussed in Appendix F. In addition, 

Section 7 presents the sensitivity/uncertainty 

analysis for the analyses. 

3.3.2 Groundwater Flow and Transport 

Modeling 

A two-dimensional unsaturated/saturated 

PORFLOW model was used to simulate water and 

contaminant transport in the subsurface at the 

INTEC facility (see Figure 3-15). By using a twodimensional model, it was possible to digitize the 

problem domain using a detailed 250 × 103 node 

grid. This approach allowed a detailed 

approximation of the complex geology underlying 

the facility as discussed in Section 3.3.2.2.1 at 

reasonable computational speeds. The twodimension problem domain represented a vertical 

slice in a north-south direction beginning at the 

Big Lost River, through the center of two tank 

vaults, and southward in the downgradient 

direction for a total distance of 2500 m (8,200 ft). 

The top of the model is located at land surface and 

extends to a depth of 200 m (700 ft)—well below 

the top of the water table located at 134 to 139 m 

(440. to 456 ft) bgs. A variable grid was used with 

a high density of nodes in the tank vault area 

where complex construction and highly variable 

hydraulic conductivities for the various materials 

exist. By default, the unit thickness of the twodimensional model in an east-west direction is 1 m 

(3 ft).

Figure 3-15. Illustration of the two-dimensional modeling slice used in PORFLOW. 
3-36

For the top of the model, a constant flux 

boundary was set at the infiltration rate based on 

precipitation in the area. For simulations where 

seepage from the Big Lost River is incorporated, 

two nodes located in the uppermost northern edge 

of the model at the location of the Big Lost River 

were injected with water. The injection rate is 

equal to the average seepage lost per meter of 

river, based on historical gaging data presented in 

Bennett (1990). The seepage loss rate is consistent 

with the unit thickness (1 m [3 ft]) of the model. 

The base of the model is set as a no flow 

boundary. This is based on the assumption of 

horizontal flow in the regional aquifer and 

nonexistent future pumping of the aquifer for 

water supply. 

The sides of the model are set as mixed 

boundaries. In the unsaturated zone, the 

boundaries are set as no flow. The boundaries in 

the unsaturated zone are located at sufficient 

distances so as to not impact movement of water 

from the tank vault area. Constant head boundary 

conditions are used in the saturated portion of the 

model based on the regional potentiometric map at 

the location of the boundary. 

Initial conditions are discussed in detail in 

Section 3.3.2.4 since different modeling 

simulations used different initial conditions. For 

the first simulation, initial conditions for steadystate runs of the base-case model assumed 

unsaturated conditions except in the regional 

aquifer, recharge rate equal to precipitation, no 

Big Lost seepage, and no perched water zones. 

For contaminant transport simulations, the 

initial contaminant concentrations in the model 

domain and the boundaries are set equal to zero. 

Consequently, all predicted contaminant 

concentrations in the model domain are the result 

of releases from the Tank Farm vaults and 

associated piping. This is a reasonable assumption 

for the unsaturated zone considering that the 

presence of existing contamination, particularly 

for future estimates, is minimal in the region based 

on existing sampling data. 

For the source area releases, the conceptual 

model of the tank vault included initial hydraulic 

conductivity values that represented the materials 

of the vaults including grout, concrete, and steel. 

At selected time periods, based on degradation 

studies presented in Section 3.1.2, the hydraulic 

conductivities were changed to the degraded 

values. Contaminants were released as a result of 

the higher hydraulic conductivity values for the 

tank vault materials. 

Because of the large contrast in hydraulic 

conductivity values for the tank vaults compared 

with the surrounding geologic material, it was 

necessary to use small time steps during the 

simulations to minimize computational errors. 

This resulted in extensive run times for the model 

simulations. To reduce run times because of the 

numerous simulations for both the transport 

predictions and the subsequent uncertainty 

analysis, the model DUST-MS, as detailed in 

Section 3.3.1, was used to compute release rates of 

radionuclides out of the TFF vault. As discussed in 

Section 3.3.1.4, a comparative analysis between 

DUST-MS and PORFLOW was conducted. The 

DUST-MS and PORFLOW transport simulation 

results compared favorably. 

For contaminant transport simulations, the 

calibrated unsaturated and saturated flow model 

was used to set the transport velocity vector field. 

All transport simulations were conducted under 

steady-state conditions. It was assumed that the 

percolation ponds had been taken out of service 

and the resulting percolation was no longer a 

factor in the transport of water in the subsurface. 

Only steady infiltration from precipitation and 

seepage from the Big Lost River, at rates 

discussed in the following sections, were 

incorporated into the water transport model. 

As noted earlier, DUST-MS was used to 

predict the release rates from the various waste 

forms. Radionuclide releases coincided with 

degradation rates of the waste forms discussed in 

Section 3.1.2.2. Release rates calculated by 

DUST-MS were incorporated as input in 

PORFLOW, with the initial radionuclide releases 

beginning at the time of degradation. For example, 

the outer vault began to release radionuclides after 

100 yr and the tanks/piping after 500 yr. These 

assumptions are conservative, based on the best 

estimates provided from the degradation analysis. 
3-37

Because of the small region of the two tanks 

simulated in the 2-D model compared with the 

large model domain, releases from only one tank 

were simulated. Radionuclide concentrations 

downgradient were subsequently doubled to 

incorporate releases from both tanks. Earlier 

simulations conducted using releases from both 

tanks yielded similar results. 

For initial conditions, the radionuclide 

concentrations were set equal to zero. All model 

boundaries also were set with initial radionuclide 

concentrations of zero. Contaminants were 

released from nodes that coincided with the 

location of the piping, sand pad, vaults, and tanks. 

Transport simulations were conducted for a time 

period of 1,000 yr. 

An important component of the conceptual 

model for the transport of contaminants is the 

relationship of dispersion in the two-dimensional 

model. Since the model uses a unit thickness of 

1 m, there is no lateral dispersion of contaminants 

beyond this thickness. It is assumed that 

contaminants disperse or diffuse out of the lateral 

boundary at the same rate as contaminants move 

into the unit thickness of the model domain. 

Essentially, all dispersion of contaminants occurs 

in the longitudinal direction, with zero transverse 

dispersion. This is a conservative approach to 

predicting downgradient contaminant 

concentrations. 

Since the model domain is located in the 

center of the tanks, the highest contaminant 

concentrations are located in the source area of the 

model. Although it is assumed that the same 

amount of contaminants disperse in a transverse 

direction out of, as into the model domain, in 

reality there would be the same loss of 

contaminants, albeit small, in the transverse 

direction. Consequently, downgradient 

contaminant concentrations will be slightly higher 

for the two-dimensional simulation compared to a 

fully three-dimensional simulation. 

Radionuclide concentrations were observed 

downgradient in the regional aquifer at the 

location where the highest concentrations occur. 

During unsaturated flow simulations, it was 

observed that the perched water zones deflected 

the contamination. Consequently, the usual 100-m 

downgradient location was inappropriate to 

quantify the maximum impacts of radionuclide 

releases from the tanks and vaults. Based on the 

radionuclide concentrations distribution, the 

maximum observed concentrations occur 

approximately 600 m downgradient from the 

tanks. 

3.3.2.1 Numerical Model Selection. The 

PORFLOW code (Runchal 1997; ACRi 2000) was 

selected to perform the proposed fluid and 

contaminant transport simulations for the TFF PA. 

PORFLOW was selected based on the following 

criteria: 

• Application to the multiphase contaminant 

transport problem at INTEC 

• Model validation 

• A history of applications similar to the model 

simulations to be performed at INTEC 

facilities. 

PORFLOW is a comprehensive mathematical 

model used for the simulation of multi-phase fluid 

flow, heat transfer, and mass transport processes in 

variably saturated porous and fractured media. The 

code can simulate transient or steady-state 

problems in one, two, or three dimensions using 

either Cartesian or cylindrical geometry. The 

geologic medium may be anisotropic and 

heterogeneous and may contain distinct embedded 

elements such as discrete fractures or boreholes 

within a porous matrix. In partially saturated 

zones, liquids and gases are assumed to co-exist. 

The degree of saturation of each phase is 

determined at each grid node as part of the 

solution. The dependent variable, or its change 

from the current state, approximates the flux 

terms. Finally, several options are provided for the 

incorporation of sources or sinks of fluid, heat, or 

mass. Fluid injection or withdrawal, sources or 

sinks of heat, or chemical species may occur 

anywhere in the interior of the domain of interests. 

For chemical species, the sources can be limited 

by their inventory, solubility, or both. 

PORFLOW numerically solves a variable set 

of equations for general transport, multi-phase 
3-38

pressure, temperature, and one or more chemical 

species. The method of nodal point integration is 

used to integrate the governing differential 

equations by temporal and spatial discretization 

over each control volume (element) of the physical 

domain. It leads to solutions that automatically 

conserve fluid, heat, and mass locally within every 

grid element, as well as for the entire flow domain. 

The storage terms are approximated by a modified 

Newton-Raphson method (Runchal 1997). 

The PORFLOW code is particularly well 

suited for simulating the transport of fluid and 

contaminants from the Tank Farm waste forms, 

through the underlying unsaturated zone, and into 

the regional aquifer. PORFLOW can be used to 

simulate the release of contaminants from the 

waste form and the subsequent transport through 

the underlying geological media. Because of the 

number of simulations conducted for the numerous 

radionuclides discussed in Section 4, time 

restraints required the use of DUST-MS for 

computing the release of radionuclides from the 

waste forms. PORFLOW was designed especially 

for problems of fluid and mass transport in 

geologic media. 

Several features of PORFLOW are especially 

important regarding the proposed modeling 

simulations for the INTEC area. First, data exist 

for the waste inventory and solubility of the 

contaminants that will be used as part of the PA. 

This information can be input to PORFLOW to 

estimate the contaminant source emanating from 

the Tank Farm waste forms. These sources terms 

are automatically input into the underlying 

unsaturated zone, transported to the regional 

aquifer based on hydraulic and retardation 

characteristics of the unsaturated zone, and then 

transported by the regional aquifer. PORFLOW 

can simulate transport under partially or 

completely saturated conditions. Second, the 

geologic media underlying the facility consists of 

fractured basalts and unconsolidated porous 

media. PORFLOW incorporates the planar 

geologic features, such as fractures or faults, with 

different length scales and properties to distinguish 

these features from the parent geologic media. 

This feature will be useful for simulating flow in 

the fractured basalt instead of assuming a 

representative hydraulic conductivity for the entire 

basalt matrix. Finally, PORFLOW can calculate 

contaminant flux values at any location in the 

problem domain. This feature allows the user to 

assess the flux of contaminants from the waste 

forms, the amount of contaminants that are being 

held in the unsaturated zones, and the flux of 

contaminants entering the regional aquifer. 

PORFLOW can simulate the decay of a 

radionuclide and up to three daughters in a 

conservative manner. Using the law of radioactive 

disintegration and the half-life input by the user, 

PORFLOW computes the remaining 

concentrations of the parent radionuclides and the 

subsequent increase and/or decrease in the 

daughters’ concentrations. The code maintains a 

mass balance of the initial radionuclide and the 

subsequent concentrations of the parent and 

daughters at selected time intervals. 

PORFLOW has been extensively peerreviewed (Runchal 1997). INEEL, Battelle Pacific 

Northwest, and Professor R. Allan Freeze of the 

University of British Columbia have formally 

reviewed PORFLOW and its derivatives. 

Additionally, it has been reviewed by ANDRA 

(France), Bae-SEMA (United Kingdom), British 

Petroleum (United Kingdom), Exxon Production 

Research, Failure Analysis Associates Inc., Fluor 

Daniels Inc., Gaz de France (France), SAIC, Shell 

Oil, SOHIO, and Westinghouse Hanford 

Company. Over 100 publications and project 

reports on the benchmarking, verification, and 

applications of PORFLOW are currently available. 

PORFLOW has a history of use specific to the 

PA presented in this report. The code was 

successfully used for the Radiological 

Performance Assessment for the Z-Area Saltstone 

Facility and E-Area at the Savannah River Site 

(Westinghouse Savannah River Company 1992). 

This use of PORFLOW is very similar to the 

proposed application at the INEEL Site. 

PORFLOW also was used at RWMC at the 

INEEL, near the INTEC area (Maheras et al. 

1994). 

In summary, PORFLOW is a robust 

simulation code that has a direct application to the 

multi-phase contaminant transport problem at 

INTEC. The code has been extensively validated 
3-39

by independent researchers and has been used to 

model similar hydrogeologic systems at other 

DOE facilities.

3.3.2.2 General Modeling Assumptions. 

Based on a review of the available technical 

information for INTEC and the surrounding sites 

at the INEEL, three general conclusions can be 

drawn regarding the ability to accurately simulate 

hydrologic conditions underlying INTEC and 

subsequent predictive estimates relative to the 

necessary assumptions. The first conclusion is that 

there is a large amount of data available from 

previous investigations. These data provide 

important information about the stratigraphy and 

hydrologic parameters that control the subsurface 

movement of water and contaminants. Conversely, 

the second conclusion, based on a thorough review 

of the available technical data, is that there is a 

large degree of uncertainty regarding this 

information. The uncertainty is due to conflicting 

information from the available technical 

references, such as large ranges in the values of 

certain hydraulic parameters, conflicting 

lithologies, and the selection of appropriate 

technical data that results in a bias relative to the 

hydrologic properties representing certain 

subsurface conditions. The final conclusion is that 

there is a significant lack of important technical 

data necessary to reduce the uncertainty of model 

simulations for existing hydrologic conditions and 

future predictions. The uncertainties associated 

with the available technical data and the lack of 

important data will be discussed in detail in the 

following sections.

These three conclusions required certain 

assumptions to provide a reasonable model of 

present and future hydrologic conditions at 

INTEC. Once again, these assumptions will be 

discussed in detail in the following sections. In 

general terms, however, the following rules were 

used as guidelines for selecting the assumptions 

inherent in the model simulation performed as part 

of this PA. First, where appropriate, existing 

hydrologic data were used extensively to describe 

subsurface hydraulic conditions. For example, the 

distribution of basalt units and interbeds closely 

followed the geologic cross sections presented in 

several USGS reports (Anderson 1991; Anderson 

and Bartholomay 1995; Anderson, Ackerman, and 

Liszewski 1996; Anderson, Liszewski, and Cecil 

1997; Anderson, Kuntz, and Davis 1999). Second, 

because of the conflicting nature of the data from 

several references, attempts were made to confirm 

the data used in the model from different sources. 

For example, if the drilling logs indicate that a 

selected interbed consists of both fine and coarse 

grain sediments, separate hydraulic information 

such as the hydraulic conductivity or unsaturated 

hydraulic characteristic was evaluated to confirm 

the lithologic information. Finally, if there is no 

available information to confirm lithologic or 

hydraulic data, then a conservative approach was 

adapted that resulted in using the value that 

yielded the highest transport rates. For example, if 

the only available information for a particular 

geologic unit is a range in hydraulic conductivity 

values, then the upper portion of the range is used 

in the model. 

3.3.2.2.1 Subsurface Geology 

Assumptions—As noted in the geology section, 

the geology underlying INTEC consists of a series 

of basalt flows and sedimentary interbeds with an 

alluvial veneer. There are over 30 individual 

geologic units that compose the unsaturated zone 

and upper regional aquifer at the site. Previous 

modeling programs at the INEEL (Thomas 1988; 

Maheras et al. 1997; Rodriguez et al. 1997) 

simplified the subsurface geology to perform 

three-dimensional modeling. It was assumed for 

this PA that a two-dimensional model can be used 

to simulate the movement of water and 

contaminants in the subsurface. The advantage of 

the two-dimensional model is that greater detail is 

possible for defining the subsurface features.

The PA model is based on a USGS cross 

section of the subsurface geology at INTEC 

presented by Anderson (1991) and shown in 

Figure 2-12. This detailed geologic cross section 

shows the alluvium, 18 basalt flows, and nine 

continuous and discontinuous sedimentary 

interbeds. With the exception of basalt flows E 

through I, which form the upper regional aquifer 

and were combined into a single hydrologic unit, 

all of the remaining geologic units are represented 

as individual units within the problem domain. 

The USGS cross section runs north-south 

across INTEC and is aligned to parallel the 
3-40

regional groundwater flow direction. Hydraulic 

properties are based on the lithologic descriptions 

and published values and are assigned to the 

individual units in accordance with their locations. 

3.3.2.2.2 Infiltration Assumptions—

A fundamental assumption for simulating the 

perched groundwater zones discussed in the 

subsequent section is that the amount of 

infiltrating water, either from precipitation, the Big 

Lost River, or human activities, is based on 

reasonable values from the site literature. Surface 

sources are set at a maximum reasonable value 

regardless of the resultant formation of perched 

zones. The controlling factor for the development 

of perched groundwater zones will be the vertical 

permeability contrasts in subsurface units that 

result in perched water.

3.3.2.2.3 Perched Groundwater 

Assumptions—The mechanisms surrounding 

the hydrologic and hydraulic factors that control 

the perched water zones underlying INTEC are not 

well understood and remain controversial. The 

source of water causing the perched zones and the 

subsurface units responsible for restricting the 

downward movement of infiltrating water are not 

fully understood. Further, it is not known if 

perched water zones existed before site activities 

began. Finally, there is disagreement on the 

subsurface units responsible for restricting the 

downward movement of infiltrating water. Most 

scientific studies state that fine-grain interbeds 

form the base of the perched water zones. 

Conversely, the interbeds and highly fractured 

basalts adjacent to the interbed zones commonly 

are the highest permeable regions in basalt flows. 

Cecil et al. (1991) state that lithologic features 

contributing to the contrasts in the vertical 

hydraulic conductivity of basalt layers and 

sedimentary interbeds in the unsaturated zone 

provide the mechanisms for the development of 

perched groundwater zones.

Perched water at TRA, INTEC, and RWMC 

can be attributed to at least four lithologic features. 

First, the vertical hydraulic conductivity of a 

sedimentary interbed might be smaller than that of 

an overlying basalt flow. Second, alterations in the 

baked zone between two flows can contribute to 

reduced vertical hydraulic conductivity. Third, 

dense, unfractured basalt may inhibit unsaturated 

groundwater movement, contributing to the 

formation of perched groundwater zones. Fourth, 

sedimentary and chemical filling of fractures near 

the upper contact of a basalt flow reduces the 

vertical hydraulic conductivity. Selecting which of 

these mechanisms is responsible for the observed 

perched zones is an important assumption in the 

model simulation. 

The assumptions involving the perched zones 

are perhaps the most important in developing and 

interpreting the numerical model. They control the 

amount of water input into the system as well as 

the hydraulic characteristics of the underlying 

formation. All model results should be viewed in 

light of these initial assumptions. 

For the initial phase of the model, it is 

assumed that there are no perched water zones due 

solely to infiltrating water from precipitation. This 

assumption is supported by age-dating studies 

using 36Cl ratios conducted by Cecil et al. (1992). 

The pore water from soil cores was extracted and 

analyzed for 36Cl. 36Cl was produced during 

nuclear weapons tests conducted over the oceans 

from 1952 to 1958. In the subsequent 40 yr, 

infiltrating water from precipitation containing the 

higher concentrations of 36Cl has only penetrated 

to a depth of 1.3 m (4.3 ft). Based on this 

evidence, it appears unlikely that precipitation is a 

significant source of recharge for the perched 

groundwater underlying the site. 

The best estimate of the subsurface 

stratigraphy, hydraulic parameters, and volume of 

infiltrating water from the various sources initially 

was coded into the model. Subsurface stratigraphy 

and the volume of infiltrating water remained 

constant during the model simulations. For 

subsequent simulations involving seepage losses 

from the Big Lost River and the percolation ponds, 

it was assumed that the development of perched 

water zones was due to variations in the vertical 

hydraulic conductivity of selected subsurface 

units. Consequently, the development of perched 

zones and the thickness and aerial extent of 

perched water are determined by (a) comparing 

the model results with the existing hydrologic 

information and (b) adjusting the vertical 

permeability of the appropriate subsurface unit 
3-41

with measured perched water to match model 

predictions with actual conditions. 

Based on the lithologic features presented by 

Cecil et al. (1991) and discussed in detail in 

Section 2.1.5.4, sedimentary interbeds in the 

unsaturated zone exhibit lower permeability and 

restrict the movement of infiltrating water. 

Selection of the appropriate interbeds for reduced 

permeability and subsequent formation of perched 

water zones was based on the known location of 

perched water at the site in relation to the 

individual interbeds and the limited amount of 

lithologic and hydraulic information available. For 

example, 12 soil borings in the C-D interbed 

reported fine-grain material as the dominant 

lithology for this unit. Hydraulic data indicated 

high residual moisture content indicative of finegrain material. And, finally, wells completed at a 

similar depth as the C-D interbed reported perched 

water. Consequently, this unit was assigned a 

lower permeability consistent with the lithology 

that was sufficient to form perched water zones. 

A similar approach was taken with assigning 

permeability values for the other sedimentary 

interbeds. For the lower perched zones, there was 

a limited amount of lithologic and hydraulic 

information available. Therefore, the location of 

perched water zones, as defined by the available 

monitoring wells, was the primary factor for 

determining the vertical permeability. 

3.3.2.2.4 Hydraulic Transport 

Parameter Estimations—Transport parameter 

estimations involve the hydrodynamic dispersion 

and molecular diffusion terms in the transport 

equation. Both terms can have a significant effect, 

particularly in the unsaturated zone where flux 

rates may be low but pore water velocities may be 

significant. Relatively high dispersion and 

diffusion terms can result in significant transport 

and spreading of contaminants that may or may 

not be representative of actual contaminant 

transport conditions.

The one-dimensional form of the advectivedispersive equation for nonreactive dissolved 

containment under steady-state conditions is 

t

C

l

C

l

C Dl l ∂∂ = ∂∂ − ν ∂∂

2

2

 (3-12) 

where 

Dl = coefficient of hydrodynamic 

dispersion in the longitudinal 

direction 

C = solute concentration 

l = curvilinear coordinate direction 

νl = average linear velocity. 

The coefficient of hydrodynamic dispersion 

can be expressed in terms of two components: 

* Dl = α l

ν + D (3-13)

where 

αl = dynamic dispersion or dispersivity 

v = velocity 

D* = coefficient of molecular diffusion for 

the solute in a porous medium 

(Freeze and Cherry 1979). 

The hydrodynamic dispersion is dependent on 

the pore water velocity as illustrated in 

Equation (3-13). In the unsaturated zone, low 

moisture contents can result in a low effective 

porosity, resulting in a relatively high pore-water 

velocity. When these high pore-water velocities 

are multiplied by the dispersivity, the net result 

can yield unrealistically high transport rates in the 

unsaturated zone. 

Previous modeling studies at the site set the 

dispersivity values to zero to yield a conservative 

estimate of the maximum concentrations of 

contaminants (Maheras et al. 1994). This approach 

was based on the limited availability of data on 

unsaturated water movement and virtually no data 

on dispersivity values in the unsaturated zone. 

In evaluating the best estimate of dispersivity 

values for the unsaturated zone modeling, several 

factors were incorporated into the analysis. The 
3-42

transport scale is on the order of 130 m for 

transport through the unsaturated zone to the 

underlying water table. There are 20 separate 

zones that are assigned a specific hydraulic 

conductivity value that differs from the adjacent 

zones. The variations in hydraulic conductivity 

inherently impart dispersion on the transport of 

solutes. The scale of transport in the individual 

geologic units or zones is on the order of a few to 

tens of meters. As shown by Mills et al. (1985), 

the scale of the transport distances for the 

individual stratigraphic units results in a range of 

longitudinal dispersivity of 0.01 to 1.0 m. The 

results of an extensive review of the literature for 

measured dispersivity values in porous and 

fractured material are presented in Section 7. The 

range in measured dispersivity values is 

significantly less than the calculated site values for 

the Snake River Plain Aquifer. In addition, the 

measurement scale for the literature values is 

closer to the scale of transport for the individual 

geologic units in the unsaturated zone underlying 

the site. 

Longitudinal dispersivity values of 0.29 m for 

sediments and 1.85 m for basalts were assigned for 

transport simulations. Transverse dispersivities 

were set at one half the value of the longitudinal 

dispersivity. This approach minimizes the amount 

of dispersion in the unsaturated zone resulting in 

conservative estimates of radionuclide 

concentrations but allows the contaminant plumes 

to spread during transport. 

3.3.2.3 Numerical Model Description and 

Parameterization 

3.3.2.3.1 Alluvium—The upper alluvial 

unit is consistently identified as a coarse-grain unit 

consisting of predominantly sand and gravel. 

Lithologic logs from CPP-2 and CPP-3, along with 

the PW series wells (PW-1 through PW-6; 

Figure 3-19) and numerous USGS wells, all show 

the coarse-grain nature of the alluvium. 

Consequently, the hydraulic parameters in the 

model are set to reflect the nature of the coarsegrain deposits. Unfortunately, there is little 

technical data available for the hydraulic 

properties because of the unsaturated nature of the 

alluvium. The average hydraulic conductivity for 

the alluvial sediments, as reported by Thomas 

(1988), indicates a relatively low value of 

0.16 m/d (0.52 ft/d). Based on the lithologic logs, 

this value appears to be low. Therefore, adapting a 

conservative approach, the hydraulic conductivity 

value for the alluvium was increased to 80 m/d 

(300 ft/d), a value consistent with the lithologic 

descriptions as described by Freeze and Cherry 

(1979).

Total and effective porosities for the alluvium 

are based on values presented by Magnuson 

(1995). The moisture characteristic curve, adapted 

from a publication by Blumb, Murphy, and Everett 

(1992), represents coarse-grain sediment similar to 

the alluvium. Low residual moisture contents 

reported in the RI/FS also were used to select the 

representative curve (Rodriguez et al. 1997). The 

curve for the unsaturated hydraulic conductivity as 

a function of moisture content was developed 

using the methodology presented by Van 

Genuchten (1978). 

3.3.2.3.2 Basalts—Hydraulic data for 

the basalts in the unsaturated zone are very limited 

at the site. This is due in part to the lack of field 

testing of hydraulic properties in the unsaturated 

basalt. Conventional test methods such as 

pumping or slug tests are not applicable for 

unsaturated basalts. This lack of data is further 

exacerbated by the difficulty in predicting the 

unsaturated hydraulic parameters, such as moisture 

characteristic curves and unsaturated hydraulic 

conductivity, as a function of moisture contents in 

a fractured medium. 

Because of these problems, it is necessary to 

estimate the hydraulic properties indirectly based 

on a number of limiting assumptions. Hydraulic 

conductivities in the basalts are directly related to 

the size and distribution of fractures with the 

individual flow units. Characterizing these basalt 

fractures using conventional drilling techniques is 

difficult at best. Consequently, recent research has 

focused on the characteristics of the basalt flows 

as a general guide to predicting the magnitude of 

the hydraulic conductivity. Comparing the range 

and frequency distribution of hydraulic 

conductivity, Anderson, Kuntz, and Davis (1999) 

found that three main rock types at the INEEL 

affect the magnitude of the hydraulic conductivity. 

These rock types are (1) thin, tube-fed pahoehoe 
3-43

flows, (2) thick, tube-fed pahoehoe flows, and 

(3) near-vent volcanic deposits consisting of shelly 

pahoehoe and slab pahoehoe flows and bedded 

scoria, splatter, and ash. The hydraulic 

conductivity of thin, tube-fed pahoehoe flows 

ranges from 2.4E−02 to 7E+03 m/d (7.9E−02 to 

2.3E+04 ft/d) compared with 3E−03 to 

1.5E+03 m/d (1E−02 to 4.9E+03 ft/d) for the 

thick, tube-fed pahoehoe flows. The higher 

hydraulic conductivity range for the thin flows is 

due to voids associated with the contacts, rubble 

zones, and cooling fractures of thin, tube-fed flow 

that are more predominant in the thinner flows. 

The thick, greater-than-10 m (30 ft), tube-fed 

flows are characterized by fewer contacts, fewer 

rubble zones, and less extensive cooling fractures. 

Where they have ponded in topographic 

depressions, these units are so massive that their 

hydraulic conductivity values may be extremely 

low. Near-vent volcanic deposits represent only 

about 10 to 20% of most lava fields, and thick, 

ponded, tube-fed flows are rare. Hydraulic 

conductivity values associated with near-vent 

deposits are similar to those of the more permeable 

thin, tube-fed pahoehoe flows. 

Anderson (1991) provides a detailed 

discussion of the characteristics of the individual 

basalt units underlying INTEC. These 

characteristics include the number of flows that 

compose a basalt unit, the thickness of the basalt 

units, and the probable source vent for the 

individual flows. Using this lithological 

information as a guide, the basalt units were 

divided into broad categories: basalt units 

comprising thin flows and units comprising thick 

flows. Basalt units comprising thick flows are 

initially assigned a hydraulic conductivity value of 

10 m/d (30 ft/d). These values were modified for 

the perched groundwater simulations, as discussed 

in subsequent sections. This thicker-flow basalt 

includes the flow groups BC, DE3, DE3-4, DE4, 

DE5, and I. The remaining basalt units comprising 

thin flows are assigned a hydraulic conductivity 

value of 1,000 m/d (3,000 ft/d). Both hydraulic 

conductivity values assigned to the basalt units are 

not the highest values but represent the upper 

range of values reported. 

Although there is a large range of hydraulic 

conductivity values reported for the individual 

units, the RI/FS for INTEC (Rodriguez et al. 1997) 

reports a range in hydraulic conductivity for the 

BC group (a thick basalt flow) of 0.4 to 2.1 m/d (1 

to 6.9 ft/d). Conversely, the CPP-02 and 

CPP-DISP well tests indicate hydraulic 

conductivity values for 1,000 to 823 m/d (3,000 to 

2,700 ft/d) for DE3, believed to consist of thinflow units. The noted exception to this generalized 

approach was for thick-flow Basalt Group I that 

had several hydraulic tests conducted and yielded 

a higher hydraulic conductivity value of 170 m/d 

(560 ft/d). 

3.3.2.3.3 Sedimentary Interbeds—

Sedimentary interbeds exhibit a wide range of 

lithologic and hydraulic characteristics, both 

within the same interbed unit and between 

individual units. Hydraulic conductivity values for 

individual units can range as high as four or five 

orders of magnitude. Further, there are often 

inconsistencies between reported lithologies and 

hydraulic conductivities for the respective units. In 

addition, there is a lack of detailed information for 

several of the interbed units—especially the 

deeper units. Consequently, several basic 

assumptions were implemented in selecting the 

hydraulic properties based on drilling logs, 

hydraulic tests, and moisture contents.

The first step in determining the appropriate 

hydraulic parameters for an individual unit is a 

review of the lithologic descriptions from the 

drilling logs. These descriptions can range from 

clay to gravel for the same interbed unit. 

Therefore, several logs were evaluated and the 

dominant lithology type was identified. 

Lithologies were then compared with available 

hydraulic testing data from laboratory core testing 

and field hydraulic testing. Commonly, the field 

tests incorporated both the sedimentary interbed 

and the surrounding basalt, resulting in an average 

value for the two geologic units. Next, available 

moisture content data were compared with the 

lithologic and hydraulic data to select the most 

representative unsaturated hydraulic parameters. It 

was assumed that the limited number of available 

in-situ moisture measurements were representative 

of the residual moisture content. (Higher moisture 

contents are often associated with finer-grained 

sediments as a result of higher capillary forces.) 

Finally, based on the best determination of the 
3-44

lithologic properties and residual moisture 

contents, moisture characteristic curves were 

selected from the available literature that 

represented the type of material comprising the 

individual interbed. 

Based on this approach, the C-CD and D-DE2 

interbeds were assigned relatively low vertical 

permeability values of 0.005 and 0.0025 m/d (0.02 

and 0.0082 ft/d), respectively. Assigning low 

permeable values to these units is important 

because the permeability will control the 

movement of infiltrating water in the subsurface, 

restricting the downward movement of water, and 

forming a perched water zone. The lower 

permeability values for these upper sedimentary 

interbeds are supported by existing perched water 

at these depths in both the northern and southern 

portion of the INTEC area. 

The other sedimentary units assigned a low 

permeability value were interbeds DE4 and DE4-

DE5. These units are discontinuous across the site 

but together can restrict the downward movement 

of infiltrating water across the entire site. Both 

units are thin and located within massive basalt 

flows. By lowering their permeabilities, the effects 

of restricting infiltration and forming perched 

water bodies is consistent with the factors 

resulting in perched water as outlined by Cecil et 

al. (1991). As discussed earlier, whether the low 

vertical permeability is a result of a sedimentary 

interbed or alterations in the baked zone between 

two flows that can contribute to reduced vertical 

hydraulic conductivity, the net result is the same. 

The remaining interbeds were assigned 

hydraulic characteristics for coarser-grained 

sediments such as sands and gravels. Based on the 

modeling simulations and the resulting formation 

of perched water, these hydraulic values may be 

modified to reflect actual conditions. 

3.3.2.4 Model Development. An incremental 

modeling process was used to assess the impacts 

of key input parameters on the resulting numerical 

simulations. This process involved a base-case 

simulation in which the only source input is the 

infiltration from precipitation. Once a mass 

balance was achieved, a second simulation was 

conducted, inputting the losses caused by 

infiltration from the Big Lost Creek. Results were 

evaluated to check the mass balance and the 

validity of the initial assumptions outlined in the 

previous sections. Upon completion of the second 

simulation, the base model was assumed to be 

complete and represented hydrologic conditions 

before human impact at the site. Subsequent 

modeling used this flow base case as a starting 

point for the simulation of human impacts, such as 

seepage from percolation ponds and the simulation 

of contaminant migration in the subsurface. Based 

on the modeling results and assumptions, 

hydraulic parameters, such as hydraulic 

conductivity, were modified in the subsequent 

modeling simulations. These changes are noted in 

the following discussions.

3.3.2.4.1 Flow Base-Case Model 

Infiltration Simulations—The first modeling 

simulation involved the implementation of a twodimensional model as shown in Figure 3-16. A 

detailed cross section based on the work of 

Anderson (1991) was discretized to include the 

major basalt flow units, sedimentary interbeds, and 

surface alluvium. Using the existing technical 

literature and the assumptions outlined in the 

previous section, the best available hydraulic 

parameters were incorporated in the model. This 

approach allowed an evaluation of the existing 

technical data to test the relevance of the 

assumption that no perched water was present in 

the subsurface due solely to precipitation.

Results the base-case model simulation show 

that surface infiltration dominates flow in the 

unsaturated zones. Downward movement of water 

at a flux rate comparable to the infiltration rate 

was observed in the unsaturated zone. Movement 

was dominantly vertical in the unsaturated zone at 

a rate and direction consistent with infiltration 

conditions. Saturated zone simulations will be 

modified in later simulations. 

Moisture contents in the unsaturated portion of 

the problem domain yielded values that were 

consistent with the moisture characteristic curves 

for the individual units. All of the moisture content 

values input to the PORFLOW code represented 

the residual moisture contents (Θr) for the 

moisture characteristic curves of the individual 

hydrologic units. These results are consistent with 
3-45

Figure 3-16. Model grid superimposed on geologic cross section for groundwater model. 

the initial unsaturated hydraulic parameters input 

into the model indicating that the unsaturated 

hydraulic conductivity at Θr is sufficient to 

transport precipitation through the unsaturated 

zone without developing a significant increase in 

moisture conditions. More importantly, no perched 

water zones developed, which is consistent with 

the initial assumptions. 

Water balances for the initial simulation were 

consistent with the amount of precipitation 

entering the upper boundary of the model, plus the 

upgradient groundwater inflow equaling the 

downgradient groundwater discharge. 

3.3.2.4.2 Base-Case Model 

Simulations of Big Lost River Losses—The 

next modeling simulation involved inputting water 

into the model to simulate seepage losses from the 

Big Lost River. Seepage losses from the Big Lost 

River occur sporadically over time as a result of 

periodic flows in the river channel. These flows 

are dependent on precipitation, releases from 

upstream dams, and irrigation diversions. For 

modeling purposes, it was assumed to be a 

constant source, based on seepage records for the 

site. This approach is consistent with earlier 

modeling simulations presented in the RI/FS 

report (Rodriguez et al. 1997).

The amount of water that seeps into the 

ground from the Big Lost River near INTEC was 

based on estimates provided by Bennett (1990). 

Using stream flow data from gaging stations 

located at Lincoln Blvd. (Site 9) and the playas 

site (Site 14) for the period of 1965 through 1987, 

Bennett reports average seepage losses of 82,000 

to 210,000 m3

/d (2,000 to 5,100 acre-ft/mo). 

Dividing the seepage losses by the distance 

between the two gaging stations, 30.1 km 

(18.7 mi) yields an effective seepage rate of 2.72 

and 6.98 m3

/d. For conservative purposes, the 

higher seepage value was used in the simulation. 

Results of the simulation, using the additional 

water input from the Big Lost River seepage, show 

the development of several perched zones directly 

below the river and extending southward into the 
3-46

INTEC area, as illustrated in Figure 3-17. The 

upper perched zone is formed by the combination 

of C-D, D-DE2, and DE2-DE3 interbeds acting as 

the restricting layer. This upper perched zone is 

consistent with the shallow perched water found at 

several locations across the INTEC site. The 

extent of the alluvial perched zone is controlled by 

the permeability of the restricting layer. It is 

reasonable to decrease the vertical permeability of 

these interbeds within hydraulic conductivity 

values consistent with the lithology to increase the 

southern extent of the alluvial perched zone. This 

issue will be addressed during subsequent 

simulations when water from man-made sources is 

introduced into the system. However, the model 

predictions suggest that the source of water in the 

northern INTEC facility could be solely from 

seepage losses along the Big Lost River. 

There are three lower perched layers that form 

as the result of interbeds DE3-DE4, DE4-DE5, 

and DE6. Unfortunately, due to the lack of 

monitoring wells, there is little available 

information to confirm these perched zones. One 

exception is Well USGS-50, which is an open 

borehole from 109 to 123 m (358 to 404 ft) bgs. 

Based on the well completion interval, it is 

possible that the lower perched zones emanating 

from the Big Lost River seepage could be 

responsible for the groundwater observed at Well 

USGS-50. Once again, it is reasonable to decrease 

the vertical permeability of the restricting 

interbeds and increase the southern extent of these 

lower perched zones to intersect Well USGS-50. 

This modeling supports the case that the Big 

Lost River is at least a partial source of 

groundwater for the northern INTEC area. 

Furthermore, results suggest that perched water 

existed, at least on a limited aerial extent, in the 

INTEC area before human influences. The issue of 

modifying subsurface permeability units to extend 

the perched zones will be determined in 

subsequent modeling simulations after evaluating 

the results of inputting man-made sources of 

water. 

3.3.2.4.3 Percolation Ponds—The 

next step in the modeling process is to evaluate the 

impact of the percolation ponds on the underlying 

perched zones. These ponds have been in service 

since 1984, receiving service water from INTEC 

operations. According to the RI/FS, approximately 

96% of all water recharging the perched water 

bodies from man-made sources is a result of

seepage from the percolation ponds (Rodriguez et 

al. 1997). The percolation ponds are unlined ponds 

that were excavated into the surface alluvium. The 

two ponds have been used alternatively, receiving 

an approximate average volume of 5.85E+03 m3/d 

(1.54E+06 gal/d).

There appears to be a discrepancy in the 

amount of water recharged to the perched zones 

Figure 3-17. Model prediction of the hydrologic conditions illustrating the perched water zones resulting 

from the Big Lost River seepage. 
3-47

from the percolation ponds according to the RI/FS. 

This document reports a recharge value because of 

seepage from the percolation ponds of 

7.16E+03 m3

/d (1.89E+06 gal/d), compared with 

the estimated average volume discharged to the 

ponds of 5.85E+03 m3/d (1.55E+06 gal/d). The 

higher seepage number cannot be confirmed; 

however, the amount of water discharged to the 

percolation ponds reported in the RI/FS is 

confirmed in a separate report by Cecil et al. 

(1991). 

The seepage volume from the percolation 

ponds does not account for any evaporative losses 

from the ponds. The RI/FS reports an annual pan 

evaporation rate of 109 cm/yr (42.9 in./yr) for the 

area. This pan evaporation rate is supported by 

regional data presented by Dunne and Leopold 

(1978). Using this pan evaporation value and 

assuming that only one pond is used at a time, the 

resulting losses due to evaporation are 5.75 m3

/d 

(1,520 gal/d). This evaporation loss amounts to 

less than 1% of the seepage volume and, therefore, 

is not considered significant. 

The model consists of a 1-m wide vertical 

cross section in a north-south direction through the 

west pond. For input into the model, the total 

amount of discharge to the ponds was divided by 

two to incorporate the discharge to the individual 

ponds and then multiplied by the ratio of the area 

of the pond area intersected by the model by the 

total area of the west pond. As a result, only water 

lying directly above the model in the region of the 

pond was input into the model. Water was injected 

into the boundary nodes along the top of the model 

in the area of the pond at a total rate equal to the 

percolation rate calculated above. 

To determine the size and location of the 

perched zones, the subsurface lithology is 

compared to the existing well information used to 

select where the permeability should decrease in 

selected subsurface units to form a restricting 

layer. One well that provides a detailed picture of 

subsurface perched zones is Well USGS-51. 

Detailed neutron logging of this well, as presented 

by Cecil et al. (1991), shows three perched zones 

located beneath the percolation ponds. The base of 

the upper perched zone is located at a depth of 

32 m (10E+02 ft). This zone coincides with the CD and C-CD interbeds. The base of the middle 

perched zone is located at a depth of 54 m (180 ft) 

and coincides with the DE2-DE3 interbed. The 

base of the lower perched zone is located at a 

depth of 102 m (335 ft) and roughly approximates 

the DE6 and DE7-DE8 interbeds. Consequently, 

the vertical permeability of these interbeds was 

decreased to the lower end of the reported range 

for the simulations conducted in this model 

simulation. 

Modeling results presented in Figure 3-18 

show the development of several additional 

perched zones, plus the enhancement of existing 

perched zones predicted by the previous Big Lost 

River simulation. A large upper perched zone 

directly below the percolation ponds is evident 

because of the C-D and D-DE2 interbeds. Other 

perched zones have developed on the DE2-DE3 

and DE4 interbeds. In addition, the existing 

perched zones developed by seepage from the Big 

Lost River show an increase in thickness and 

extent because of the additional water provided by 

the percolation ponds. The lower perched zones 

exhibit a larger degree of variability because of the 

lithology. These perched zones are formed by 

interbeds DE6 and DE7-DE8 that are 

stratigraphically located at different depths. This 

difference in depths results in a confining layer 

that ranges from 106.8 to 126.7 m (350.4 to 

415.7 ft) bgs, yielding a lower perched zone that 

changes with depth across the site. The lower 

perched zone is discontinuous but extends across 

the majority of the site. 

3.3.2.5 Model Calibration. There are only a 

limited number of wells completed in the 

unsaturated zone at the site. Compounding the lack 

of wells is the numerous perched zones possible at 

the site. Consequently, a combined approach was 

used for calibrating the model. This approach 

initially relied on an accurate and detailed 

definition of the hydrostratigraphy, incorporation 

of the best available hydraulic information, and the 

input of accurate sources of water to the model as 

previously discussed. Finally, two simulations, the 

first with the Big Lost River seepage and then a 

combined simulation with both the Big Lost River 

and the southern percolation ponds, were 

compared with the existing well information for 

calibration purposes.
3-48

Figure 3-18. Model prediction of the hydrologic conditions illustrating the perched water zones resulting 

from the Big Lost River and percolation ponds seepage. 

The remaining portion of this section 

discusses the assumption, available data, and 

calibration of the unsaturated portion of the model. 

Figure 3-17 shows the moisture distribution 

resulting from Big Lost River seepage. There are 

several perched zones that develop as a result of 

water from the Big Lost River. An upper perched 

zone that forms as a result of the lower 

permeability values assigned to C-D, D-DE2, and 

the DE2-DE3 interbeds extends approximately 

250 m (820 ft) southward across the INTEC 

facility. The extent of perched water in the upper 

unsaturated zone is in general agreement with the 

Hydrologic and Geochemical Assessment (INEL 

1983), which states that seepage from the Big Lost 

River may extend laterally outward from 30 to 

500 m (100 to 2,000 ft) from the river. In addition 

to the upper perched zone, there is a discontinuous 

lower perched zone that forms as a result of the 

lower permeability values assigned to the DE3(4)-

DE4, DE4, DE4-DE5, and DE6 interbeds. 

Although these interbeds were assigned similar 

permeability values, their distribution, extent, and 

thickness (as defined by the geologic cross 

section) results in varied distributions and 

thicknesses of the perched water zones. For 

several of these lower perched zones, there are no 

monitoring wells to confirm or deny their 

existence. 

Figure 3-18 illustrates the distribution of 

perched zones resulting from both seepage from 

the Big Lost River and the percolation ponds at the 

south end of the INTEC facility. The number, 

extent, and thickness of perched zones increase 

significantly because of additional water seepage 

from the percolation ponds. This is expected 

because of the large volume of water associated 

with the percolation ponds compared with the Big 

Lost River. A large perched zone is located 

directly beneath the percolation pond because of 

the low permeability of the interbed units C-D and 

D-DE2. In addition, a smaller perched zone 

developed below this large perched zone because 

of the lower permeability associated with the DE2-

DE3 interbed. The increased extent of the upper 

perched zones resulting from Big Lost River 

Tank Farm Facility 
3-49

seepage reflects the increase in soil moisture 

because of the percolation pond seepage. Finally, 

the lower perched zones initially established by 

Big Lost River seepage also increase in thickness 

and extent because of the additional percolation 

pond seepage. 

To calibrate the development of perched 

groundwater zones caused by seepage from the 

Big Lost River and the percolation ponds, 

Table 3-7 lists the available monitoring wells, 

including surface elevations, well depths, screen 

intervals, and water level elevations. 

The bulk of the monitoring wells are located 

in the upper portion of the unsaturated zone, with 

only four wells available to define the lower 

perched zones. In addition to the well data, the 

lithologic and neutron log of Well USGS-51 from 

Cecil et al. (1991) is presented in Figure 2-23. 

This well log shows the presence of several 

perched zones that can be used to calibrate the 

perched zone below the percolation ponds. 

The following subsections compare the 

available well information and geophysical data to 

calibrate the upper and lower perched zones 

predicted by the numerical simulations. All well 

locations discussed in the following sections are 

shown in Figure 3-19. Monitoring wells presented 

in Table 3-7 are designated in general as 

ICPP-MON-P--#. These wells also are referenced 

to as MW wells at the INEEL. The nomenclature 

in the following discussions are designated as 

MW, which can be directly referenced to the 

ICPP-MON-P well number designations in 

Table 3-7 and Figure 3-19. 

3.3.2.5.1 Upper Perched Zones—For 

the purpose of this report, the upper perched zones 

are defined as saturated water zones above, and 

including, the DE2-DE3 interbed. Figure 3-18 

illustrates a vertical cross section of the site 

showing relative moisture contents resulting from 

the Big Lost River and percolation ponds seepage. 

The model predicts two upper perched zones. The 

northern perched zone is a thin, perched water 

body that results from the Big Lost River seepage. 

This perched water body extends approximately 

700 m (2,000 ft) from the edge of the Big Lost 

River. Compared with the modeling result for 

seepage solely from the Big Lost River as shown 

in Figure 3-17, this perched zone extends another 

450 m (1,500 ft) because of the additional water in 

the unsaturated zone provided by seepage from the 

percolation ponds. The other upper perched zone 

is a thick saturated body that results from 

percolation pond seepage. As noted in previous 

sections, there is a substantial quantity of water 

available for seepage from the percolation ponds 

based on estimates presented in the RI/FS report 

(Rodriguez et al. 1997).

Figure 3-19 compares the extent of the upper 

perched zones with the existing monitoring well 

network. Note that this map differs from the 

previous definition of the upper perched water 

presented in Figure 2-21. For purposes of model 

calibration, it was determined that the vertical 

hydraulic conductivity of the D-DE2 interbed 

provided the controlling influence on the extent 

and thickness of the upper perched zone. Available 

lithologic and hydraulic data indicate that this 

interbed exhibits the lowest vertical hydraulic 

conductivity values for the upper perched zone. 

Interbeds C-D and DE2-DE3 also are important 

for restricting infiltration and forming perched 

water zones. However, based on the water level 

information presented in Table 3-7, the base of the 

upper perched zone coincides with the depth and 

location of the D-DE2 interbed. 

Changes of 0.001 m/d (0.04 in./d) in the 

vertical hydraulic conductivity of the D-DE2 

interbed results in changes of hundreds of meters 

in the lateral extent of the northern upper perched 

zone. Consequently, for calibration purposes, the 

vertical hydraulic conductivity of this interbed was 

adjusted until the lateral extent of the northern 

perched zone matched the extent predicted by the 

monitoring wells. In addition, water levels for both 

the upper perched zones also were compared with 

the predicted head values. For the extent of the 

northern perched zone, it was assumed that the 

southern edge is located between Wells MW-02, 

-04, -05, and -18 (wells showing perched water) 

and Wells MW-08 and-11 (wells reported to be 

dry). 
3-50

Table 3-7. Perched water wells at INTEC. 

Well No. 

Other Well 

Designation 

Surface 

Elevation (m) 

Well 

Depth 

(m bgs) 

Screen Interval 

(m amsl) 

Average 

Water Level 

Elevation 

(m) 

Range in Water 

Level Elevation 

(m) 

CPP-33-1 (5 cm) 1498.15 30.2 1468.0–1471.0 Dry–1468.4 

CPP-33-2 (5 cm) 1497.85 32.3 1465.6–1471.7 1468.2 1467.3–1468.0 

CPP-33-3 (5 cm) 1498.00 37.2 1460.8–1463.9 1462.7 

CPP-33-4 (5 cm) CPP-33-4E 1497.14 36.0 1461.1–1467.2 1466.7 

CPP-37-4 1465.8 

CPP-55-06 (5 cm) 1497.33 34.4 1462.8–1468.9 1465.0 1465.2 

MW-1 (2.5 cm) ICPP-MON-P-001 1498.55 120.4 1386.1–1389.1 Dry 

MW-1 (10 cm) 1498.55 120.4 1396.1–1399.2 1402.1 1400.6–1402.42 

MW-2 ICPP-MON-P-002 1497.5 38.7 1463.4–1466.4 1464.9 1465.9 

MW-3 (2.5 cm) ICPP-MON-P-003 1498.4 46.1 1462.4–1462.9 Dry 

MW-3 (5 cm) 1498.4 46.1 1456.3–1459.4 Dry–1456.5 

MW-4 (2.5 cm) ICPP-MON-P-004 1497.0 39.9 1457.5–1458.0 1457.9 

MW-4 (5 cm) 1497.0 39.9 1463.3–1466.3 1464.9 1465.2 

MW-5 ICPP-MON-P-005 1498.5 43.0 1460.0–1466.1 1464.9 1463.6 

MW-6 ICPP-MON-P-006 1498.5 49.1 1456.8–1462.9 1461.8 1461.9 

MW-7 (2.5 cm) ICPP-MON-P-007 1498.8 53.9 1467.1–1467.6 Dry 

MW-7 (5 cm) 1498.8 53.9 1455.5–1458.6 1456.6 1442.4(?)–1456.9 

MW-8 ICPP-MON-P-008 1497.1 43.0 1459.0–1462.0 Dry–1459.4 

MW-9 (2.5 cm) ICPP-MON-P-009 1499.5 48.2 1467.3–1467.7 Dry 

MW-9 (5 cm) 1499.5 48.2 1459.9–1462.9 1460.3 Dry 

MW-10 (2.5 cm) ICPP-MON-P-010 1498.0 55.2 1474.2–1474.7 Dry–1474.6 

MW-10 (5 cm) 1498.0 55.2 1452.0–1455.0 1453.9 1453.9 

MW-11 (2.5 cm) ICPP-MON-P-011 1498.1 45.9 1463.5–1463.9 Dry 

MW-11 (5 cm) 1498.1 45.9 1456.6–1458.1 1457.9 Dry 

MW-12 (2.5 cm) ICPP-MON-P-012 1497.5 46.6 1451.7–1452.2 Blocked 

MW-12 (5 cm) 1497.5 46.6 1461.2–1464.3 Dry 

MW-13 ICPP-MON-P-013 1499.6 39.0 1467.6–1469.1 Dry 

MW-14 ICPP-MON-P-014 1499.3 42.1 1467.6–1470.7 Dry–1467.3(?) 

MW-15 ICPP-MON-P-015 1499.2 43.6 1459.3–1465.4 1461.9 1462.9 

MW-16 ICPP-MON-P-016 1499.3 38.4 1466.7–1469.8 Dry 

MW-17 (2.5 cm) ICPP-MON-P-017 1499.1 116.1 1415.6–1418.7 Dry 

MW-17 (5 cm) 1499.1 116.1 1440.7–1443.7 1441.0 1441.89 

MW-17 (10 cm) 1499.1 116.1 1383.0–1389.4 1388.1 1388.91 

MW-18 (2.5 cm) ICPP-MON-P-018 1497.8 150.6 1371.6–1377.7 1373.7 Dry 

MW-18 (5 cm) 1497.8 150.6 1460.1–1463.2 Dry 

MW-18 (10 cm) 1497.8 150.6 1351.2–1358.0 1356.7 1357.3–1357.7 

MW-19 ICPP-MON-P-019 

MW-20 (2.5 cm) ICPP-MON-P-020 1498.0 46.2 1465.8–1468.8 Dry–1465.9 

MW-20 (5 cm) 1498.0 46.2 1452.8–1457.5 1457.3 1457.3 

PW-1 1498.85 36.3 1462.6–1468.4 1474.9 1476.4–1478.3 

PW-2 1498.70 39.9 1458.8–1464.9 1461.8 1461.5–1461.2 

PW-3 1498.55 37.5 1461.1–1467.1 1462.4 1462.6 

PW-4 1498.03 45.7 1452.3–1464.5 1475.2 1472.1–1472.8 

PW-5 1500.80 39.3 1461.5–1467.6 1477.7 1476.4–1478.0 

PW-6 1499.74 38.1 1461.6–1467.7 1462.4 1462.8–1463.6 

USGS-50 1498.7 123.4 1375.3–1389.9 1378.0 Dry–1378.0 

Sources: INEL 1993; Rodriguez et al. 1997; DOE-ID 1994. 
3-51

Figure 3-19. The extent of upper perched water at the INTEC facility based on perched water well data in 

Table 3-7. 
3-52

As shown in Figure 3-19, the extent of the 

perched zone resulting from the Big Lost River 

seepage is approximately 670 m (2,200 ft). The 

southern extent of the Big Lost River perched unit 

is located between Wells MW-02, -04, -05, and 

-18 and Wells MW-08 and -11—consistent with 

the calibration requirements discussed in the 

previous paragraph. Since the volume of water 

required to form a perched water zone from other 

possible sources has not been clearly defined in 

previous investigations, it appears more likely that 

the assumption of the Big Lost River as the source 

of the upper northern perched zones is valid. For 

example, the RI/FS for the site reports that the 

seepage from the sewage lagoon is only 2.1% of 

the estimated volume of water recharging the 

perched water bodies at the INTEC site. This 

compares with 95.8% for the percolation ponds. 

Water system leaks that could possibly provide 

water to the previously defined perched zone in 

Well MW-18 are only 0.6%. In comparison, 

reported infiltration from precipitation accounted 

for only 1.1% of the estimated volume of water 

recharging the perched zones and previous model 

simulations showed that no perched water zones 

developed at the site due solely to precipitation. 

The elevation of the upper perched water 

emanating from the Big Lost River is predicted by 

the numerical simulation ranges from 

approximately 1,460 m (4,790 ft) amsl to 

approximately 1,450 m (4,760 ft) amsl in the 

southern portion of this perched unit. Table 3-7 

shows water-level elevations in the northern 

portion of this perched unit are as high as 1,468 m 

(4,817 ft) amsl and as low 1,453 m (4,767 ft) amsl 

in the southern portion of the perched unit. 

Because the monitoring wells used for calibration 

purposes are screened at various depths and the 

model predicts generally lower water-level 

elevations as the perched unit moves southward, 

there is a reasonable agreement between the model 

prediction and the actual site data for the upper 

perched zone emanating from the Big Lost River. 

The upper perched zone resulting from the 

percolation ponds’ seepage also is shown in 

Figure 3-19. The model predicts that this upper 

perched water body is the largest perched zone at 

the site—consistent with the fact that the 

percolation ponds are the largest source of water 

seepage in the area. The model further predicts 

that water-level elevations range from 1,475 m 

(4,839 ft) amsl directly below the ponds, to 

1,444 m (4,738 ft) at the base of the perched zone. 

The southern extent of this perched unit is not 

shown in the figure but is located 420 m (1,400 ft) 

south of the southern edge of the pond. For 

calibration purposes, the first set of wells, 

compared with the model prediction, is the 

perched water (PW)-series wells, PW-1 through 

PW-6, (Figure 3-19). These wells show waterlevel elevations that range from 1,462 to 1,478 m 

(4,797 to 4,849 ft) amsl. Most of the PW-series 

wells are located adjacent to the ponds, where the 

steep decline in water levels is evident in the 

model simulation. The observed water levels in 

these adjacent PW-series wells are comparable to 

the water levels based on the model prediction. 

The variation of the water-level data from the PWseries wells is due in part to different screen 

intervals measuring different hydrologic units. 

The second set of wells used to evaluate the 

depth and extent of the upper perched zone 

associated with the percolation ponds includes 

MW-07, -13, -15, -16, and -17. The model predicts 

that all five wells are located within the extent of 

the upper perched zone. However, Wells MW-13 

and MW-16 are dry, while the other three wells 

contain water. The explanation appears to be the 

location of the screen interval for these wells. The 

dry wells, MW-13 and MW-16, are screened 

above the perched water with the base of the well 

screens located at 1,468 and 1,467 m (4,817 and 

4,813 ft) amsl, respectively. The saturated wells 

are screened 8 to 28 m (30 to 92 ft) lower and 

intersect the perched water zone, predicted by the 

model. This evidence is strong confirmation on the 

northern extent of the upper perched zone 

associated with the percolation ponds. 

3.3.2.5.2 Lower Perched Zones—For 

the purpose of this report, the lower perched zones 

are defined as those located below the DE2-DE3 

interbed. Figure 3-18 illustrates the location of 

perched groundwater zones based on the results of 

the unsaturated flow model. The lower perched 

zones are discontinuous and controlled by the 

discontinuous low-permeable interbed units. 

Water sources for these lower perched zones 
3-53

include seepage losses from the Big Lost River 

and the percolation ponds.

There are only a few wells completed in the 

deeper perched zones at the site that can be used to 

evaluate the accuracy of the model predictions. 

Monitoring Wells MW-1 (10-cm inside diameter), 

MW-17 (10-cm inside diameter), MW-18 (2.5-cm 

inside diameter), and USGS-50 are completed in 

the lower perched aquifer and showed water-level 

elevations that ranged from 1,356.7 to 1,441.0 m 

(4,451.3 to 4,727.9 ft) in 1995. This compares 

with the model that predicts a range of water-level 

elevation for the lower perched zone of 

approximately 1,367 to 1,436 m (4,485 to 

4,712 ft). This is a favorable comparison between 

actual and predicted water levels considering the 

large degree of variability in both the location of 

the deep perched zone and the reported water-level 

data. However, there are several lower perched 

zones that have no wells screened at their location 

or depth. Consequently, it is not possible to 

calibrate the existence of these lower perched 

zones. The existence of several lower perched 

zones relies on the accuracy of the geologic cross 

section used as a basis of the model and the 

relative permeability values based on limited 

lithologic data. 

Comparing individual water-level data for the 

limited number of monitoring wells with the 

predicted water-level elevations based on model 

predictions yields a reasonable agreement. 

Monitoring Well MW-01 (10 cm) has a waterlevel elevation of approximately 1,402 m 

(4,600. ft) amsl. The model results show a perched 

zone resulting from the low-permeable DE4-DE5 

interbed extending southward from the Big Lost 

River. The water-level elevation predicted by the 

model is approximately 1,400 m (4,600 ft)—only 

a 2-m (7-ft) difference than the water level in 

MW-01. Well MW-17 (10 cm) has a water-level 

elevation of 1,388 m (4,554 ft) amsl, compared 

with the model that shows a water-level elevation 

of 1,381 m (4,531 ft) for the perched zone resting 

upon the DE6 interbed. Well MW-18 (2.5 and 10 

cm) has water-level elevations of 1,374 and 

1,357 m (4,508 and 4,452 ft) amsl, respectively. 

The model predicts several small scattered perched 

zones at the location that approximate the reported 

data. Since the MW-18 wells have previously 

reported dry conditions, the scattered thin perched 

zones are consistent with the limited amount of 

perched water in this area. 

Well USGS-50 has a water-level elevation of 

1,378 m (4,521 ft) amsl, compared with 1,380 m 

(4,530 ft) for the perched zone controlled by the 

DE-6 interbed—a difference of only 2 m (7 ft). 

Well USGS-50 is located at the southern edge of 

this DE-6 perched zone, perhaps explaining why 

this well has occasionally been observed to be dry. 

A second source of important data for the 

calibration of the lower perched zones is the 

detailed geophysics conducted on Well USGS-51 

(Figure 2-23). This well was repeatedly logged in 

1984, 1985, and 1986, showing the development 

of several perched zones with depth. The lowest 

perched zone associated with the DE6 interbed as 

predicted by the model is not evident in the 

geophysical logs from Well USGS-51. The next 

lowest perched zone, from 83 to 102 m (270 to 

335 ft) bgs, coincides with the DE4 perched zone 

predicted by the model that shows saturated 

conditions from approximately 83 to 101 m (270 

to 331 ft) bgs. Well USGS-51 intersects the DE2-

DE3 interbed perched zone at the thinnest 

saturated portion, and therefore is not evident in 

the geophysical log for the well. The other two 

perched zones identified in the geophysical log 

correspond to the upper perched zone associated 

with the percolation ponds. 

The final source of data for the calibration of 

the lower perched zones is information presented 

in the RI/FS report (Rodriguez et al. 1997), which 

lists perched zones for several USGS wells. This 

information is reproduced in Table 3-8. There is 

concern regarding this information because the 

majority of the USGS wells listed in the table are 

completed in the underlying regional aquifer. It is 

hard to conceive water levels above the regional 

aquifer, considering the substantially higher 

transmissivities exhibited by the regional aquifer. 

There is information provided, however, that the 

reported water levels occurred during the drilling 

of the wells. 

USGS Wells-40, -41, -43, and -44 have water 

levels from 1,382 to 1,395 m (4,534 to 

4,577 ft) amsl. Comparing the perched zones 
3-54

Table 3-8. Perched layer, depth to water, and water table elevations for the wells encountering the lower 

perched zone (Rodriguez et al. 1997). 

Wella

Water level 

(m bgs) Year Interbed 

Interbed Depth

(m) 

Water Table Elevation

(m amsl) 

MW-1 98 1995 Unknown Unknown 1401 

MW-17 111 1995 DE8 >116 1389 

MW-18 124 1995 DE8 125–129 1388 

USGS-40 104 1956 DE6 112–116 1395 

USGS-41 117 1956 DE8 123–131 1382 

USGS-43 110 1957 DE6 112–131 1388 

USGS-44 117 1957 DE8 125 1382 

USGS-50 117 1995 DE8 120–123 1382 

USGS-51 82 1986 DE6 117–122 1428 

USGS-52 117 1960 DE8 126–130 1379 

a. MW wells also are designated as ICCP-MON-P at the INEEL. 

developed solely by seepage from the Big Lost 

River, the DE6 interbed, resulting in a perched 

layer, yields the highest water level of 1,381 m 

(4,531 ft) amsl. This comparison indicates that the 

DE6 perched zones associated with the Big Lost 

River as predicted by the model are real and 

supported by the available data. 

In summary, there is general agreement in the 

location, depth, and water-level elevations 

between the numerically predicted perched zones 

and the perched zones defined by the available 

monitoring well network. There is uncertainty 

regarding the exact locations of all of the perched 

zones because of the limited number of wells 

available for calibration. The biggest difference in 

modeling results presented in this report and 

previous modeling exercises is the number and 

locations of sources providing water for the 

formation of perched zones. The approach 

presented in this report relies on estimates of 

precipitation and seepage from the Big Lost River 

and the percolation ponds with the adjustment of 

interbed hydraulic conductivity to calibrate the 

location and extent of the perched zones. Based on 

this approach, the existing distribution of perched 

zones can be reasonably matched. 

Previous investigations relied on several 

sources of water for the formation of perched 

zones. The INTEC RI/FS used a similar approach 

of including water from several sources at the site. 

However, these additional sources provide only a 

few percent of the total source of water available 

for the formation of perched water and are not 

significant when compared to the total amount of 

water provided by precipitation, the Big Lost 

River, and the percolation ponds. 

To reiterate, the fundamental approach used in 

the numerical simulation of the unsaturated zone 

was to rely on the USGS definition of the site 

hydrostratigraphy, estimates of precipitation, and 

Big Lost River and percolation pond seepage 

values confirmed by several sources from the 

available literature. The hydraulic conductivity of 

the low-permeable interbed was varied within 

reasonable ranges based on lithologic descriptions 

to predict the location and extent of the numerous 

perched water bodies underlying INTEC. 

3.3.2.5.3 Transport Verification—

The disposal of tritium in the percolation ponds 

and the monitoring data from surrounding PWseries wells provide a tracer test to compare the 

transport of tritium predicted by the model with 

actual conditions. According to Cecil et al. (1991), 

the first percolation pond was placed into service 

in February 1984. It is further reported that 960 Ci 

of tritium was discharged to the percolation ponds 
3-55

from 1984 through 1988. Orr and Cecil (1991) 

report a tritium disposal rate of 185 Ci/yr from 

1986 to 1988. In November 1988, the PW-series 

wells surrounding the percolation ponds reported 

tritium values in the perched groundwater ranging 

from 5.5 to 36.7 pCi/mL (Cecil et al. 1991).

Using the Big Lost River seepage model as the 

flow base case, the percolation ponds simulation 

was run for approximately 5 yr to simulate 

wastewater from the percolation ponds entering 

the unsaturated zone. Reported tritium disposal 

volumes were averaged over the 5-yr run period. 

Results, shown in Figure 3-20, illustrate the 

contaminant plume in the unsaturated zone. Based 

on the model predictions, tritium concentrations 

range from 10 to 80 pCi/mL. This is higher than 

that reported by Cecil et al. (1991) (5.5 to 

36.7 pCi/mL), yet comparable to observed results. 

3.3.2.5.4 Final Flow and Transport 

Model—During the radionuclide transport 

simulations, the percolation ponds will no longer 

be in service. Consequently, the water transport 

portion of the model is based on the hydraulic 

values selected during the calibration of the model 

but without the input of the percolation ponds. The 

extent of the perched water is due solely to the Big 

Lost River seepage and precipitation. The extent 

of future perched water is similar to the extent 

shown in Figure 3-17.

It was necessary to calibrate the water 

transport model using the percolation ponds 

because the well data used in the calibration were 

collected during the operation of the ponds. By 

using the calibrated model without the percolation 

ponds, a realistic simulation of future subsurface 

conditions was obtained. The PORFLOW flow 

and transport input files are provided in 

Appendix G. 

3.3.3 Radon Transport Model 

This report considers the flux of radon from a 

grouted slab of residual waste, which has finite 

thickness in one spatial direction and is of infinite 

thickness in the other two spatial directions (i.e., a 

1-dimensional model). The slab is submersed in 

air containing no radon. This model is 

conservative for the TFF LLW site for several 

reasons: 

Figure 3-20. Distribution of tritium resulting from discharge into percolation ponds (units in pCi/mL).
3-56

• No credit for the tank or vault is assumed.b

These materials provide a significant barrier 

for radon transport to the surface of the 

facility. 

• Submersion of the slab in radon-free air 

allows for the greatest possible diffusive flux 

out of the slab. 

This conservative model is chosen to simplify 

the calculations, making them more transparent 

and easier to understand. 

3.3.3.1 The Transport Model. The model 

presented herein is a one-dimensional diffusiononly model for radon transport out of a grouted 

heel slab. Advective flow is not considered in this 

model because cementitious materials are too 

impermeable for pressure-driven flow to be a 

significant means of radon transport in the slab 

(Renken and Rosenberg 1995; Tanner 1990; 

Rogers and Nielson 1992). Also, since the waste 

is being buried at depth, it is difficult to envision a 

scenario where a pressure gradient across the slab 

could be established. Figure 3-21 shows a 

schematic diagram of this model.

The equation describing the diffusive 

transport of radon in this slab is 

( ) () () G Rn Rn Rn Rn Rn D C x,t C x,t S

t

C x,t = ∇ − λ + ∂∂ 2 (3-14) 

where 

CRn ( ) x,t = time and space-dependent 

radon concentration in the 

air-filled pore space in the 

slab (atoms/m3

) 

 

b. This assumption is somewhat modified for the short-lived 

nuclides (220Rn and 219Rn) to show compliance. It is shown 

that a very moderate overburden (i.e., much less than the 

10 ft of overburden currently projected) is sufficient to 

prevent release of these two short-lived isotopes flowing in 

the presence of the more dominant isotope, 222Rn. 

DG = bulk diffusion coefficient for 

radon in the grouted slab (m2

/s) 

λRn = radioactive decay constant for 

radon (s-1) 

SRn = source of radon (atoms/m3·s) 

(footnote c) 

222Radon (half-life of 3.82 d) is much longerlived than either 220Rn (half-life of 55.6 s) or 219Rn 

(half-life of 3.96 s). Hence, the diffusive transport 

in the slab will be dominated by 222Rn because of 

its relative abundance on an atomic fraction basis. 

This fact means that 220Rn and 219Rn can be 

treated as trace gases flowing in the presence of a 

more dominant gas. The shorter-lived isotopes 

would need to be present in activity

concentrations at least three or four orders of 

magnitude larger than 222Rn to invalidate this 

assumption. Inventories given in Section 2.3 do 

not show these trace isotopes to be present in such 

high quantities. Therefore, the rest of this section 

will consider 222Rn. A later section of this chapter 

will demonstrate that the shorter-lived isotopes 

cannot penetrate even a modest barrier. 

For the TFF, the steady-state solution is of 

interest. Hence, the time derivative is neglected 

and Equation (3-14) becomes 

( ) ( ) G Rn Rn Rn Rn = D ∇ C x − λ C x + S 2 0 (3-15) 

which has the general solution of the form 

( )

A exp( ) λ / D x

A exp λ / D x λ

S C ( x )

2 Rn G

1 Rn G

Rn

Rn Rn

+ − ⋅

= + ⋅ (3-16) 

where A1 and A2 are constants that are determined 

by boundary conditions. 

The boundary conditions for this problem are 

 

c. This term can be treated as a constant in this analysis 

because the radium concentration in the slab is varying much 

more slowly (from ingrowth and decay) than the radon 

concentration. 
3-57

Figure 3-21. A diagram of the one-dimensional slab model.

0

0

=

x=

Rn

dx

dC ( x ) (3-17) 

CRn ( x = T ) = 0 (3-18) 

Equation (3-17) is the condition that no flow is 

allowed out of the back surface of the slab, which 

allows for greater diffusive transport out of the 

front face of the slab. Equation (3-18) provides for 

continuity of concentration across the cement/air 

interface. This situation is analogous to the 

grouted heel setting on the ground surface in open 

air. Using these boundary conditions, the solution 

for the radon concentration is 

( ) ( ) ( ) ( ) ( )















λ ⋅ + − λ ⋅

λ ⋅ + − λ ⋅ − λ = exp / D T exp / D T

S exp / D x exp / D x C x

Rn G Rn G

Rn G Rn G

Rn

Rn Rn 1 (3-19) 

and the solution for the radon flux at the surface of 

the slab isd

( ) ( ) ( ) ( ) ( )















⋅ + − ⋅

⋅ − − ⋅ = = D T D TD T D T

J x T pS D D

Rn G Rn G

Rn G Rn G

Rn Rn G Rn G

exp / exp /

exp / exp /

/

λ λ

λ λ

λ

 (3-20) 

By substituting appropriate values for the 

parameters in Equation (3-19), a value for the flux 

of radon from this slab can be evaluated. 

 

d. The derivation of this expression is performed in 

Appendix H. 

3.3.3.2 Appropriate Values for Radon Gas 

Transport Parameters. Table 3-9 lists the 

parameter values and the source of the information 

that was used for this analysis.

Substituting the transport parameter values 

listed in Table 3-9 along with the 226Rn 

concentration (1.92E−04 Ci/tank; 

1.73E+06 pCi/m3

) into Equation (3-20), the flux of 

222Rn out of the surface of the slab is 0.39 pCi/m2

/s 

after converting to pCi from Bq (1pCi=0.037 Bq). 

This value is more than an order of magnitude 

below the standard of 20 pCi/m2

/s given in DOE 

Manual 435.1-1 (DOE 2001a). The reason that 

such a conservative model was still able to meet 

the standard is that there is simply not much 226Ra 

in the slab. Also, cement materials do not emanate 

or transport radon very well. 

The situation for the short-lived isotopes 220Rn 

and 219Rn is not as simple as was seen previously 

for 222Rn. This is due to the fact that these shortlived isotopes are flowing in the presence of the 

more long-lived, dominant isotope and are being 

swept along at higher diffusive velocities than 

would be the case if they were present alone (see 

Bird, Stewart, and Lightfoot 1960). Where the 

treatment of multiple isotopic flow is discussed 

herein, we assume that mass differences in the 

isotopes are negligible for such small distances of 

travel. 

Air

(zero radon concentration)

Grouted Heel

Slab

x=0 x=T Direction of Increasing

Slab Thickness

Radium

Contamination
3-58

Table 3-9. Values of the gas transport parameters for the grouted heel along with the source of the 

information. 

Parameter Value Source Notes 

Bulk diffusion coefficient 

for radon in grout (DG) 

4.6E−07 m2

/s Renken and Rosenberg 

(1995) 

Conservative upper limit 

for concrete 

Emanation coefficient (ε) 0.3 Gadd and Borak (1995) Conservative upper limit 

for concrete 

Slab thickness (T) 2 ft (0.6 m) Tank Farm Closure Study 

(INEEL 1999a) 

Design minimum 

thickness of grouted heel 

Decay constant for 226Ra 

(λRa) 

1.37E−11/s Parrington et al. (1996) 

Decay constant for 222Rn 

(λRn) 

2.09E−06/s Parrington et al. (1996) 

Porosity (p) 0.1 Gadd and Borak (1995) 

Assuming that 222Rn dominates the diffusion 

process in the slab,e

 the steady state radon 

transport equation for the shorter-lived isotopes is 

* * * * G C ( x ) C ( x ) S

C ( x )

C ( x ) = D ∇ 2 − λ + 2

2

0 (3-21) 

where the subscript 2 indicates a value for 222Rn 

and the subscript * indicates a value for the 

shorter-lived isotope in question (i.e., either 220Rn 

or 219Rn). The solution for the flux of these 

isotopes at the surface of the slab is then given 

byf

: 

J ( x T ) S

S

J ( x T ) * * * = λ

λ = = 2

2 2

 (3-22) 

By substituting appropriate parameter values 

(i.e., half-lives [Parrington et al. 1996] and 

appropriate inventories for radium [Staiger and 

Millet 2000]) into Equation (3-22), the radon flux 

out of the surface of the slab can be calculated. 

When this substitution is made, however, the bare 

slab in open-air conservative calculation no longer 

meets the standard (by about two orders of 

magnitude) because the diffusive flux from 222Rn 

 

e. This was shown to be a reasonable assumption in the 

previous section. 

f. The derivation of this expression is given in Appendix H. 

is able to drive these nuclides out of the system 

more rapidly than would otherwise be the case. 

The following sections show, however, that even 

a modest overburden will prevent the migration of 

these two nuclides out of the system before they 

radioactively decay. 

The analytical model used to demonstrate the 

inability of the short-lived isotopes to migrate 

from the system adds a diffusion barrier atop of 

the slab model shown previously. This model is 

shown pictorially in Figure 3-22. 

Here, Ts is the system thickness and is the 

compliance point for the calculation. For 

simplicity and to make use of the relations derived 

for a bare slab, the interface condition in this 

model is that the flux at the interface between the 

slab and the overburden is identical to that 

determined by Equation (3-19). This is a 

conservative assumption because radon flows 

more easily out of a bare slab than out of one 

covered by an overburden. Also, for simplicity, 

the slab and overburden are assumed to have 

equal porosities. Again, the shorter-lived isotopes 

are flowing mainly in the presence of the 

dominant isotope, 222Rn. 

The steady-state transport equation for 222Rn 

across the overburden is given as 

D C ( x ) C ( x ) = S∇ 2−S − λ2 2−S 2 0 (3-23) 
3-59

Figure 3-22. A diagram of the 1-dimensional slab model with overburden. 

where 

DS = bulk diffusion coefficient for radon 

in the overburden (m2

/s) 

C2-S = 

222Rn concentration as a function of 

position in overburden (atoms/m3

) 

λ2 = radioactive decay constant for 222Rn 

(s−1

). 

Here the subscript S denotes properties of the 

overburden (i.e., soil) and the subscript 2 denotes 

properties of 222Rn. With the boundary conditions 

that the concentration of radon goes to zero at the 

outer boundary and the flux of radon into the 

overburden is given by Equation (3-20), the 

following equation gives the flux of 222Rn out of 

the overburden: 

[ ] ( ) ( ) ( ) [ ] ( ) ( ) ( )

[ ] ( ) ( )

2 G s 222

2 s s 2 s s

2 G s

2 s s 2 s 2 s

2 s s 2 s s 2 s s

2 S 2 G s

J ( x T ) RF

exp λ / D (T T ) exp λ / D (T T )

2 J ( x T )

exp 2 λ / D T exp λ / D T exp λ / D T

exp 2 λ / D T exp λ / D T exp λ / D T J ( x T ) J ( x T )

= = ⋅

⋅ − + − ⋅ − = = ⋅

− ⋅ ⋅ + − ⋅

− ⋅ ⋅ + − ⋅ = = =

−

−

− −

 (3-24) 

where 

RF222 = reduction factor for radon due to 

decay while traversing the 

overburden. 

Defining the mean migration time (tm) as the 

average time it takes a radon atom to traverse the 

overburden, it follows that 

222

222 222 2 λ

− = −λ ⇒ = ln( RF ) RF exp( tm ) tm (3-25) 

and 

( )















λ

λ = 222

222

RF exp ln RF * * (3-26) 

Since the radioactive decay constants for 220Rn and 

219Rn are well over three orders of magnitude 

larger than that of 222Rn, and the natural log term 

is always less than zero, the reduction factor for 

the short-lived nuclides is vanishingly small for all 

but the thinnest of overburdens (i.e., where RF222

approaches unity). For reference, a 0.5-m-thick 

overburden of soil with a diffusion coefficient of 

7E−06 m2

/s (the diffusion coefficient for radon in 

air, a conservative upper limit for radon diffusion 

in any medium [Rogers and Nielson 1992]) would 

have a reduction factor of about 1096, with the 

reduction factor for 219Rn being even smaller. 

3.3.4 Volatile Radionuclide Model 

The release of gaseous species of 3

H and 14C 

to the air above the TFF was assumed to be 

controlled by gaseous diffusion in the air-filled 

pore space. Assuming that the concentration 

Air 

(zero radon concentration) 

Grouted Heel

Slab

x=0 x=T Direction of Increasing 

Slab Thickness 

Radium 

Contamination 

x=T s

Overburden
3-60

gradient is constant over time, a conservative 

release rate, Qi (in Ci/yr), can be estimated from a 

modified one-dimensional flux equation of the 

form 

A z

C

Q D ,i

i e,i ×













 = × 0 (3-27) 

where 

De,i = effective diffusion coefficient of 

radionuclide i, in pore spaces 

(m2

/yr) 

C0,i = initial air concentration of 

radionuclide i in the pores of the 

waste zone (Ci/m3

) 

z = average length of diffusion path to 

surface (m) 

A = total area of waste facility (m2

). 

The expression above represents a maximum 

release rate because it assumes that the initial 

inventory is immediately available for release and 

that it is not significantly depleted over a year of 

release. Realistically, depletion from volatilization 

may greatly alter the concentration in the waste 

during a period of 1 yr. 

To estimate C0,i for 3

H and 14C, the chemical 

form of these radionuclides in the waste form must 

be taken into account. Tritium released to the 

pores of the waste form is assumed to be water 

vapor (i.e., tritiated water [HTO]). This 

assumption is conservative because the dose 

conversion factor for inhalation of HTO is many 

orders of magnitude greater than that for HT. 14C 

is assumed to be solely associated with gaseous 

CO2 in the air-filled pores. 

To estimate the air concentration of HTO at 

the source, C0,H-3, the concentration of 3

H in the 

pore water is assumed to equal its concentration in 

water vapor. The amount of water vapor, and thus 

HTO vapor, in the pores is a function of 

temperature and relative humidity. At 100% 

relative humidity, which can be assumed for the 

pores because of the presence of residual water, 

the concentration of water vapor, Cv,H2O in g/m3

, 

was estimated from 

MW

RT

P C v v,H O = × 3

2 10 (3-28) 

where 

103

 = unit conversion factor (L/m3) 

Pv = vapor pressure at given 

temperature (atm) 

R = gas constant (0.082 L-atm/ mol-oK) 

T = temperature (oK) 

MW = molecular weight of water (18 

g/mol). 

Equation (3-27) does not account for 

radioactive decay or the porosity of the 

overburden. The fluxes were determined for each 

radionuclide at the time of closure, which is 

conservative. 

Assuming a subsurface temperature of about 

283o

K (10oC), corresponding to a vapor pressure 

of water of 1.2E−02 atm (9.1 mm Hg), the 

concentration of water vapor in the voids is 

approximately 9.3 g/m3

. 

The waste form was assumed to have a 

porosity and water content equal to the values 

associated with the alluvium (for conservatism) to 

estimate the volume of water in the waste. The 

inventory was assumed to be diluted within this 

volume. Therefore, assuming a porosity of 0.24 a 

waste volume of approximately 27.8 m3

 (982 ft3), 

and a 50-ft tank with a 6-in. heel, there is 

estimated to be about 6.7 m3 (240 ft3

) of water in 

the waste form. 

Assuming all of the 3

H disposed of in the TFF 

is initially associated with this water, an activity 

concentration, Cp,H-3, of 0.10 Ci/m3

 is estimated for 

the pore water. This value is based on the 

estimated, combined 3

H inventory at closure of 

5.6E−01 Ci in the tank (Table 2-17), 2.47E−23 Ci 

in the sand pad (Table 2-18), and the 2.0 m3

(71 ft3

) of water estimated in the pores. If the 
3-61

concentration of 3H in pore water is equivalent to 

its concentration in water vapor in the pores, the 

air-filled pore concentration, C0,H-3, can be 

calculated from 

w

v,H O ,H p,H

C

C C

ρ − = − × 2 0 3 3 (3-29) 

where 

ρw = density of water (i.e., 106

 g/m3). 

The estimated 3

H concentration in the pore gas 

following this procedure is 7.8E−07 Ci/m3

. The 

effective diffusion coefficient, De,H2O , for H2O 

through the vadose zone to the ground surface was 

estimated as 

( ) O v H2 O a, H2 e, D = D × 0.66 × ε − θ (3-30) 

where 

Da,H2O = diffusivity of H2O in air 

0.66 = tortuosity coefficient 

ε = total porosity (0.49) 

θv = volumetric water content (0.073). 

Equation (3-29) is the empirically-derived 

Penman relation (Hillel 1980) describing the 

relationship of effective diffusivity in soils to the 

unhindered diffusivity in air. The diffusion 

coefficient of water vapor in air is approximately 

754 m2

/yr (8,120 ft2/yr) (CRC Press 1981). Thus, 

the effective diffusion coefficient for water vapor 

in soil is approximately 208 m2/yr (2,240 ft2/yr). 

Using a pore gas 3

H concentration of 

7.8E−07 Ci/m3

 in Equation (3-22), an average 

diffusion path length of 3.0 m (i.e., 10 ft of cover 

soil), a tank area of 182 m2

 (1,960 ft2), and a 

diffusion coefficient for water vapor in soil of 

208 m2

/yr (2,240 ft2/yr), the maximum release rate 

of HTO to the atmosphere at closure is estimated 

to be 9.8E−03 Ci/yr. This is an extremely 

conservative value because of (a) the simplified 

computational approach that does not account for 

the first-order dependence of the release rate on 

the pore water concentration, and (b) the 

assumption that HTO diffuses through the vadose 

zone without further exchange of 3

H with bound 

water in the soil. A more complicated realistic 

model was not warranted because the relatively 

short half-life of 3

H alone depletes the source 

sufficiently to prevent excessive doses from 3

H at 

the end of the institutional control. Calculations of 

offsite doses before cessation of institutional 

control are based on this conservative release rate. 

To estimate the pore-water concentration of 

14CO2 at the source (Cp,C-14), a sorption coefficient 

of 5 m3/kg was used to partition the 14C from the 

waste to the pore water. The initial concentration 

of 14C in the pore water at closure is estimated to 

be 1.7E−06 Ci/m3

 based on a total activity at 

closure of 0.49 Ci (Tables 2-17 and 2-18), a 

porosity of 0.24, and a total tank waste volume of 

27.8 m3

 (982 ft3). 

The effective diffusivity of CO2 in soil was 

estimated to be 121 m2

/yr (130E+01 ft2/yr) based 

on diffusivity in the air of 440 m2

/yr (4,700 ft2/yr). 

From Equation (3-27), the maximum flux of 14C is 

estimated to be 0.012 Ci/yr, assuming a diffusion 

path length of 3.0 m (10 ft) of cover soil and a 

waste area of 182 m2 (1,960 ft2

). 

These release rates were adjusted to account 

for the eleven 300,000-gal tanks (i.e., they were 

multiplied by 11). They were then input into the 

CAP-88 PC model (EPA 1992) to determine doses 

to the general public from the atmospheric 

pathway. 

The receptor was located 100 m (300 ft) from 

the TFF. The analysis used 1992 wind data from 

the INEEL Grid-3 meteorological station 

(obtained from the NOAA Office at the INEEL). 

Standard default values for agricultural products, 

based on the rural definition in CAP88-PC, were 

used in the analysis. The outputs from the 

modeling runs are provided in Appendix I. 

3.3.5 Dose Model 

The methodology used to calculate the allpathways dose is based on the methodology 

present in reports by NRC (1977), Peterson 

(1983), and Maheras et al. (1997). This all-
3-62

pathways scenario assumes that a receptor receives 

radiation doses by consuming 

• Contaminated groundwater 

• Contaminated animal products (i.e., milk and 

beef from cattle that consume contaminated 

water and pasture grass irrigated with 

contaminated groundwater) 

• Contaminated leafy vegetables and produce 

(i.e., all vegetation other than leafy 

vegetables). 

Radionuclide concentrations as a function of 

time at the receptor well (calculated using the 

hydrological transport model, described 

previously) were used as input to these dose 

models. The receptor is located at the INEEL 

boundary during the operational and institutional 

control periods. During this time, the INEEL Site 

boundary is maintained and access by the public is 

not allowed. During the post-institutional control 

period, the member of the public is assumed to 

have a well located in the groundwater at the point 

of maximum concentration, which for analyses 

presented in this PA is located 600 m (2,000 ft) 

from the facility. Table 3-10 lists the parameters 

used in the all-pathways dose analysis, their 

distribution types and ranges, and the sources of 

information used to establish these ranges. These 

parameter distributions are used in this analysis, 

which is done stochastically for two reasons: 

• To facilitate sensitivity analyses 

• To examine the uncertainty on the final dose 

estimates. 

The 95% confidence level will be used as the 

metric for comparison to the standards to “provide 

reasonable expectation” that the performance 

objective is met (DOE 2000). 

For reference, Table 3-11 shows the parameter 

values used by Maheras et al. (1997) for a similar 

dose assessment at the INEEL. The values in 

Table 3-11 generally fall within the ranges found 

in Table 3-10. However, since many of the 

parameter distributions in Table 3-10 were from 

work performed for the Yucca Mountain Project (a 

site very similar in climate and geography to the 

INEEL) (LaPlante and Poor 1997), and since 

values from Maheras et al. (1997) were mostly 

national averages, some deviation is expected. The 

new values will be more representative of an arid 

environment and lifestyle. 

3.3.6 The All-Pathways Dose Assessment 

Model 

Figure 3-23 presents a block diagram of the 

all-pathways dose assessment model used herein, 

along with a listing of the input parameters to this 

dose model and where they reside in the 

calculation. The following paragraphs and 

equations describe the mathematical details of this 

model. The model is very similar to the model 

used by Maheras et al. (1997). 

The dose from direct consumption of 

groundwater as drinking water was calculated as 

follows: 

∑

μ = − × × × i DW GW i W i Ci

pCi D C U DCF

6 10 (3-31) 

where 

DDW = committed effective dose 

equivalent from one year’s 

consumption of contaminated 

groundwater (mrem/yr) 

CGW-i = concentration of radionuclide i 

in groundwater (μCi/L) 

UW = human consumption rate of 

water (L/yr) 

DCFi = ingestion dose conversion factor 

for radionuclide i (mrem/pCi). 

The dose through water ingestion by beef and 

milk cattle assumes that cattle drink contaminated 

water. The receptor is then assumed to drink milk 

and eat meat from the cattle that drank the 

contaminated water. The dose from this pathway 

was calculated with the following equations: 
3-63

Table 3-10. The input parameters and their distributions used in these analyses. 

Parameter Distributiona Justification 

Human Consumption Rate 

of Water (UW) 

Uniform on 

[184, 511]L/yr 

Lower end is national average estimated daily intake of 

water-based foods and beverages (including drinking 

water) from Yang and Nelson (1986); upper end is 

national average intake of all liquid food stuffs from 

Yang and Nelson (1986) 

Beef Cattle Consumption 

Rate of Water (QW-BC) 

Uniform on 

[50, 60] L/d 

Lower end is GENII-S default (Leigh et al. 1993); upper 

end is based on International Atomic Energy Agency 

(1994) 

Milk Cattle Consumption 

Rate of Water (QW-MC) 

Uniform on 

[60, 100] L/d 

Lower end is GENII-S default (Leigh et al. 1993); upper 

end is based on International Atomic Energy Agency 

(1994) 

Beef Cattle Consumption 

Rate of Fodder (QF-BC) 

Uniform on 

[6.0, 12.0] kg/d 

Lower end is based on International Atomic Energy 

Agency (1994); upper end is GENII-S default (Leigh et 

al. 1993) (Note: dry weights) 

Milk Cattle Consumption 

Rate of Fodder (QF-MC) 

Uniform on 

[11.0, 14.6] kg/d 

Lower end is GENII-S default (Leigh et al. 1993); upper 

end is based on International Atomic Energy Agency 

(1994) (Note: dry weights) 

Human Consumption Rate 

of Beef (UB) 

Normal 

μ=32 kg/yr 

σ= 0.4 kg/yr 

National average from Yang and Nelson (1986) 

Human Consumption Rate 

of Milk (UM) 

Uniform on 

[20, 180] L/yr 

Range based on lower bound in LaPlante and Poor 

(1997); centroid value based on Hoffman, Gardner, and 

Eckerman (1982) and matches that in Yang and Nelson 

(1986) 

Human Consumption Rate 

of Produce (UP) 

Normal 

μ=166 kg/yr 

σ= 40 kg/yr 

Mean is from LaPlante and Poor (1997); standard 

deviation based on LaPlante and Poor (1997) 

Human Consumption Rate 

of Leafy Vegetables (ULV) 

Uniform on 

[4.3, 28.3] kg/yr 

Based on Maheras et al. (1997) and LaPlante and Poor 

(1997) 

Irrigation Rate (I) Uniform on 

[3.5, 8.5] L/m2

/d 

Lower bound is from LaPlante and Poor (1997); upper 

bound is from Maheras et al. (1997) 

Wash-off Constant (k) Uniform on 

[0.018, 0.032] L/mm 

±25% from value in Maheras et al. (1997) 

Leafy Vegetable 

Interception Fraction per 

Unit Yield (r/YLV) 

Uniform on 

[0.056, 0.096] m2

/kg 

±25% from value in Maheras et al. (1997) 

Produce Interception 

Fraction per Unit Yield 

(r/YP) 

Uniform on 

[0.028, 0.048] m2

/kg 

±25% from value in Maheras et al. (1997) 

Fodder Interception 

Fraction per Unit Yield 

(r/YF) 

Uniform on 

[1.5, 2.5] m2

/kg 

±25% from value in Maheras et al. (1997) 

Soil Areal Density (P) Uniform on 

[180, 270] kg/m2

From LaPlante and Poor (1997) for areal density in top 

15 cm of soil; centroid value matches that used in 

Maheras et al. (1997) 
Table 3-10. (continued). 

3-64

Parameter Distributiona Justification 

Irrigation Time (tI) Uniform on 

[60, 240] d 

Based on LaPlante and Poor (1997) for all vegetation 

types; Maheras et al. (1997) used 90 days, which is in 

range 

Buildup Time for 

Radionuclides in Soil (tb) 

Constant at 365 d Conservative estimate from Maheras et al. (1997) 

Irrigation Fraction of Year 

(fI) 

Irrigation Time/ 365 d First principles 

Leafy Vegetable 

Translocation Factor (TLV) 

Constant at 1.0 Conservative estimate from default values in GENII-S 

(Leigh et al. 1993) 

Produce Translocation 

Factor (TP) 

Uniform on 

[0.075, 0.125] 

±25% from GENII-S default value (Leigh et al. 1993) 

Leafy Vegetable Fraction 

of Activity after Processing 

(DFLV) 

[0.375, 0.625] ±25% from value in Maheras et al. (1997) 

Produce Fraction of 

Activity after Processing 

(DFP) 

Constant at 1.0 Conservative estimate assuming little of produce activity 

is on surface of vegetation (i.e., not washable) 

Site Grown Factor for 

Vegetables and Produce 

(FV) 

Uniform on 

[0.25, 0.7] 

Lower bound is estimate from intruder analysis 

(Section 5); upper bound is from Maheras et al. (1997) 

Site Grown Factor for Beef 

(FB)

Uniform on 

[0.25, 0.442] 

Lower bound is estimate from intruder analysis 

(Section 5); upper bound is from Maheras et al. (1997) 

Site Grown Factor for Milk 

(FM)

Uniform on 

[0.25, 0.399] 

Lower bound is estimate from intruder analysis 

(Section 5); upper bound is from Maheras et al. (1997) 

a. μ=mean and σ=standard deviation. 
3-65

Table 3-11. Parameter values used in the all pathway dose calculation in Maheras (1997). 

Parameter Value Reference 

Human Consumption Rate of Water (UW) 258 L/yr Yang and Nelson (1984) 

Beef Cattle Consumption Rate of Water (QW-BC) 50 L/d NRC (1977) 

Milk Cattle Consumption Rate of Water (QW-MC) 60 L/d NRC (1977) 

Beef Cattle Consumption Rate of Fodder (QF-BC) 12 kg/d National Council on Radiation Protection 

and Measurement (1984) 

Milk Cattle Consumption Rate of Fodder (QF-MC) 16 kg/d National Council on Radiation Protection 

and Measurement (1984) 

Human Consumption Rate of Beef (UB) 85 kg/yr Rupp (1980) 

Human Consumption Rate of Milk (UM) 112 L/yr Rupp (1980) 

Human Consumption Rate of Produce (UP) 176 kg/yr Rupp (1980) 

Human Consumption Rate of Leafy Vegetables 

(ULV) 

18 kg/yr Rupp (1980) 

Irrigation Rate (I) 8.47 L/m2

/d Maheras et al. (1997) 

Wash-off Constant (k) 0.025/mm Peterson (1983) 

Leafy Vegetable Interception Fraction per Unit 

Yield (r/YLV) 

0.076 m2/kg Baes and Orton (1979) and Baes et al. 

(1984) 

Produce Interception Fraction per Unit Yield 

(r/YP) 

0.032 m2

/kg Baes and Orton (1979) and Baes et al. 

(1984) 

Fodder Interception Fraction per Unit Yield 

(r/YF) 

2.0 m2

/kg Baes and Orton (1979) and Baes et al. 

(1984) 

Soil Areal Density (P) 225 kg/m2

 DOE (1987) 

Irrigation Time (tI) 90 d Maheras et al. (1997) 

Buildup Time for Radionuclides in Soil (tb) 365 d Maheras et al. (1997) 

Irrigation Fraction of Year (fI) 0.25 Maheras et al. (1997) 

Leafy Vegetable Translocation Factor (TLV) 1.0 Ng et al. (1978) 

Produce Translocation Factor (TP) 0.1 Ng et al. (1978) 

Leafy Vegetable Fraction of Activity after 

Processing (DFLV) 

0.5 Ng et al. (1978) 

Produce Fraction of Activity after Processing 

(DFP) 

1.0 Ng et al. (1978) 

Site Grown Factor for Vegetables and Produce 

(FV) 

0.7 EPA (1989) 

Site Grown Factor for Beef (FB) 0.442 EPA (1989) 

Site Grown Factor for Milk (FM) 0.399 EPA (1989) 
3-66

G

R

O

U

N

D

W

A

T

E

R

C

O

N

C

E

N

T

R

A

T

I

O

N

Uptake by Livestock

- Beef Cattle Consumption

 Rate of Water

- Milk Cattle Consumption

 Rate of Water

- Beef Cattle Consumption

 Rate of Fodder

- Milk Cattle Consumption

 Rate of Fodder

- Site Grown Factor for Beef

- Site Grown Factor for Milk

DOSE TO

HUMANS

- Human 

 Consumption

 Rate of Water

- Human 

 Consumption

 Rate of Beef

- Human 

 Consumption

 Rate of Milk

- Human 

 Consumption

 Rate of Produce

- Human 

 Consumption

 Rate of Leafy 

 Vegetables

Irrigation of 

Crops/Fodder

- Irrigation Rate

- Irrigation Time

- Fraction of Year Irrigation

- Buildup Time for 

 Radionuclides in Soil

- Soil Areal Density

Uptake by Crops

- Wash-off Constant

- Leafy Vegetable Interception Fraction

 per Unit Yield

- Produce Interception Fraction per Unit

 Yield

- Leafy Vegetable Translocation Fraction

- Produce Translocation Fraction

- Leafy Vegetable Fraction of Activity

 after Processing

- Produce Fraction of Activity after

 Processing

- Site Grown Factor for Vegetables

 and Produce

Uptake by Livestock 

Feed (Fodder)

- Fodder Interception Fraction

 per Unit Yield

Figure 3-23. A block diagram of the all-pathways dose assessment and a description of where the 

parameter values enter the calculation. 

Direct ingestion of beef contaminated by drinking 

water: 

∑

μ − = − × − × × × × × i B DW GW i W BC b B i B Ci

pCi D C Q F U DCF F

6 10 (3-32) 

where 

DB-DW = committed effective dose 

equivalent from one year’s 

consumption of beef contaminated 

by drinking water (mrem/yr) 

QW-BC = water consumption rate by beef 

cattle (L/yr) 

Fb = beef transfer coefficient (d/kg) 

UB = human consumption rate of beef 

(kg/yr) 

FB = fraction of beef produced locally 

(unitless). 

Direct ingestion of milk contaminated by drinking 

water: 

∑

μ − = − × − × × × × × i M DW GW i W MC m M i M Ci

pCi D C Q F U DCF F

6 10 (3-33) 

where 

DM-DW = committed effective dose 

equivalent from one year’s 

consumption of milk 

contaminated by drinking water 

(mrem/yr) 
3-67

QW-MC = water consumption rate by milk 

cattle (L/yr) 

Fm = milk transfer coefficient (d/L) 

UM = human consumption rate of milk 

(L/yr) 

FM = fraction of milk produced locally 

(unitless). 

The dose to humans from ingestion of 

contaminated leafy vegetables and produce was 

calculated assuming two contamination routes: 

direct deposition of contaminated irrigation water 

on plants and deposition of contaminated irrigation 

water on soil followed by root uptake by plants. 

Leafy vegetables and produce were treated 

separately. The dose through direct deposition was 

calculated as follows: 

Leafy vegetables—direct deposition: 

∑





























× × × × ×

× +

− × × ×

=

− + −

− i 6

LV i LV LV V

i

I k I )t i ( λ

LV

GW i

LV D

μCi

10 pCi U DCF DF T F

λ k I

1 e

Y

C I r

D (3-34) 

where 

DLV-D = committed effective dose 

equivalent from one year’s 

consumption of leafy vegetables 

contaminated by direct 

deposition (mrem/yr) 

I = irrigation rate (L/m2/d) 

r = interception fraction (unitless) 

YLV = agricultural yield for leafy 

vegetables (kg/m2

, wet weight) 

λi = radioactive decay constant for 

radionuclide i (d-1) 

k = wash-off constant (mm-1) 

tI = irrigation time (d) 

ULV = human consumption rate of 

leafy vegetables (kg/yr) 

DFLV = fraction of activity remaining 

after preparation and processing 

of leafy vegetables (unitless) 

TLV = translocation factor for leafy 

vegetables (unitless) 

FV = fraction of leafy vegetables and 

produce produced locally 

(unitless). 

Produce—direct deposition: 

∑





























× × × × ×

× +

− × × ×

=

− + −

− i 6

P i P P V

i

I k I )t i ( λ

P

GW i

P D

μCi

10 pCi U DCF DF T F

λ k I

1 e

Y

C I r

D (3-35) 

where 

DP-D = committed effective dose 

equivalent from one year’s 

consumption of produce 

contaminated by direct 

deposition (mrem/yr) 

YP = agricultural yield for produce 

(kg/m2

, wet weight) 

UP = human consumption rate of 

produce (kg/yr) 

DFP = fraction of activity remaining 

after preparation and processing 

of produce (unitless) 

TP = translocation factor for produce 

(unitless). 

The product kI is also known as the 

weathering-rate constant because of wash-off 

(Peterson 1983). This quantity describes the rate at 

which material is removed from plant surfaces by 

water and is analogous to λe , the weathering rate 

constant used in non-irrigation situations. The 

value of kI was calculated as follows: 
3-68

. / d m

, mm

, L

m

m day

. L kI . / mm 0 212

1

1000

1000

8 47 1 0 025

3

2 = × × × =

 (3-36) 

The dose from deposition of contaminated 

irrigation water on soil followed by root uptake by 

plants and human consumption of plants was 

calculated using the following equations. Credit 

was not taken for leaching of radionuclides from 

the root zone of the plants. Also, buildup from 

successive years of farming was not taken into 

account. 

Leafy vegetables—root uptake: 

∑





























× × × ×

× − × × ×

=

− × −

− i 6

i LV i V

i

) bt i ( λ

GW i I

LV R

μCi

10 pCi CR U DCF F

λ

1 e

P

C I f

D (3-37) 

Produce—root uptake: 

∑





























× × × ×

× − × × ×

=

− × −

− i 6

i P i V

i

) bt i ( λ

GW i I

P R

μCi

10 pCi CR U DCF F

λ

1 e

P

C I f

D (3-38) 

where 

DLV-R = committed effective dose 

equivalent from one year’s 

consumption of leafy vegetables 

contaminated by root uptake 

(mrem/yr) 

DP-R = committed effective dose 

equivalent from one year’s 

consumption of produce 

contaminated by root uptake 

(mrem/yr) 

fI = fraction of the year that crops 

are irrigated (unitless) 

P = areal density (kg [dry weight 

soil]/m2

) 

tb = buildup time for radionuclides 

in soil (d) 

CRI = concentration ratio for 

radionuclide i (pCi/kg [wet 

weight plant] per pCi/kg [dry 

weight soil]). 

The dose to humans from ingestion of animal 

products contaminated by contaminated feed was 

calculated assuming two contamination routes: 

direct deposition and root uptake. Meat and milk 

were treated separately. The dose through direct 

deposition was calculated as follows: 

Beef—direct deposition: 

∑





























× × × × ×

× +

− × × ×

=

−

− + −

− i 6

F BC B B i B

i

I k I )t i ( λ

F

GW i

B D

μCi

10 pCi Q TF U DCF F

λ k I

1 e

Y

C I r

D

 (3-39) 

DB-D = committed effective dose 

equivalent from one year’s 

consumption of beef 

contaminated by direct 

deposition onto fodder 

(mrem/yr) 

YF = agricultural yield for fodder 

(kg/m2

, dry weight) 

QF-BC = beef cow consumption rate of 

fodder (kg [dry]/d) 

TFB = transfer factor (fodder to beef)

UB = human consumption rate of beef 

(kg/yr) 

FB = fraction of beef produced 

locally (unitless). 

Milk—direct deposition: 

∑





























× × × × ×

× +

− × × ×

=

−

− + −

− i 6

F MC M M i M

i

I k I )t i ( λ

F

GW i

M D

μCi

10 pCi Q TF U DCF F

λ k I

1 e

Y

C I r

D

 (3-40) 
3-69

where 

DM-D = committed effective dose 

equivalent from one year’s 

consumption of milk 

contaminated by direct 

deposition onto fodder 

(mrem/yr)

QF-MC = milk cow consumption rate of 

fodder (kg [dry]/d) 

TFM = transfer factor (fodder to milk) 

UM = human consumption rate of milk 

(L/yr) 

FM = fraction of milk produced locally 

(unitless). 

The dose through deposition on soil followed 

by root uptake was calculated using the following 

equations. As with produce and leafy vegetables, 

credit was not taken for leaching of radionuclides 

from the root zone of plants. Here again, buildup 

from successive years of farming also was not 

taken into account. 

Beef—root uptake: 

∑





























× × × × × ×

× − × × ×

=

− × −

− i 6

i F B B i B

i

) bt i ( λ

GW i I

B R

μCi

10 pCi CR Q TF U DCF F

λ

1 e

P

C I f

D

 (3-41) 

Milk—root uptake: 

∑





























× × × × × ×

× − × × ×

=

− × −

− i 6

i F M M i M

i

) bt i ( λ

GW i I

M R

μCi

10 pCi CR Q TF U DCF F

λ

1 e

P

C I f

D

 (3-42) 

where 

DB-R = committed effective dose 

equivalent from one year’s 

consumption of beef 

contaminated by root uptake 

(mrem/yr) 

DM-R = committed effective dose 

equivalent from one year’s 

consumption of milk 

contaminated by root uptake 

(mrem/yr) 

fI = fraction of the year that crops 

are irrigated (unitless) 

P = areal density (kg [dry weight 

soil]/m2

) 

tb = buildup time for radionuclides 

in soil (d) 

CRi = concentration ratio for 

radionuclide i (pCi/kg [wet 

weight plant] per pCi/kg [dry 

weight soil]) 

QF = animal consumption rate of 

fodder (kg [dry]/d). 

Secondary and indirect pathways, such as 

inhalation of contaminated irrigation water, 

inhalation of contaminated dust, or external 

exposure from radionuclides deposited on the soil, 

were omitted from this scenario. These pathways 

were assumed to contribute relatively minor 

amounts when compared to direct pathways such 

as direct ingestion of contaminated water. This has 

been found to be the case for similar dose 

assessments for contamination introduced via the 

groundwater pathway (LaPlante and Poor 1997). 

Note that the total human dose to be compared 

to the standards is the total received from all of the 

aforementioned pathways. 

The equations described above are solved 

using a Formula Translation (FORTRAN) 

computer code developed for the TFF PA and 

presented in Appendix J. The code conducts 

1,000 dose realizations based on data sampled 

from the distributions provided in Table 3-10. The 

results are then ordered from lowest to highest, 

based on dose. After this, the 950th realization on 

the list is chosen for the 95% confidence level 

(95% of the realizations are less than this value).
4-1 

4. RESULTS OF ANALYSIS

This section describes the results of the 

analysis of performance for the TFF, conducted in 

accordance with the conceptual models and 

methodologies described in Section 3. Predicted 

releases to the environment, resulting groundwater 

concentrations, radon flux analysis, volatile 

radionuclide atmospheric transport, and the results 

of the dose analysis are presented. 

Prior modeling of the TFF PA was conducted 

and a bounding analysis was completed. The 

results of these prior bounding analyses are 

provided in Appendix F. These analyses were reevaluated based upon comments received from 

NRC and DOE. The focus of the new groundwater 

modeling analyses presented in this section are 

based on the four radionuclides (i.e., 90Sr/90Y, 

99Tc, 129I, and 14C) determined from previous 

modeling (see Appendix F) to provide the majority 

of the groundwater pathway doses. 

The source terms for the groundwater 

transport pathway are described first in 

Section 4.1. The results of the groundwater 

transport calculations are presented in Section 4.2. 

The radon flux results are presented in Section 4.3, 

and the volatile radionuclide atmospheric transport 

results are provided in Section 4.4. The 

groundwater protection doses and all-pathways 

results are presented in Section 4.5. 

4.1 Groundwater Transport 

Source Term 

As discussed in previous sections, DUST-MS 

was used to predict the release rates from the 

various waste forms at the INTEC facility. This 

approach was implemented to minimize 

computational time required by PORFLOW to 

predict release rates. The large hydraulic 

conductivity contrasts of the waste forms 

compared with the surrounding geological 

material required small time steps and 

subsequently time-consuming simulations using 

PORFLOW. However, comparative simulations 

were conducted to predict release rates using both 

DUST-MS and PORFLOW. The results showed 

that both models predicted similar contaminant 

transport rates using PORFLOW with the DUSTMS source term or the source term calculated by 

PORFLOW. These results showed that either 

source model yielded similar predictions on the 

rate and concentration of radionuclides released to 

the environment. In addition, the DUST-MS 

releases were compared to simple leaching 

models. These comparisons are provided in 

Section 7. 

The results of the source term models for the 

groundwater pathway are described in this section. 

The contaminant releases for radon and the 

volatile radionuclides were described in Section 3. 

The results of the radon and volatile radionuclide 

analyses are provided in Sections 4.3 and 4.4, 

respectively. 

4.1.1 Radionuclide Releases from the 

Piping, Sand Pad, and Tanks 

Previous sections have 

• Described the capabilities and features of the 

DUST-MS computer code 

• Summarized the conceptual model used to 

model release rates out of the tank, sand pad, 

and concrete vault floor 

• Provided DUST-MS input parameters used in 

all simulations and nuclide-specific input 

parameters. 

The releases from piping at the TFF are based 

on the model presented in Section 3 for the 

degradation of the piping and the release of 

contaminants. In the previous analysis of the 

piping (see Appendix F), the piping was assumed 

to be released by diffusion to breaks in the piping. 

In this analysis, the piping is assumed to degrade 

in a manner similar to the tanks. Therefore, 

releases of contaminants from the piping begin to 

occur at 500 yr post-closure, the same time as 

releases from the tanks. The piping releases were 

assumed to follow the same release function as 

that for the tanks according to the DUST-MS 

results. 
4-2 

These releases were input into the 

groundwater model to determine groundwater 

concentrations, as discussed in Section 4.2. 

Figure 4-1 displays the results from the 

DUST-MS simulations. The results are reasonable, 

considering the relative magnitudes of the partition 

coefficients and the types of releases. That is, 

releases from the sand pad occur earlier than 

releases from the tanks/piping. This is reasonable 

because radionuclides in the sand pad have a 

shorter transport path than radionuclides in the 

tank waste. 

The output from DUST-MS simulations 

includes the total mass flux. These values were 

checked and they indicate a mass balance between 

releases from the source (initial inventory) and 

mass transported through the mass flux location. 

Additionally, maximum release rates compare 

favorably with hand-calculated values. Thus, the 

DUST-MS results also are quantitatively 

consistent with the expected values. 

As is shown in Figure 4-1, two release peaks 

are present, the earlier peak being the sand pad 

releases beginning at 100 yr and the later peaks 

due to releases from the tanks/piping beginning at 

500 yr. 14C also is not released in significant 

quantities due to the large distribution coefficient 

(Kd) values present in the grouted waste and 

concrete. 

The time of peak releases for each 

radionuclide and the peak release rates are 

provided in Table 4-1. 

The releases determined using the DUST-MS 

computer model were input into the PORFLOW 

groundwater model as time history fluxes 

(Runchal 1997). The groundwater modeling 

results are presented in Section 4.2. 

4.2 Groundwater Transport 

Results 

Contaminant transport simulations were 

conducted for four radionuclides (i.e., 90Sr, 99Tc, 

129I, and14C) that were determined to provide the 

majority of the groundwater pathway doses for the 

1,000-yr compliance period. Prior modeling 

analyses for the TFF, presented in Appendix F, 

assessed the other radionuclides for a period of 

one million years. These results indicate that the 

remaining radionuclides are not of concern for the 

TFF. 

10 -15

10 -10

10 -5

10 0

10 0 10 1 10 2 10 3 1 0 4 10 5 10 6

Sr-90 conservative inventory

Tc-99 conservative in ventory

I-129 conservative in ventory

C-14 conservative inventory

Release Rate (Ci/yr)

Tim e (yr)

Figure 4-1. Releases of 90Sr, 99Tc, 129I, and 14C from the sand pad and tanks and piping.
4-3 

Table 4-1. Peak releases rates, groundwater concentration, and doses for the groundwater pathway 

analyses. 

Peak Release Rate 

Peak Groundwater 

Concentration 

Peak Drinking Water 

Dose Peak All-Pathway Dose 

Nuclide 

Time 

(yr) Ci/yr 

Time 

(yr) pCi/L 

Time 

(yr) mrem/yr 

Time 

(yr) mrem/yr 

90Sr/90Y 232 0.005 551 0.01 551 0.001 551 0.006 

99Tc 12,206 5.0E!04 1.459E+04 116 1.459E+04 0.12 1.459E+04 0.87 

129I 867 1.3E!05 890 3.8 890 0.77 890 1.35 

14C 1.1998E+05 1.6E!14 1.216E+05 4.1E!09 1.216E+05 6.3E!12 1.216E+05 2.9E!11 

Radionuclide releases were placed in the 

PORFLOW model domain where the sand pad and 

tanks were located. Degradation of the piping, 

vaults, and tanks were incorporated into the 

releases determined using the DUST-MS model. 

Radionuclide concentrations were observed in the 

regional aquifer at a point of maximum 

concentration. 

Figure 4-2 shows a cross section of the 

unsaturated/saturated zone contaminant 

distribution of 99Tc resulting from contaminant 

releases. 99Tc was used in the figure because it has 

a long half-life and a source release that results in 

the center line of the plume being visible 

throughout the model domain. Therefore, 99Tc was 

the best radionuclide for visual presentation. Other 

radionuclides were checked to make sure this was 

the point of entry of the plume center line. 

Based on this contaminant distribution, the 

maximum observed contaminant concentrations 

occurs approximately 600 m (2,000 ft) 

downgradient from the tanks.a

 At this location, the 

nodes for the upper 10 m (30 ft) of the aquifer 

were averaged to obtain a contaminant 

concentration representative of a well water 

screen. Results of individual contaminant 

concentrations at this location for a one million 

year simulation are presented in Figure 4-3. 

The model simulations were conducted using 

separated sources for the piping, sand pad, and 

 

a. Model results were output in metric units. Therefore, in 

Section 4, measurements are given in metric units first, 

followed by U.S. customary units. 

tanks. These sources were simulated separately to 

assess the impact of the individual sources. The 

individual sources were then summed to obtain the 

total concentration in groundwater for the 

subsequent dose calculations. 

4.2.1 General Transport Characteristics 

The hydrostratigraphy, location of perched 

water zones, and transport characteristics of the 

individual radionuclides all impacted the results of 

the transport simulations. The perched water 

zones, or the absence thereof, have an important 

impact on the migration of contaminants through 

the vadose zone. Perched zones from seepage of 

the Big Lost River act as sources of water to the 

surrounding unsaturated material. As contaminants 

approach the perched zones, the total hydraulic 

head (defined as both capillary and gravitational 

forces) redirects the migration path around the 

perched water. While some contaminants enter the 

perched zones, the bulk of the flow is diverted 

around the perched zones (see Figure 4-2). Breaks 

in the perched zones where the impeding interbeds 

are absent play a major role in the downward 

migration of contamination. Consequently, the 

hydrostratigraphy defined by Anderson’s USGS 

report plays a major role in the migration patterns 

of the subsurface contaminants (Anderson 1991). 

This influence of the hydrostratigraphy is 

illustrated in Figure 4-2. As noted in Section 3, the 

low permeability of the D-DE2 interbed is a major 

factor in the formation of perched water bodies in 

the upper vadose zone. The hydrologic cross 

section for the site (shown in Section 3) shows that 

this interbed is discontinuous across the site. 

Hydraulic forces divert the downward migration of 

the contaminant plume to areas where this interbed 
4-4 

Figure 4-2. Groundwater modeling domain showing 99Tc concentrations and location of maximum 

concentrations (all concentrations based on a unit source inventory). 

10-14

10-12

10-10

10-8

10-6

10-4

10-2

100

102

100 101 102 103 104 105 106

I-129

Tc-99

Sr-90

C-14

Groundwater Concentrations (pCi/L)

Time (yr)

Conservative Grout Kd's

Conservative Transport

4.1 cm/yr Infiltration

Conservative Inventory

Figure 4-3. Groundwater concentrations for the conservative analysis. 
4-5 

is absent. The net result is that the center of the 

contaminant plume (where the highest 

concentrations enter the regional aquifer) is 

approximately 600 m (2,000 ft) downgradient of 

the release point. To a lesser degree, the 

discontinuity of the lower perched zones also 

impacts the migration patterns of the 

contaminants. 

By the time the contaminants reach the water 

table, the difference in concentrations and arrival 

times below the two tanks in the groundwater 

modeling is diminished. The effects of the 

hydrostratigraphy and the lower perched water 

zones appear to cause a more uniform transport 

rate to the water table. Simulations were 

conducted using separate sources for the tanks 

(i.e., north and south tanks in the cross section); 

however, the resulting contaminant concentrations 

and arrival times were similar. Therefore, all 

contaminant runs were evaluated by placing the 

releases from two tanks at the location of the 

southern tank in the groundwater model domain. 

The timing of contaminant releases from the 

sand pad and tanks/piping is evident from the 

observation point located 600 m (2,000 ft) 

downgradient in the groundwater. As expected, the 

arrival time and the time for peak concentrations 

to occur directly relate to the time of release. 

Several factors affect the transport 

characteristics of the various radionuclides. 

Factors such as the Kd values and radionuclide 

half-lives are interrelated, and all of the factors 

impact the transport characteristics. 

4.2.2 Results for Individual 

Radionuclides 

The following sections discuss the transport 

characteristics of the individual radionuclides that 

resulted in significant aquifer concentrations and 

doses in the transport model. These radionuclides 

include 90Sr, 99Tc, 129I, and 14C. The remaining 

radionuclides were previously evaluated and were 

determined to provide no significant contribution 

to the groundwater pathway doses (see 

Appendix F). 

The peak groundwater concentrations and the 

time of the peak are listed in Table 4-1. 

4.2.2.1 Strontium-90. Peak concentrations 

for the groundwater occur for the sand pad source 

after approximately 550 yr at a concentration of 

0.01 pCi/L. The peak groundwater concentration 

resulting from release from the tanks occurs 

approximately 1,110 yr after closure at a 

concentration of 1.8E!05 pCi/L. The peak 

groundwater concentration from the piping 

releases occurs at 1,110 yr after closure at a 

concentration of 4.4E!08 pCi/L.

4.2.2.2 Technetium-99. Peak 

concentrations for the groundwater occur for the 

sand pad source after approximately 150 yr at a 

concentration of 9.3E!09 pCi/L. The peak 

groundwater concentration resulting from release 

from the tanks occurs approximately 1.46E+04 yr 

after closure at a concentration of 116 pCi/L. The 

peak groundwater concentration from the piping 

releases occurs at 1.46E+04 yr after closure at a 

concentration of 0.27 pCi/L.

4.2.2.3 Iodine-129. Peak concentrations for 

the groundwater occur for the sand pad source 

after approximately 421 yr at a concentration of 

8.9E!04 pCi/L. The peak groundwater 

concentration resulting from release from the tanks 

occurs approximately 888 yr after closure, at a 

concentration of 3.82 pCi/L. The peak 

groundwater concentration from the piping 

releases occurs at 888 yr after closure at a 

concentration of 0.02 pCi/L.

4.2.2.4 Carbon-14. Peak concentrations for 

the groundwater occur for the sand pad source 

after approximately E+05 yr at a concentration of 

3.7E!14 pCi/L. The peak groundwater 

concentration resulting from release from the tanks 

occurs approximately 1.2E+05 yr after closure at a 

concentration of 4.1E!09 pCi/L. The peak 

groundwater concentration from the piping 

releases occurs at 1.2E+05 yr after closure at a 

concentration of 2.4E!16 pCi/L.

4.3 Radon Analysis Results 

As discussed in Section 3, calculations were 

performed for a one-dimensional bare slab resting 
4-6 

on the ground surface to determine the radon flux 

at the surface of this slab. These calculations are 

highly conservative with respect to the actual 

performance of a grouted heel at the TFF. Because 

the slab is modeled in open air, no credit is taken 

for the vault or tank. Also, the radium inventory at 

the end of the 1,000-yr performance period is used 

for the source in the calculations, which is 

conservative for all times before the end of the 

performance period because of ingrowth from 

parent radionuclides. Modeling the slab in open air 

means radon can more easily diffuse out of the 

slab than would otherwise be expected. The model 

did not include pressure-driven transport of radon 

out of the slab for two reasons. First, cementitious 

materials are generally too impermeable for 

pressure-driven transport to be significant (Renken 

and Rosenberg 1995; Tanner 1990; Rogers and 

Nielson 1992). Second, it is difficult to envision a 

situation where a significant pressure gradient is 

maintained across the slab when it is buried at 

depth. In all cases for the radon analysis, the point 

of compliance is the surface of the TFF soils. 

The conservative inventory 226Ra 

concentration (1.92E−04 Ci/tank; 

1.73E+06 pCi/m3

) resulted in a flux of 222Rn out of 

the surface of the slab of 0.39 pCi/m2

/s after 

converting to pCi from Bq (1 pCi is equal to 

0.037 Bq). This value is more than an order of 

magnitude below the standard of 20 pCi/m2

/s, 

given in DOE Manual 435.1-1 (DOE 2001a). The 

reason that such a conservative model was still 

able to meet the standard is that there is simply not 

much 226Ra in the slab. Also, cement materials do 

not emanate or transport radon very well. 

Since the radioactive decay constants for 220Rn 

and 219Rn are well over three orders of magnitude 

larger than that of 222Rn and the natural log term is 

always less than zero, the reduction factor for the 

short-lived nuclides is vanishingly small for all but 

the thinnest of overburdens (i.e., where the 

reduction factor for 222Rn approaches unity). For 

reference, a 0.5-m-thick (2-ft) overburden of soil 

with a diffusion coefficient of 7E−06 m2

/s (the 

diffusion coefficient for radon in air, a 

conservative upper limit for radon diffusion in any 

medium [Rogers and Nielson 1992]) would have a 

reduction factor of about 1096, with the reduction 

factor for 219Rn being even smaller. 

4.4 Volatile Radionuclide 

Analysis Results 

Based on the conservative inventory, volatile 

radionuclide release rates determined in Section 3 

were input into the CAP88-PC computer model 

(EPA 1992) to determine the potential doses from 

the atmospheric pathway. The receptor was 

located 100 m (300 ft) from the TFF. The analysis 

used 1992 wind data from the INEEL Grid-3 

meteorological station. Standard default values for 

agricultural products, based on the rural definition 

in CAP88-PC, were used in the analysis. The 

outputs from the modeling runs are provided in 

Appendix I. 

The results of the CAP88-PC atmospheric 

transport runs for 3

H releases were determined to 

be 0.027 mrem/yr and 14C doses were 

1.28 mrem/yr, with a total dose of 1.3 mrem/yr. 

The point of compliance was at 100 m (300 ft) in a 

south-southwest direction from the TFF, which 

was the point of maximum dose. The performance 

objective for the air pathway is 10 mrem/yr. 

Therefore, the air transport pathway was found to 

be approximately 1/10th of the allowable dose 

limit. 

4.5 Dose Analysis Results 

4.5.1 Groundwater Protection 

The groundwater pathway analysis was 

evaluated using conservative parameters. The 

analysis consisted of the conservative grout 

distribution coefficients, conservative transport 

parameters, 4.1 cm/yr (1.6 in./yr) infiltration, and 

the conservative inventory discussed previously in 

Sections 2 and 3. 

The drinking water doses were calculated 

using Equation (3-31) in Section 3. The drinking 

water rate of 730 L/yr was used in the analysis. 

The point of compliance was at the point of 

maximum groundwater concentrations in the 

aquifer, 600 m (2,000 ft) south of the TFF. 

The drinking water doses are shown in 

Figure 4-4. The peak drinking water doses and 

their time of peak for each radionuclide are listed 

in Table 4-1. 
4-7 

10-15

10-13

10-11

10-9

10-7

10-5

10-3

10-1

101

100 101 102 103 104 105 106

I-129 

Tc-99 

Sr-90/Y-90

C-14

Drinking Water Dose (mrem/yr)

Time (yr)

Conservative Grout Kd's

Conservative Transport

4.1cm/yr Infiltration

Conservative Inventory

Figure 4-4. Drinking water doses for the conservative analysis. 

4.5.1.1 Strontium-90. Peak drinking water 

doses occur for the sand pad source after 

approximately 550 yr at a dose of 0.001 mrem/yr. 

The peak drinking water doses resulting from 

release from the tanks occur approximately 

1,110 yr after closure at a dose of 

2.0E!06 mrem/yr. The peak drinking water dose 

from the piping releases occurs at 1,110 yr after 

closure at a dose of 4.9E!09 mrem/yr.

4.5.1.2 Technetium-99. Peak drinking 

water doses occur for the sand pad source after 

approximately 150 yr at a dose of 

9.9E!12 mrem/yr. The peak drinking water dose 

resulting from release from the tanks occurs 

approximately 1.46E+04 yr after closure at a dose 

of 0.12 mrem/yr. The peak drinking water dose 

from the piping releases occurs at 1.46E+04 yr 

after closure at a dose of 2.9E!04 mrem/yr.

4.5.1.3 Iodine-129. Peak drinking water 

doses occur for the sand pad source after 

approximately 421 yr at a dose of 

1.8E!04 mrem/yr. The peak drinking water dose 

resulting from release from the tanks occurs 

approximately 888 yr after closure at a dose of 

0.77 mrem/yr. The peak drinking water dose from 

the piping releases occurs at 888 yr after closure at 

a dose of 0.004 mrem/yr.

4.5.1.4 Carbon-14. Peak drinking water 

doses occur for the sand pad source after 

approximately 1E+05 yr at a dose of 

5.7E!17 mrem/yr. The peak drinking water dose 

resulting from release from the tanks occurs 

approximately 1.2E+05yr after closure at a dose of 

6.3E!12 mrem/yr. The peak drinking water dose 

from the piping releases occur at 1.2E+05 yr after 

closure at a dose of 3.6E!19 mrem/yr.

The doses reported above are for the entire 

simulation time of one million years after closure 

of the TFF. The compliance period of 1,000 yr 

exhibited a peak dose of 0.77 mrem/yr from 129I. 

To ensure that the drinking water pathway 

doses are bounding, a sensitivity/uncertainty 

analysis was conducted. A discussion of the 

sensitivity/uncertainty results is provided in 

Section 7. 

4.5.2 All-Pathways Dose 

The groundwater all-pathway dose analysis 

was evaluated using conservative parameters. The 

analysis consisted of the conservative grout 

distribution coefficients, conservative transport 

parameters, 4.1 cm/yr (1.6 in./yr) infiltration, and 

the conservative inventory discussed previously in 

Sections 2 and 3. 
4-8 

10-15

10-13

10-11

10-9

10-7

10-5

10-3

10-1

101

100 101 102 103 104 105 106

I-129 

Tc-99 

Sr-90/Y-90

C-14

All Pathways Dose (mrem/yr)

Time (yr)

Conservative Grout Kd's

Conservative Transport

4.1cm/yr Infiltration

Conservative Inventory

Figure 4-5. All-pathway doses for the conservative analysis. 

The all-pathways doses are shown in 

Figure 4-5. The peak all-pathways doses and the 

tie of the peak for each radionuclide are 

summarized in Table 4-1. 

4.5.2.1 Strontium-90. Peak all-pathways 

doses occur for the sand pad source after 

approximately 550 yr at a dose of 0.005 mrem/yr. 

The peak all-pathways doses resulting from 

release from the tanks occur approximately 

1,110 yr after closure at a dose of 

1.0E!05 mrem/yr. The peak all-pathways dose 

from the piping releases occurs at 1,110 yr after 

closure at a dose of 2.5E!08 mrem/yr.

4.5.2.2 Technetium-99. Peak all-pathways 

doses occur for the sand pad source after 

approximately 150 yr at a dose of 

7.0E!11 mrem/yr. The peak all-pathways dose 

resulting from release from the tanks occurs 

approximately 1.46E+04 yr after closure at a dose 

of 0.87 mrem/yr. The peak all-pathways dose 

from the piping releases occurs at 1.46E+04 yr 

after closure at a dose of 2.0E!03 mrem/yr.

4.5.2.3 Iodine-129. Peak all-pathways doses 

occur for the sand pad source after approximately 

421 yr at a dose of 3.1E!04 mrem/yr. The peak 

all-pathways dose resulting from release from the 

tanks occurs approximately 888 yr after closure at 

a dose of 1.4 mrem/yr. The peak all-pathways 

dose from the piping releases occurs at 888 yr 

after closure at a dose of 0.006 mrem/yr.

4.5.2.4 Carbon-14. Peak all-pathways doses 

occur for the sand pad source after approximately 

1E+05yr at a dose of 2.6E!16 mrem/yr. The peak 

all-pathways dose resulting from release from the 

tanks occur approximately 1.2E+05 yr after 

closure at a dose of 2.9E!11 mrem/yr. The peak 

all-pathways dose from the piping releases occurs 

at 1.2E+05 yr after closure at a dose of 

1.7E!18 mrem/yr.

The doses reported above are for the entire 

simulation time of 1 million years after closure of 

the TFF. The compliance period of 1,000 yr 

exhibited a peak dose of 1.4 mrem/yr from 129I. 

To ensure that the all-pathway doses are 

conservative, a sensitivity/uncertainty analysis 

was conducted. A discussion of the 

sensitivity/uncertainty results is provided in 

Section 7.
5-1 

5. INADVERTENT INTRUDER ANALYSES 

This section describes the potential intruder 

scenarios for the TFF. The intruder scenarios are 

screened based on the physical aspects of the 

facility (i.e., depth to the waste). 

Intruder scenarios considered for use in this 

PA were limited to those previously described for 

low-level radioactive waste performance 

assessments (NRC 1982; Kennedy and Peloquin 

1988). These intruder scenarios include both acute 

and chronic exposure scenarios. Acute exposure 

scenarios involve exposures of short duration, and 

include an intruder-construction scenario, a 

discovery scenario, and a drilling scenario. 

Chronic, longer-duration, intruder scenarios 

include the intruder-agriculture, intruder-resident, 

and post-drilling scenario. The Addendum to 

Radioactive Waste Management Complex LowLevel Waste Radiological Performance 

Assessment (Maheras et al. 1997) considered two 

additional intruder scenarios at the INEEL: 

chronic intruder-radon and chronic biointrusion. 

All eight of these scenarios were screened for use 

in this PA and are discussed in this section. 

The intruder scenarios used to support the 

waste classification limits in 10 CFR 61 (2003) 

were the intruder-construction, intruder-discovery, 

and intruder-agriculture scenarios. The more 

restrictive of the scenarios, intruder-construction 

and intruder-agriculture, were used for setting the 

Class A and Class C waste classification limits. 

The intruder-discovery scenario was used for 

setting the Class B waste limits for short-lived 

radionuclides (Oztunali and Roles 1986; NRC 

1982). 

When waste is disposed at depths greater than 

5 m (20 ft), the intruder scenarios defined by NRC 

for surface excavation may not directly apply.a

Disposal with special waste forms or at a depth 

greater than 5 m (20 ft) (which applies to the waste 

 

a. Model inputs and outputs are given in metric units. 

Therefore, in Section 5, measurements are given in metric 

units first, followed by U.S. customary units. 

in the tanks and sand pads addressed by this PA), 

resembles the conditions anticipated for “Greater 

Confinement Disposal” operations for low-level 

waste at many of the operating DOE burial 

grounds (Oztunali and Roles 1986). These 

conditions require the definition of additional 

intruder scenarios. Kennedy and Peloquin (1988) 

and Oztunali and Roles (1986) present intruderdrilling and post-drilling scenarios to deal with 

waste disposed at depths greater than 5 m (20 ft). 

The remainder of this section explores the 

intruder scenario definitions, presents the basis for 

the selection of appropriate intruder scenarios for 

the TFF, and explains the development of the 

mathematical representations and parameters of 

the intruder scenarios for both the acute and 

chronic exposures. 

It should be noted that all of the intruder 

scenarios presented in this section were based on 

the conservative inventories presented in 

Section 2.3. 

5.1 Intruder Scenario 

Screening 

5.1.1 Intruder Inventory 

The intruder analysis considers that active 

institutional control will be maintained over the 

disposal site for at least 100 yr after facility 

closure. This consideration eliminates from 

concern any radionuclide with a half-life less than 

about 5 yr, because the inventory in the waste at 

100 yr after facility closure will be reduced to 

innocuous levels by radioactive decay (unless it 

appears in the initial inventory in extremely high 

concentrations). However, radionuclides with halflives less than 5 yr cannot be neglected if the 

radionuclide appears in a decay chain, since its 

activity may increase with time because of the 

decay of a parent radionuclide, unless the parent is 

short-lived. 

Using the 5-yr half-life screening criteria 

results in the elimination of the following 

radionuclides from further consideration: 102Rh, 
5-2 

119mSn, 134Cs, 150Eu, 153Gd, 155Eu, and 171Tm. In 

addition, the following radionuclides in short 

decay chains were eliminated from further analysis 

since the parent and progeny each have half-lives 

less than 5 yr: 106Ru/106Rh, 109Cd/109mAg, 

110mAg/110Ag, 125Sb/125mTe, and 144Ce/144mPr/144Pr. 

Additional radionuclides were screened from 

further consideration because their half-lives 

indicate that they are either stable or have such 

long half-lives that their specific activity and thus, 

contribution to dose, would be insignificant. These 

radionuclides include 142Ce (5E+16 yr), 149Sm 

(1E+15 yr), 144Nd (2.4E+15 yr), and 148Sm 

(2E+14 yr). 

5.1.2 Acute Intruder Scenarios 

As suggested in the PA guidance (DOE 

1999b), three potential acute intruder exposure 

scenarios were considered: the intruderconstruction, intruder-discovery, and intruderdrilling scenarios. These scenarios are described 

below. Scenarios not considered applicable to the 

TFF were screened out from further consideration. 

5.1.2.1 Intruder-Construction Scenario. 

The intruder-construction scenario involves an 

inadvertent intruder who chooses to excavate or 

construct a building on the disposal site. In this 

scenario, the intruder is assumed to dig a basement 

excavation to a depth of approximately 3 m (10 ft) 

(Oztunali and Roles 1986). It is assumed that the 

intruder does not recognize the hazardous nature 

of the material excavated. He or she is exposed to 

radioactive constituents in the waste during the 

excavation of the basement. The intruder also is 

exposed to the exhumed waste by inhalation of 

resuspended contaminated soil and external 

irradiation from contaminated soil. Potato cellar 

excavation also was considered since these cellars 

are common in the agricultural region surrounding 

the INEEL. However, potato cellars are relatively 

shallow, with a typical depth of 1 m (3 ft) 

(Maheras et al. 1994, 1997). 

The depth to the residual waste at the bottom 

of the tanks and/or sand pads (i.e., 3 m [10 ft] of 

soil plus 9 m [30 ft] of concrete) would preclude 

direct contact with the waste from the 3-m (10-ft) 

excavation. However, approximately 30% of the 

contaminated piping associated with the tanks is 

located within 3 m (10 ft) of the surface. 

Due to the disposal depth of the waste in the 

tanks at the TFF (i.e., greater than 10 m [30 ft]) 

and the depth of the sand pad contamination, the 

intruder-construction scenario was not considered 

applicable to these waste. However, the intruderconstruction scenario was considered for the 

radionuclide inventory located in the associated 

piping less than 3 m (10 ft) from the surface. 

5.1.2.2 Intruder-Discovery Scenario. The 

intruder-discovery scenario is conceptualized as a 

modification of the intruder-construction scenario. 

The basis for the intruder-discovery scenario is the 

same as the intruder-construction scenario except 

that the exposure time is reduced (Oztunali and 

Roles 1986). The scenario involves the intruder 

excavating a basement to a 3-m (10-ft) depth. The 

intruder is assumed to recognize that he or she is 

digging into very unusual soil immediately upon 

encountering the vault/tank/piping system and 

leaves the site. Consequently, the exposure time is 

reduced. The time typically applied to this 

scenario is 6 hr. 

The depth to the residual waste at the bottom 

of the tanks and/or sand pads (i.e., 3 m [10 ft] of 

soil plus 9 m [30 ft] of concrete) would preclude 

direct contact with the waste from the 3-m (10-ft) 

excavation. However, approximately 30% of the 

contaminated piping associated with the tanks is 

located within 3 m (10 ft) of the surface. 

The intruder-discovery scenario was not 

considered for further analysis since the exposure 

time of 6 hr would result in limited doses in 

comparison to the intruder-construction scenario 

with an exposure time of 160 hr which was 

retained for analysis of the piping radionuclide 

inventory. 

5.1.2.3 Intruder-Drilling Scenario. The 

intruder-drilling scenario assumes the short-term 

exposure of a hypothetical intruder to drill cuttings 

from a borehole penetrating the waste disposal 

site. This scenario involves wastes buried below 

the depth of typical construction excavations.
5-3 

Oztunali and Roles (1986) indicate that for 

waste below 10 m (30 ft), the only applicable 

intrusion scenario is the intruder-drilling scenario. 

They also note that for grouted waste or waste 

disposed of in reinforced concrete structures, a 

time period of 500 yr after site closure is assumed 

as the effectiveness limit for this waste form. 

Therefore, this scenario is not considered 

applicable for drilling through reinforced concrete 

until 500 yr post-closure. However, they also note 

that the scenario is assumed to be fully applicable 

for grouted waste at any time. Because the waste 

could be contacted and moved to the surface, 

thereby exposing the intruder, the intruder-drilling 

scenario was retained for analysis in the TFF PA. 

5.1.3 Chronic Exposure Scenarios 

As suggested in the PA guidance (DOE 

1999b), five potential chronic intruder exposure 

scenarios were considered: the intruderagriculture, intruder-resident, intruder-radon, 

biointrusion, and post-drilling scenarios. These 

scenarios are described below. Those scenarios not 

considered applicable to the TFF were screened 

out from further consideration. 

5.1.3.1 Intruder Post-Construction 

Scenario. The chronic intruder post-construction 

(i.e., agriculture) scenario is an extension of the 

acute intruder-construction scenario. It is assumed 

in this scenario that an intruder lives in the 

building constructed as part of the intruderconstruction scenario and engages in agricultural 

activities on the contaminated site. The intruder is 

exposed to contamination by inhalation of 

resuspended contaminated soil, inhalation of 

gaseous radionuclides released from the waste, 

external irradiation, ingestion of contaminated 

soil, ingestion of contaminated beef and milk, and 

ingestion of contaminated vegetables. 

As stated earlier, the intruder-construction 

scenario was only considered applicable to the 

piping associated with the TFF that is less than 

3 m (10 ft) from the surface. The depth of the 

residual waste in the tanks and the sand pads 

prevents its excavation to the surface. Therefore, 

the intruder-agriculture scenario was retained for 

the piping inventory. 

5.1.3.2 Intruder-Resident Scenario. The 

intruder-resident scenario assumes that the 

intruder constructs a residence on the waste form 

(i.e., concrete vault) after an excavation or some 

natural process exposes it. This scenario was not 

considered applicable to the TFF because of the 

depth of the waste and the shielding provided by 

the overlying grout and concrete. This shielding, 

along with the shielding provided by the house 

foundation, would reduce the external dose rates to 

very low levels. Therefore, the intruder-resident 

scenario was not considered for further analysis.

5.1.3.3 Intruder-Radon Scenario. The 

RWMC PA (Maheras et al. 1997) evaluated the 

intruder-radon scenario. The intruder-radon 

scenario assumes that the intruder excavates a 

20 × 10 × 3-m (70 × 30 × 10-ft) basement over the 

waste while constructing a home. The intruder is 

exposed to 222Rn and its short-lived progeny while 

residing in the home. The exposure from radon 

emanating from the waste and migrating into the 

home is evaluated. 

DOE Manual 435.1-1 (DOE 2001a) states that 

the intruder dose analyses are to exclude the total 

effective dose equivalent (TEDE) contribution 

from radon in air. Therefore, the RWMC chronic 

intruder-radon scenario was not considered for 

further analysis. 

5.1.3.4 Biointrusion Scenario. The 

biointrusion scenario, assessed in the RWMC PA 

(Maheras et al. 1997), assumes that an intruder 

moves onto the site but does not excavate into the 

waste. Rather, radioactivity is brought to the 

surface by plants through root uptake and by 

burrowing animals. The RWMC discussion is 

given below.

Groves and Keller (1983) identified 10 species 

of small mammals nesting on or near the RWMC. 

Four species were the most numerous: deer mice 

(Peromyscus maniculatus), montane voles 

(Microtus montanus), Ord’s kangaroo rats 

(Dipodomys ordii), and Townsend’s ground 

squirrels (Spermophilus townsendii). Reynolds and 

Wakkinen (1987) studied the burrow depths of 

these four species in undisturbed soils. None of the 

deer mice burrows extended past 60 cm (20 in.), 

none of the montane vole burrows extended past 
5-4 

70 cm (30 in.), and none of the Ord’s kangaroo rat 

burrows extended past 100 cm (40 in.). The 

maximum reported burrow depths for undisturbed 

soils was 138 cm (54.3 in.) for the Townsend’s 

ground squirrel. 

Harvester ants (Pogonomyrmex salinus) also 

were considered. Harvester ants burrow deeper 

than the small mammals. Blom, Clark, and 

Johnson (1991) state that harvester ants have been 

found as deep as 2.7 m (8.9 ft) in Wyoming and at 

the Hanford Site. 

Reynolds and Fraley (1989) studied plant root 

profiles near the RWMC and determined the 

maximum rooting depth for big sagebrush 

(Artemisia tridentata) was 225 cm (88.6 in.), green 

rabbitbrush (Chrysothamnus viscidiflorus) was 

190 cm (75 in.), and Great Basin wildrye (Leymus 

cinereus) was 200 cm (80 in.). 

These studies give evidence that the 

biointrusion scenario is not applicable to the tanks 

and sand pads at the TFF because of the depth of 

the waste. The animal burrows and plant roots 

would not reach the residual waste in the tanks or 

sand pads. The animal burrows could contact 

waste in the piping located within the burrow 

depths. However, the amount of contamination 

excavated from the animal burrows would be far 

less than that involved in the intruderconstruction/agricultural scenarios and this 

scenario would be bounded by the intruderconstruction scenario for piping. Therefore, the 

biointrusion scenario was not considered further in 

this PA. 

5.1.3.5 Post-Drilling Scenario. The 

chronic post-drilling scenario is an extension of 

the acute drilling scenario. It assumes that the 

intruder occupies the site after drilling a water well 

and grows crops on a mixture of clean soil and 

contaminated drill cuttings. After exhumation of 

the waste, the exposure pathways are the same as 

for the intruder-agriculture scenario. This intruder 

scenario was retained for further analysis.

5.2 Acute Intruder-Drilling 

Scenario Definition 

The acute drilling scenario assumes that an 

inadvertent intruder drills a well into the contents 

of the tank and vault system. The intruder is 

exposed to contaminated drill cuttings spread over 

the ground and contaminated airborne dust. In the 

standard drilling scenario used in many PAs, the 

intruder is assumed to be exposed to contaminated 

drill cuttings in a mud pit. However, site-specific 

information developed through interviews with 

local well drilling contractors in the Idaho Falls 

area indicates that drillers spread the cuttings over 

the ground and do not use mud pits (Seitz 1991). 

The authors of the RWMC PA (Maheras et al. 

1997) used this site-specific deviation of the 

standard drilling scenario and it also was 

incorporated into the TFF intruder-drilling 

scenario. The assumption that the drill cuttings are 

spread over the ground will result in higher dose 

estimates than if the cuttings were assumed to be 

in a mud pit because of the decrease in the 

shielding factor. 

The drill cuttings are assumed to be spread 

over a 2,200 m2

 (24,000 ft2) lot, which 

corresponds to about one-half acre. Typical lot 

sizes located outside of the Idaho Falls city limits 

are typically 1 to 3 acres. Therefore, a 2,200-m2

 lot 

size was considered conservative for use in the 

RWMC Radiological PA (Maheras et al. 1997), 

and also was incorporated into the TFF intruderdrilling scenario. 

Well drilling contractors in the Idaho Falls 

area have reported that two types of wells are 

typically drilled: small-diameter residential wells 

and large irrigation wells. The small residential 

wells are typically 6 to 8 in. in diameter, serve a 

single residence, and also may provide enough 

water for a family garden and small quantities of 

livestock. The large-diameter irrigation wells are 

drilled to serve systems that irrigate hundreds of 

acres; the wells are located in the middle of farm 

fields, not near the farmer’s residence. Therefore, 

a farmer would not drill an irrigation well to 

acquire water for his residence. Large-diameter 

irrigation wells are currently drilled in 18-in. 

diameters but drilling contractors indicated that 
5-5 

22-in.-diameter irrigation wells would be drilled in 

the future (Seitz 1991). An acute drilling exposure 

could result from drilling either a 6-in.-diameter 

residential well or 22-in.-diameter irrigation well. 

The larger 22-in.-diameter irrigation well was 

assessed for the acute intruder-drilling scenario. 

The intruder is assumed to reside by the 

contaminated cuttings for 160 hr, the time local 

Idaho Falls well drilling contractors say it would 

take to drill and develop a 22-in.-diameter 

irrigation well (Seitz 1991). The exposure 

pathways for this acute drilling scenario include 

inhalation of resuspended drill cuttings, external 

exposure to the ground source, and inadvertent soil 

ingestion. Figure 5-1 illustrates the acute intruderdrilling scenario. 

The TFF concrete vaults are reinforced 

concrete but the grout between the tank and vault 

and also inside of the tank is not reinforced 

concrete. However, considering that the depth to 

the waste in the tank is 10 m (30 ft), that the waste 

is contained within a stainless-steel tank, and that 

the overlying concrete vault roof is reinforced, 

credit beyond 100 yr could be taken for a barrier to 

intrusion. However, the intruder-drilling scenario 

was assumed to begin 100 yr after closure. 

The activity concentration of radionuclides in 

the drill cuttings was determined as follows: 

π ρ

π

well b

well w

s i w i

r D

r t C C 2

2

,

12

, = 1 ×10 × × (5-1) 

where 

Cs,i = soil activity concentration of 

radionuclide i (pCi/g) 

1 × 1012 = factor for converting Ci to pCi 

Cw,i = waste activity concentration 

of radionuclide i at the time of 

intrusion (Ci/m3

) 

rwell = radius of the well borehole 

(m) 

tw = thickness of the waste zone (m) 

Db = total depth of the borehole (m) 

ρ = bulk density of the cuttings 

(g/m3

) 

since 

Tank w

Tank

w i

r t

Ci C , 2 π = (5-2) 

where 

CiTank = activity in the tank (Ci) 

rTank = radius of the tank (m). 

Then by substitution 

π ρ

π

π well b

well w

Tank w

Tank

s i

r D

r t

r t

Ci C 2

2

2

12

, = 1×10 × × (5-3) 

Simplifying results in 

π Tank bρ

Tank

s i

r D

Ci C 2

12

, = 1×10 × (5-4) 

This equation is independent of the waste 

thickness or grout mixing thickness. Also, the 

radius of the well is not present in the equation. 

However, it must be noted that the concentration 

in the soil will be the same for any given well 

radius but the total amount of contaminated soil 

will vary with the well radius. This equation 

assumes that the radionuclide activity is 

completely mixed with the drill cuttings. The 

radionuclide activity would not likely be 

completely mixed with the drill cuttings. However, 

when considering the different possibilities, such 

as the cuttings from the tank being covered with 

cuttings from below the tank, it seems appropriate 

to make this gross assumption of complete mixing. 

Contaminated sand pads are located beneath 

Tanks WM-185 and WM-187. Equation (5-4) can 

be expanded to include this contamination as 

follows: 
5-6 

Vault 

Water well 

22-in. diameter 

2,200 m 

contaminated 

area 

2 

Figure 5-1. Graphical representation of the acute intruder-drilling scenario. 





















× +

π ρ

× = 2 2

12 1 10

sand pad

sand pad

Tank

Tank

b s,i r

Ci

r

Ci

D

C (5-5) 

where 

Cisand pad = activity in the sand pad (Ci) 

rsand pad = radius of the sand pad (m). 

In all cases, the depletion of the tank and sand 

pad radioactivities by leaching from water 

infiltration was not considered. The only 

mechanism for loss of activity in the tank and sand 

pad was by radioactive decay. 

The depth of the borehole (Db) was assumed to 

be 122 m (400 ft) to the aquifer (Rodriguez et al. 

1997). The bulk density (ρ) was assumed to be 

1.6E+06 g/m3

 (INEEL 1999b). When the volume 

of soil and waste excavated by the 22-in.-diameter 

well (πr

2

well × Db=30 m3

) is spread over 2,200 m2, 

the final contaminated zone is approximately 

0.014 m (0.54 in.) thick. 

The unit concentration dose factor (DFinh,i) for 

the inhalation of suspended cuttings was 

calculated as follows: 

inh i inh i inh inh DF = DCF × ASR × f × I , , (5-6) 

where 

DFinh,i = committed effective dose 

equivalent for inhalation of 

radionuclide i per unit soil 

concentration (mrem/yr per pCi/g) 

DCFinh,i = inhalation dose conversion factor 

for radionuclide i (mrem/pCi) 

ASR = soil mass loading in air (g/m3

) 
5-7 

finh = occupancy factor for inhalation 

pathway (dimensionless) 

Iinh = annual intake of air (m3

/yr). 

The committed effective dose equivalent 

(CEDE) from inhalation of suspended cuttings can 

be calculated for the radionuclide soil 

concentrations by multiplying them by the 

respective unit concentration dose factor (DFinh,i). 

The dose conversion factors used were taken 

from Federal Guidance Reports 11 and 12 (EPA 

1988, 1993). The most conservative values were 

used in the intruder analysis. This approach is 

considered reasonable since no information is 

available on the chemical and physical form of the 

waste constituents. The soil mass loading was 

assumed to be 1 mg/m3

 (EG&G Idaho 1984), 

representative of construction activities. The 

inhalation occupancy factor is the fraction of time 

of a year spent onsite (0.018 or 160 hr/yr) (Seitz 

1991). The annual intake of air was assumed to be 

8,400 m3

/yr. 

The unit concentration dose factor (DFext,i) for 

external radiation dose was calculated as follows: 

DFext,i = DCFext,i × f ext × ρ (5-7) 

where 

DFext,i = unit concentration dose factor for 

external exposure to radionuclide 

i per unit soil concentration 

(mrem/yr per pCi/g) 

DCFext,i = external dose conversion factor 

for radionuclide i (mrem/yr per 

pCi/m3

) 

fext = occupancy factor for external 

radiation (dimensionless) 

ρ = soil bulk density (g/m3). 

The EDE from external radiation exposure can 

be calculated for the radionuclide soil 

concentrations by multiplying them by the 

respective unit concentration dose factor (DFext,i). 

The occupancy factor for external radiation is the 

fraction of a year spent at the site (0.018 or 

160 hr/yr) (Seitz 1991). The soil density was 

assumed to be 1.6E+06 g/m3

 (Rodriguez et al. 

1997). External dose conversion factors for 

contaminated soil 1 cm (0.4 in.) thick with infinite 

lateral extent were used in the calculations from 

Federal Guidance Report 12 (EPA 1993). 

The unit concentration dose factor (DFing,i) for 

inadvertent soil ingestion was calculated as 

follows: 

ing ,i ing ,i soil soil DF = DCF × f × I (5-8) 

where 

DFing,i = unit concentration dose factor for 

inadvertent soil ingestion of 

radionuclide i per unit soil 

concentration (mrem/yr per pCi/g) 

DCFing,i = ingestion dose conversion factor 

for radionuclide i (mrem/pCi) 

fsoil = occupancy factor for inadvertent 

soil ingestion (dimensionless) 

Isoil = annual consumption rate of soil 

(g/yr). 

The CEDE from inadvertent soil ingestion can 

be calculated for the radionuclide soil 

concentrations by multiplying them by the 

respective unit concentration dose factor (DFing,i). 

The occupancy factor for inadvertent soil 

ingestion is the fraction of a year spent at the site 

(0.018 or 160 hr/yr) (Seitz 1991). The annual soil 

ingestion rate was assumed to be 40 g/yr 

(1.4 oz/yr) (Calabrese et al. 1990). 

The dose from inhalation of 3

H was calculated 

assuming HTO was the only form present. The 

CEDE from inhalation and dermal absorption of 

HTO was calculated from 

3 3 1 5 inh,H − = × inh × inh × inh × Ca,H − D . DCF I f (5-9) 
5-8 

where 

Dinh,H-3 = CEDE from inhalation and 

absorption of HTO (mrem/yr) 

1.5 = factor to adjust the dose factor 

(DCFinh) to account for 

absorption of 3

H through the 

skin 

Ca,H-3 = activity concentration of HTO 

in air (pCi/m3

). 

The concentration of HTO in air was 

calculated assuming that all of the HTO is in the 

soil pore water. The flux from the surface is the 

product of the soil pore water 3

H concentration 

and annual evapotranspiration. This flux is 

assumed to be mixed into a 2-m (7-ft) mixing 

zone. The concentration of HTO in air is given by 

H U

E A C 3.17 10 0.5

C

mix

t

v

8 s

a,H 3

×

× × × × ×

=

−

−

θ

ρ

(5-10) 

where 

3.17×10-8 = conversion factor (yr/s) 

0.5 = fraction of time wind is blowing 

toward receptor (dimensionless) 

Cs = concentration in the soil (pCi/g) 

θv = volumetric water content 

(dimensionless) 

Et = evapotranspiration rate (m/yr) 

A = area of the contaminated zone 

(m2

) 

Hmix = height of atmospheric mixing 

zone (m) 

U = mean wind speed (m/s). 

The volumetric water content of the well 

borehole cuttings was assumed to be the same as 

the near-surface alluvium, 0.28 (taken from the 

groundwater modeling results presented in 

Section 3). The evapotranspiration rate is the 

annual quantity of water evaporated from the 

surface, which was assigned a value of 0.15 m/yr 

(Maheras et al. 1994). The height of the 

atmospheric mixing zone and mean wind speed 

were conservatively assumed to be 2 m and 2 m/s, 

respectively. 

The dose from inhalation of 14C was 

calculated in a similar fashion. All airborne 14C 

was assumed to be 14CO2. The CEDE from 

inhalation of 14CO2 was calculated as follows: 

inh,C−14 = inh × inh × inh × Ca,C−14 D DCF I f (5-11) 

where 

Dinh,C-14 = CEDE from inhalation of 

14CO2 (mrem/yr) 

Ca,C-14 = activity concentration of 14CO2

in air (pCi/m3

). 

The concentration of 14CO2 in air was estimated 

from 

H U

. . EVSN A C

mix a,C ×

× × × × =

−

−

3 17 10 0 5 8

14 (5-12) 

where 

EVSN = the 14CO2 evasion rate 

(pCi/m2

/yr). 

The carbon evasion rate was estimated from 

s C ref EVSN = × × C × E × × d , −14 ρ 6 1 10 (5-13) 

where 

Cs,C-14 = concentration of 14C in 

contaminated soil (pCi/g) 

E = carbon evasion rate constant 

(per yr) 

dref = reference soil depth (m). 
5-9 

The carbon evasion rate constant was assumed 

to be 22/yr, which is the default value used in the 

RESRAD computer model (Yu et al. 1993). The 

soil layer releasing 14CO2 was assumed to be equal 

to the depth of the contaminated zone [i.e., 

0.014 m (0.54 in.)]. The same values for the area, 

mixing height, and wind speed used for 3

H also 

were used for 14C. 

The dose conversion factors from Federal 

Guidance Reports 11 and 12 (EPA 1988 and 1993) 

are listed in Table 5-1. The parameter values used 

in the acute intruder-drilling scenario are 

summarized in Table 5-2. The resulting unit 

concentration dose factors for inhalation, external 

exposure, and inadvertent soil ingestion for the 

acute drilling scenario are provided in Table 5-3. 

The results of the analyses are given in 

Section 5.6. 

5.3 Acute IntruderConstruction Scenario 

The acute intruder-construction scenario 

assumes that an inadvertent intruder moves onto 

the TFF and excavates a 20 × 10 × 3-m 

(70 × 30 × 10-ft) excavation for a basement 

(Oztunali and Roles 1986). 

The intruder is assumed to spend 160 hr 

working on the excavation of the basement. The 

exposure pathways for this acute scenario include 

inhalation of resuspended soil, external exposure 

to the ground surface, and inadvertent soil 

ingestion. 

The determination of the unit concentration 

dose factors for each exposure pathway were 

determined using the equations in Section 5.2 for 

the acute intruder-drilling scenario with the 

exception that the external dose factors for 

contaminated soil were based on infinite thickness 

and infinite lateral extent from Federal Guidance 

Report 12 (EPA 1993). 

The activity concentration of radionuclides in 

the excavation were determined by assuming that 

the radionuclide inventory in the 30% of the 

piping located within 3 m (10 ft) of the surface 

was located over the 5-acre (20,234-m2

) area of 

the TFF. The soil activity concentration in the 

20 × 10 × 3-m (70 × 30 × 10-ft) excavation was 

then determined as follows: 

TL A v s

p p L A

s i P x T x E x

C x M x P x E

C

ρ , = (5-14) 

where 

Cs,i = soil activity concentration of 

radionuclide i (pCi/g) 

Cp = piping activity concentration of 

radionuclide i at the time of 

intrusion (pCi/kg) 

Mp = mass of residual contamination in 

the piping (kg) 

PL = length of piping within 3 m of 

the surface (m) 

EA = area of the excavation (m) 

PTL = total piping length at TFF (m) 

TA = area of the TFF 

Ev = volume of excavation (m3) 

ρs = bulk density of the soil (g/m3). 

The piping activity concentrations (Cp) were based 

on the solid tank inventory provided in Section 2. 

The mass of residual contamination in the piping 

(Mp) was 15.5 kg (34.2 lb) based on the piping 

inventory analysis provided in Appendix A. The 

total length of process piping at the TFF located 

within 3 m (10 ft)of the surface (PL) is 975.4 m 

(3,200. ft). The area of the excavation (EA) is 

200 m2

 (20 × 10 m[70 × 30-ft]). The total length of 

the process piping at the TFF (PTL) is 3230.9 m 

(10,600. ft). The area of the TFF (TA) is 5 acres 

(20,234 m2

). The volume of the excavation (EV) is 

600 m3

 (20 × 10 × 3 m [70 × 30 × 10-ft]). The 

bulk density of the soil (ρs) was assumed to be 

1E+06 g/m3

. 
5-10

Table 5-1. Dose conversion factors used in the acute intruder-drilling scenario. 

Nuclide 

Inhalation DCF 

(mrem/pCi) 

External DCF 

1 cm 

(mrem/yr per pCi/g) 

External DCF 

15 cm 

(mrem/yr per pCi/g) 

Ingestion DCF 

(mrem/pCi) 

10Be 3.54E−04 3.62E−04 1.06E−03 4.66E−06 

14C 2.09E−06 8.03E−06 1.34E−05 2.09E−06 

113mCd 1.53E−03 2.24E−04 6.39E−04 1.61E−04 

60Co 2.19E−04 2.84E+00 1.35E+01 2.69E−05 

135Cs 4.55E−06 1.96E−05 3.83E−05 7.07E−06 

154Eu 2.86E−04 1.42E+00 6.58E+00 9.55E−06 

3

H 6.40E−08 0.00E+00 0.00E+00 6.40E−08 

166mHo 7.73E−04 2.04E+00 9.15E+00 8.07E−06 

129I 1.74E−04 1.11E−02 1.29E−02 2.76E−04 

115In 3.74E−03 1.47E−04 3.96E−04 1.58E−04 

138La 1.37E−03 1.39E+00 6.63E+00 5.88E−06 

94Nb 4.44E−04 1.84E+00 8.46E+00 7.14E−06 

63Ni 3.10E−06 0.00E+00 0.00E+00 5.77E−07 

107Pd 1.28E−05 0.00E+00 0.00E+00 1.49E−07 

146Pm 1.47E−04 8.80E−01 3.94E+00 3.67E−06 

87Rb 3.23E−06 6.16E−05 1.40E−04 4.92E−06 

79Se 9.84E−06 1.08E−05 1.86E−05 8.70E−06 

146Sm 8.25E−02 0.00E+00 0.00E+00 2.04E−04 

151Sm 3.00E−05 9.75E−07 9.84E−07 3.89E−07 

121mSn 1.15E−05 1.75E−03 1.96E−03 1.55E−06 

98Tc 2.29E−05 1.66E+00 7.58E+00 4.88E−06 

99Tc 8.33E−06 5.45E−05 1.25E−04 1.46E−06 

123Te 1.05E−05 6.26E−03 6.56E−03 4.18E−06 

Short Decay Chains 

108Ag 0.00E+00 2.34E−02 1.03E−01 0.00E+00 

108mAg 2.83E−04 1.91E+00 8.61E+00 7.62E−06 

137mBa 0.00E+00 7.02E−01 3.19E+00 0.00E+00 

137Cs 3.19E−05 2.50E−04 7.36E−04 5.00E−05 

152Eu 2.21E−04 1.31E+00 6.01E+00 6.48E−06 

152Gd 2.43E−01 0.00E+00 0.00E+00 1.61E−04 
Table 5-1. (continued). 

5-11 

Nuclide 

Inhalation DCF 

(mrem/pCi) 

External DCF 

1 cm 

(mrem/yr per pCi/g) 

External DCF 

15 cm 

(mrem/yr per pCi/g) 

Ingestion DCF 

(mrem/pCi) 

93mNb 2.92E−05 1.04E−04 1.04E−04 5.22E−07 

147Pm 3.92E−05 2.22E−05 4.99E−05 1.05E−06 

126Sb 1.17E−05 3.34E+00 1.52E+01 1.07E−05 

126mSb 3.39E−08 1.83E+00 8.29E+00 9.36E−08 

147Sm 7.47E−02 0.00E+00 0.00E+00 1.85E−04 

126Sn 9.95E−05 5.32E−02 1.48E−01 1.95E−05 

90Sr 1.30E−03 2.45E−04 6.95E−04 1.42E−04 

90Y 8.44E−06 5.79E−03 2.24E−02 1.08E−05 

93Zr 3.21E−04 0.00E+00 0.00E+00 1.66E−06 

Decay Chain 1 

228Ac 3.08E−04 1.12E+00 5.16E+00 2.16E−06 

212Bi 2.16E−05 2.15E−01 1.00E+00 1.06E−06 

252Cf 1.57E−01 1.30E−04 1.76E−04 1.08E−03 

244Cm 2.48E−01 1.22E−04 1.26E−04 2.02E−03 

248Cm 1.65E+00 8.44E−05 8.78E−05 1.36E−02 

240Np 8.14E−08 1.51E+00 6.82E+00 2.37E−07 

240mNp 0.00E+00 3.89E−01 1.77E+00 0.00E+00 

212Pb 1.69E−04 1.70E−01 6.76E−01 4.55E−05 

212Po 0.00E+00 0.00E+00 0.00E+00 0.00E+00 

216Po 0.00E+00 1.98E−05 9.10E−05 0.00E+00 

236Pu 1.45E−01 1.56E−04 2.24E−04 1.17E−03 

240Pu 4.29E−01 1.16E−04 1.46E−04 3.54E−03 

244Pu 4.03E−01 7.19E−05 7.55E−05 3.32E−03 

224Ra 3.16E−03 1.16E−02 4.89E−02 3.66E−04 

228Ra 4.77E−03 0.00E+00 0.00E+00 1.44E−03 

220Rn 0.00E+00 4.58E−04 2.05E−03 0.00E+00 

228Th 3.42E−01 2.28E−03 7.79E−03 3.96E−04 

232Th 1.64E+00 2.17E−04 5.19E−04 2.73E−03 

208Tl 0.00E+00 3.66E+00 1.81E+01 0.00E+00 

232U 1.49E−02 3.51E−04 8.91E−04 1.31E−03 
Table 5-1. (continued). 

5-12 

Nuclide 

Inhalation DCF 

(mrem/pCi) 

External DCF 

1 cm 

(mrem/yr per pCi/g) 

External DCF 

15 cm 

(mrem/yr per pCi/g) 

Ingestion DCF 

(mrem/pCi) 

236U 1.25E−01 1.22E−04 2.13E−04 2.69E−04 

240U 2.27E−06 1.05E−03 1.42E−03 4.44E−06 

Decay Chain 2 

225Ac 1.08E−02 1.79E−02 6.24E−02 1.11E−04 

241Am 4.44E−01 2.15E−02 4.37E−02 3.64E−03 

217At 0.00E+00 3.64E−04 1.61E−03 0.00E+00 

213Bi 1.71E−05 1.59E−01 7.01E−01 7.22E−07 

249Cf 5.77E−01 3.94E−01 1.72E+00 4.74E−03 

245Cm 4.55E−01 9.79E−02 3.36E−01 3.74E−03 

221Fr 0.00E+00 3.61E−02 1.48E−01 0.00E+00 

237Np 5.40E−01 2.58E−02 7.77E−02 4.44E−03 

233Pa 9.55E−06 2.32E−01 9.64E−01 3.63E−06 

209Pb 9.47E−08 2.62E−04 7.62E−04 2.13E−07 

213Po 0.00E+00 0.00E+00 0.00E+00 0.00E+00 

241Pu 8.25E−03 1.79E−06 5.88E−06 6.85E−05 

225Ra 7.77E−03 7.92E−03 1.10E−02 3.85E−04 

229Th 2.15E+00 9.51E−02 3.18E−01 3.53E−03 

209Tl 0.00E+00 2.30E+00 1.08E+01 0.00E+00 

233U 1.35E−01 4.03E−04 1.35E−03 2.89E−04 

237U 3.53E−06 1.49E−01 5.19E−01 3.17E−06 

Decay Chain 3 

227Ac 6.70E+00 1.44E−04 4.89E−04 1.41E−02 

243Am 4.40E−01 5.53E−02 1.42E−01 3.62E−03 

211Bi 0.00E+00 5.53E−02 2.39E−01 0.00E+00 

251Cf 5.88E−01 1.38E−01 5.16E−01 4.85E−03 

243Cm 3.07E−01 1.46E−01 5.64E−01 2.51E−03 

247Cm 4.14E−01 3.74E−01 1.64E+00 3.42E−03 

223Fr 7.84E−03 5.79E−02 1.89E−01 8.62E−06 

239Np 2.51E−06 1.91E−01 7.29E−01 3.26E−06 

231Pa 1.28E+00 4.30E−02 1.80E−01 1.06E−02 
Table 5-1. (continued). 

5-13 

Nuclide 

Inhalation DCF 

(mrem/pCi) 

External DCF 

1 cm 

(mrem/yr per pCi/g) 

External DCF 

15 cm 

(mrem/yr per pCi/g) 

Ingestion DCF 

(mrem/pCi) 

211Pb 7.53E−06 6.07E−02 2.73E−01 5.25E−07 

211Po 0.00E+00 9.13E−03 4.18E−02 0.00E+00 

215Po 0.00E+00 2.11E−04 9.30E−04 0.00E+00 

239Pu 4.29E−01 1.05E−04 2.84E−04 3.54E−03 

243Pu 1.64E−07 2.58E−02 7.85E−02 3.34E−07 

223Ra 7.84E−03 1.51E−01 5.79E−01 6.59E−04 

219Rn 0.00E+00 6.65E−02 2.88E−01 0.00E+00 

227Th 1.62E−02 1.21E−01 4.95E−01 3.81E−05 

231Th 8.77E−07 1.32E−02 3.62E−02 1.35E−06 

207Tl 0.00E+00 4.22E−03 1.77E−02 0.00E+00 

235U 1.23E−01 1.77E−01 7.06E−01 2.66E−04 

Decay Chain 4 

242Am 5.85E−05 1.53E−02 4.99E−02 1.41E−06 

242mAm 4.26E−01 8.01E−04 1.68E−03 3.52E−03 

210Bi 1.96E−04 1.03E−03 3.47E−03 6.40E−06 

214Bi 6.59E−06 1.71E+00 8.14E+00 2.83E−07 

250Cf 2.62E−01 1.16E−04 1.18E−04 2.13E−03 

242Cm 1.73E−02 1.42E−04 1.69E−04 1.15E−04 

246Cm 4.51E−01 1.11E−04 1.16E−04 3.70E−03 

238Np 3.70E−05 6.31E−01 2.95E+00 4.00E−06 

234Pa 8.14E−07 2.20E+00 1.00E+01 2.16E−06 

234mPa 0.00E+00 1.78E−02 7.85E−02 0.00E+00 

210Pb 1.36E−02 1.54E−03 2.45E−03 5.37E−03 

214Pb 7.81E−06 2.93E−01 1.25E+00 6.25E−07 

210Po 9.40E−03 9.94E−06 4.58E−05 1.90E−03 

214Po 0.00E+00 9.75E−05 4.48E−04 0.00E+00 

218Po 0.00E+00 1.06E−05 4.91E−05 0.00E+00 

238Pu 3.92E−01 1.18E−04 1.51E−04 3.20E−03 

242Pu 4.11E−01 9.77E−05 1.28E−04 3.36E−03 

226Ra 8.58E−03 7.75E−03 3.08E−02 1.32E−03 
Table 5-1. (continued). 

5-14 

Nuclide 

Inhalation DCF 

(mrem/pCi) 

External DCF 

1 cm 

(mrem/yr per pCi/g) 

External DCF 

15 cm 

(mrem/yr per pCi/g) 

Ingestion DCF 

(mrem/pCi) 

222Rn 0.00E+00 4.74E−04 2.13E−03 0.00E+00 

230Th 3.26E−01 4.35E−04 1.19E−03 5.48E−04 

234Th 3.50E−05 8.52E−03 2.41E−02 1.37E−05 

234U 1.32E−01 1.89E−04 4.00E−04 2.83E−04 

238U 1.18E−01 8.26E−05 1.03E−04 2.55E−04 

Table 5-2. Parameter values used in the acute intruder-drilling scenario. 

Parameter Description Value 

A Area of contaminated zone (m2

) 2,200 

ASR Soil mass loading in air (g/m3

) 1E−03 

Db Total depth of well borehole (m) 122 

dref Reference soil depth (m) 0.014 

E Carbon evasion rate constant (per yr) 22 

Et Evapotranspiration rate (m/yr) 0.15 

fext Occupancy factor for external exposure (unitless) 0.018 

finh Occupancy factor for inhalation (unitless) 0.018 

fsoil Occupancy factor for soil ingestion (unitless) 0.018 

Hmix Height of mixing zone (m) 2 

Iinh Annual intake of air (m3

/yr) 8,400 

Isoil Annual consumption rate of soil (g/yr) 40 

rsand pad Radius of the sand pad (m) 7.8 

rTank Radius of the tank (m) 7.6 

rwell Radius of the well (m) 0.28 

ρ Bulk density of the well cuttings (g/m3

) 1.6E+06 

θv Volumetric water content (unitless) 0.28 

tw Thickness of waste zone (m) 0.15 

U Mean wind speed (m/s) 2 
5-15

Table 5-3. Unit concentration dose factors for acute intruder-drilling scenario. 

Unit DF Inhalation Unit DF External Unit DF Ingestion Unit DF Total 

Nuclide (mrem/yr per pCi/g) 

10Be 5.44E−05 6.62E−06 3.41E−06 6.44E−05 

14C 3.21E−04 1.47E−07 1.53E−06 3.22E−04 

113mCd 2.34E−04 4.09E−06 1.18E−04 3.56E−04 

60Co 3.35E−05 5.19E−02 1.97E−05 5.19E−02 

135Cs 6.98E−07 3.58E−07 5.16E−06 6.22E−06 

154Eu 4.39E−05 2.60E−02 6.97E−06 2.60E−02 

3

H 1.47E−05 0.00E+00 4.68E−08 1.48E−05 

166mHo 1.19E−04 3.72E−02 5.89E−06 3.73E−02 

129I 2.66E−05 2.03E−04 2.02E−04 4.31E−04 

115In 5.73E−04 2.69E−06 1.15E−04 6.91E−04 

138La 2.10E−04 2.55E−02 4.30E−06 2.57E−02 

94Nb 6.81E−05 3.36E−02 5.22E−06 3.37E−02 

63Ni 4.76E−07 0.00E+00 4.22E−07 8.98E−07 

107Pd 1.96E−06 0.00E+00 1.09E−07 2.07E−06 

146Pm 2.25E−05 1.61E−02 2.68E−06 1.61E−02 

87Rb 4.96E−07 1.13E−06 3.60E−06 5.22E−06 

79Se 1.51E−06 1.97E−07 6.35E−06 8.06E−06 

146Sm 1.27E−02 0.00E+00 1.49E−04 1.28E−02 

151Sm 4.60E−06 1.78E−08 2.84E−07 4.90E−06 

121mSn 1.77E−06 3.20E−05 1.13E−06 3.49E−05 

98Tc 3.51E−06 3.04E−02 3.57E−06 3.04E−02 

99Tc 1.28E−06 9.96E−07 1.07E−06 3.34E−06 

123Te 1.62E−06 1.14E−04 3.05E−06 1.19E−04 

Short Decay Chains 

108Ag 0.00E+00 4.26E−04 0.00E+00 4.26E−04 

108mAg 4.35E−05 3.48E−02 5.57E−06 3.49E−02 

137mBa 0.00E+00 1.28E−02 0.00E+00 1.28E−02 

137Cs 4.90E−06 4.57E−06 3.65E−05 4.60E−05 

152Eu 3.39E−05 2.40E−02 4.73E−06 2.40E−02 

152Gd 3.74E−02 0.00E+00 1.17E−04 3.75E−02 

93mNb 4.48E−06 1.91E−06 3.81E−07 6.77E−06 

147Pm 6.02E−06 4.06E−07 7.65E−07 7.19E−06 
Table 5-3. (continued). 

5-16 

Unit DF Inhalation Unit DF External Unit DF Ingestion Unit DF Total 

Nuclide (mrem/yr per pCi/g) 

126Sb 1.80E−06 6.11E−02 7.81E−06 6.11E−02 

126mSb 5.21E−09 3.34E−02 6.84E−08 3.34E−02 

147Sm 1.15E−02 0.00E+00 1.35E−04 1.16E−02 

126Sn 1.53E−05 9.72E−04 1.42E−05 1.00E−03 

90Sr 1.99E−04 4.47E−06 1.04E−04 3.08E−04 

90Y 1.29E−06 1.06E−04 7.87E−06 1.15E−04 

93Zr 4.92E−05 0.00E+00 1.21E−06 5.04E−05 

Decay Chain 1 

228Ac 4.73E−05 2.04E−02 1.58E−06 2.04E−02 

212Bi 3.31E−06 3.92E−03 7.76E−07 3.93E−03 

252Cf 2.41E−02 2.37E−06 7.92E−04 2.49E−02 

244Cm 3.80E−02 2.23E−06 1.47E−03 3.95E−02 

248Cm 2.54E−01 1.54E−06 9.95E−03 2.64E−01 

240Np 1.25E−08 2.76E−02 1.73E−07 2.76E−02 

240mNp 0.00E+00 7.10E−03 0.00E+00 7.10E−03 

212Pb 2.59E−05 3.11E−03 3.32E−05 3.17E−03 

212Po 0.00E+00 0.00E+00 0.00E+00 0.00E+00 

216Po 0.00E+00 3.62E−07 0.00E+00 3.62E−07 

236Pu 2.22E−02 2.85E−06 8.52E−04 2.31E−02 

240Pu 6.59E−02 2.12E−06 2.58E−03 6.84E−02 

244Pu 6.19E−02 1.31E−06 2.42E−03 6.43E−02 

224Ra 4.84E−04 2.12E−04 2.67E−04 9.64E−04 

228Ra 7.32E−04 0.00E+00 1.05E−03 1.78E−03 

220Rn 0.00E+00 8.36E−06 0.00E+00 8.36E−06 

228Th 5.24E−02 4.16E−05 2.89E−04 5.27E−02 

232Th 2.51E−01 3.96E−06 1.99E−03 2.53E−01 

208Tl 0.00E+00 6.69E−02 0.00E+00 6.69E−02 

232U 2.28E−03 6.41E−06 9.57E−04 3.25E−03 

236U 1.92E−02 2.23E−06 1.96E−04 1.94E−02 

240U 3.48E−07 1.92E−05 3.24E−06 2.28E−05 

Decay Chain 2 

225Ac 1.66E−03 3.26E−04 8.11E−05 2.06E−03 

241Am 6.81E−02 3.92E−04 2.66E−03 7.12E−02 
Table 5-3. (continued). 

5-17 

Unit DF Inhalation Unit DF External Unit DF Ingestion Unit DF Total 

Nuclide (mrem/yr per pCi/g) 

217At 0.00E+00 6.65E−06 0.00E+00 6.65E−06 

213Bi 2.63E−06 2.91E−03 5.27E−07 2.91E−03 

249Cf 8.86E−02 7.20E−03 3.46E−03 9.92E−02 

245Cm 6.98E−02 1.79E−03 2.73E−03 7.43E−02 

221Fr 0.00E+00 6.58E−04 0.00E+00 6.58E−04 

237Np 8.29E−02 4.71E−04 3.24E−03 8.66E−02 

233Pa 1.46E−06 4.23E−03 2.65E−06 4.23E−03 

209Pb 1.45E−08 4.78E−06 1.55E−07 4.95E−06 

213Po 0.00E+00 0.00E+00 0.00E+00 0.00E+00 

241Pu 1.27E−03 3.28E−08 5.00E−05 1.32E−03 

225Ra 1.19E−03 1.45E−04 2.81E−04 1.62E−03 

229Th 3.29E−01 1.74E−03 2.58E−03 3.34E−01 

209Tl 0.00E+00 4.20E−02 0.00E+00 4.20E−02 

233U 2.08E−02 7.37E−06 2.11E−04 2.10E−02 

237U 5.42E−07 2.72E−03 2.32E−06 2.72E−03 

Decay Chain 3 

227Ac 1.03E+00 2.63E−06 1.03E−02 1.04E+00 

243Am 6.76E−02 1.01E−03 2.65E−03 7.12E−02 

211Bi 0.00E+00 1.01E−03 0.00E+00 1.01E−03 

251Cf 9.03E−02 2.52E−03 3.54E−03 9.63E−02 

243Cm 4.71E−02 2.66E−03 1.84E−03 5.16E−02 

247Cm 6.36E−02 6.82E−03 2.50E−03 7.29E−02 

223Fr 1.20E−03 1.06E−03 6.30E−06 2.27E−03 

239Np 3.85E−07 3.48E−03 2.38E−06 3.48E−03 

231Pa 1.97E−01 7.85E−04 7.73E−03 2.05E−01 

211Pb 1.16E−06 1.11E−03 3.84E−07 1.11E−03 

211Po 0.00E+00 1.67E−04 0.00E+00 1.67E−04 

215Po 0.00E+00 3.86E−06 0.00E+00 3.86E−06 

239Pu 6.59E−02 1.91E−06 2.58E−03 6.84E−02 

243Pu 2.52E−08 4.71E−04 2.44E−07 4.71E−04 

223Ra 1.20E−03 2.76E−03 4.81E−04 4.45E−03 

219Rn 0.00E+00 1.21E−03 0.00E+00 1.21E−03 

227Th 2.48E−03 2.21E−03 2.78E−05 4.72E−03 
Table 5-3. (continued). 

5-18 

Unit DF Inhalation Unit DF External Unit DF Ingestion Unit DF Total 

Nuclide (mrem/yr per pCi/g) 

231Th 1.35E−07 2.40E−04 9.87E−07 2.41E−04 

207Tl 0.00E+00 7.71E−05 0.00E+00 7.71E−05 

235U 1.88E−02 3.24E−03 1.94E−04 2.23E−02 

Decay Chain 4 

242Am 8.97E−06 2.79E−04 1.03E−06 2.89E−04 

242mAm 6.53E−02 1.46E−05 2.57E−03 6.79E−02 

210Bi 3.00E−05 1.89E−05 4.68E−06 5.36E−05 

214Bi 1.01E−06 3.12E−02 2.07E−07 3.12E−02 

250Cf 4.02E−02 2.12E−06 1.56E−03 4.18E−02 

242Cm 2.65E−03 2.59E−06 8.38E−05 2.74E−03 

246Cm 6.93E−02 2.03E−06 2.70E−03 7.20E−02 

238Np 5.68E−06 1.15E−02 2.92E−06 1.15E−02 

234Pa 1.25E−07 4.03E−02 1.58E−06 4.03E−02 

234mPa 0.00E+00 3.25E−04 0.00E+00 3.25E−04 

210Pb 2.08E−03 2.82E−05 3.92E−03 6.03E−03 

214Pb 1.20E−06 5.36E−03 4.57E−07 5.36E−03 

210Po 1.44E−03 1.82E−07 1.39E−03 2.83E−03 

214Po 0.00E+00 1.78E−06 0.00E+00 1.78E−06 

218Po 0.00E+00 1.94E−07 0.00E+00 1.94E−07 

238Pu 6.02E−02 2.16E−06 2.34E−03 6.25E−02 

242Pu 6.30E−02 1.78E−06 2.45E−03 6.55E−02 

226Ra 1.32E−03 1.42E−04 9.68E−04 2.43E−03 

222Rn 0.00E+00 8.67E−06 0.00E+00 8.67E−06 

230Th 5.00E−02 7.95E−06 4.00E−04 5.04E−02 

234Th 5.38E−06 1.56E−04 9.97E−06 1.71E−04 

234U 2.03E−02 3.45E−06 2.07E−04 2.05E−02 

238U 1.82E−02 1.51E−06 1.86E−04 1.84E−02 
5-19 

The results of the intruder-construction 

analyses are given in Section 5.6. 

5.4 Chronic Intruder PostDrilling Scenario Definition 

The chronic post-drilling scenario assumes 

that an inadvertent intruder moves onto the TFF 

and drills a residential well into the waste. The 

drilling portion of the scenario evaluates a 6-in. 

residential well. This type of well serves a single 

residence and provides sufficient water for a 

family garden and small quantities of livestock. As 

described in the acute drilling scenario, largediameter wells are drilled to serve irrigation 

systems (i.e., hundreds of acres) that are located in 

the middle of farm fields, not near a farmer’s 

residence. Therefore, in the chronic post-drilling 

scenario, the residence/home garden is evaluated 

using the traditional drinking water well diameter 

of 6 in. 

The drill cuttings that are brought to the 

surface are assumed to be spread over 2,200 m2

(24,000 ft or approximately one-half acre) of land 

surface. The waste is assumed to be mixed to a 

depth of 0.61 m (24 in.). The mixing depth of 

0.61 m is based on using a deep tilling plow to 

increase the depth of the root zone and to break up 

soil compaction. To minimize erosion these plows 

are used in areas of southeast Idaho with highly 

erodible soils. Deep tilling plows have shanks that 

till to a depth of 24 in. and are sold at Idaho Falls 

implement dealers (Maheras et al. 1997). 

The chronic post-drilling scenario assumes 

that the intruder is exposed to the drill cuttings 

during plowing and cultivation (i.e., dust 

inhalation). In addition, the intruder is assumed to 

ingest contaminated food products from the garden 

and from beef and milk cattle consuming 

contaminated forage. The intake of contaminated 

forage by cattle was adjusted according to the 

fraction of feed grown on contaminated cuttings 

and the necessary remaining feed obtained from 

uncontaminated ground. Figure 5-2 illustrates the 

chronic intruder post-drilling scenario. 

The activity concentration of radionuclides in 

the contaminated zone is calculated as follows: 

ρ

π

c c

well w

s i w i A D

r t C C

2

,

12

, = 1×10 × × (5-15) 

Figure 5-2. Chronic intruder post-drilling scenario.
5-20 

where 

Cs,i = soil activity concentration of 

radionuclide i (pCi/g) 

1 × 1012 = factor for converting Ci to pCi 

Cw,i = waste activity concentration of 

radionuclide i at the time of 

intrusion (Ci/m3

) 

rwell = radius of the well borehole (m) 

tw = thickness of the waste zone 

(m) 

Ac = area over which contamination 

is spread (m2

) 

Dc = final depth of the contaminated 

zone after tilling (m) 

ρ = bulk density of the soil (g/m3) 

since 

2

, 2 and c cont

Tank w

Tank

w i A r

r t

Ci C π

π = = (5-16) 

where 

CiTank = activity in the tank (Ci) 

rTank = radius of the tank (m) 

rcont = radius of the contaminated 

zone (m). 

Then by substitution 

π ρ

π

π cont c

well w

w

s i

r D

r t

r t

Ci C 2

2

2

Tank

12 Tank

, = 1×10 × × (5-17) 

Simplifying results in 

π cont c ρ

well

s i

r r D

Ci r C 2 2

Tank

2

12 Tank

, = 1×10 × (5-18) 

The equation can be expanded to include the 

contribution from sand pad contamination as 

follows: 

















π ρ

+

π ρ

= × ×

sand pad cont c

sand pad sand pad

Tank cont c

Tank well

s,i

r r D

Ci r

r r D

Ci r

C

2 2

2

2 2

2

12 1 10

 (5-19) 

where 

Cisand pad = activity in the sand pad (Ci) 

rsand pad = radius of the sand pad (m). 

The unit concentration dose factor (DFing,soil,i) 

for inadvertent soil ingestion was calculated using 

Equation (5-8). The annual soil ingestion rate from 

outdoor air was assumed to be 10 mg/d or 

3.65 g/yr (Konz et al. 1989). 

The intruder was assumed to be indoors 50% 

of the time, outdoors 25% of the time, and offsite 

for 25% of the time, resulting in a soil ingestion 

occupancy factor of 0.75. 

The unit concentration dose factor (DFinh,i) for 

the inhalation of suspended cuttings was 

calculated as follows: 

[ ] ( ) inh,outdoor inh,indoor inh inh

inh,i inh ,i

f f S I

DF DCF ASR

+ × ×

= × × (5-20) 

where 

DFinh,i = unit concentration dose factor for 

inhalation of radionuclide i per 

unit soil concentration (mrem/yr 

per pCi/g) 

DCFinh,i = inhalation dose conversion factor 

for radionuclide i (mrem/pCi) 

ASR = soil mass loading in air (g/m3) 

finh,outdoor = occupancy factor for inhalation 

pathway outside the home 

(dimensionless) 
5-21 

finh,indoor = occupancy factor for inhalation 

pathway inside the home 

(dimensionless) 

Sinh = shielding factor for dust inside the 

home (dimensionless) 

Iinh = annual intake of air (m3/yr). 

The CEDE from inhalation of suspended 

cuttings can be calculated for the actual 

radionuclide soil concentrations by multiplying the 

actual radionuclide soil concentrations by the 

respective unit concentration dose factor (DFinh,i). 

The inhalation occupancy factor assumes that 

indoor dust loading is 40% of outdoor levels. This 

results in an effective inhalation occupancy factor 

of 0.45 [i.e., indoor time fraction (0.5) × (0.4) + 

outdoor time fraction (0.25)]. 

The unit concentration dose factor (DFext,i) for 

external radiation dose was calculated as follows: 

[ ] ( ) × + × ρ

= ×

ext,indoor ext,indoor ext,outdoor

ext ,i ext,i

f S f

DF DCF

 (5-21) 

where 

DFext,i = unit concentration dose factor 

for external exposure to 

radionuclide i per unit soil 

concentration (mrem/yr per 

pCi/g) 

DCFext,i = external dose conversion factor 

for radionuclide i (mrem/yr per 

pCi/m3

) 

fext,indoor = occupancy factor for external 

radiation while inside the home 

(dimensionless) 

Sext,indoor = shielding factor for indoor 

exposures (dimensionless) 

fext,outdoor = occupancy factor for external 

radiation while outside the 

home (dimensionless) 

ρ = soil bulk density (g/m3). 

The EDE from external radiation exposure can 

be calculated for the radionuclide soil 

concentrations by multiplying them by the 

respective unit concentration dose factor (DFext,i). 

The external exposure occupancy factor in this 

case includes a shielding factor for the residence 

of 0.7 for all radionuclides except 137Cs/137mBa, 

which was assigned an external exposure 

occupancy factor of 0.35. This was based on a 

separate analysis of the gamma exposure and 

attenuation by the foundation of the house. The 

analysis supporting this factor is provided in 

Appendix K. 

Ingestion doses were calculated for ingestion 

of contaminated vegetables, beef, and dairy 

products. Since the area of contamination is small 

(i.e., 2,200 m2

 [24,000 ft2]) and cannot fully 

support beef cattle or milk cows, the consumption 

rate of the contaminated pasture was adjusted to 

reflect the maximum amount of feed that could be 

produced on the lot. The adjustment factors 

provided in the RWMC PA (Maheras et al. 1997) 

were used in this analysis. The RWMC PA

determined that 9% of the total pasture eaten by 

the cows would be contaminated. The 

consumption rate for contaminated pasture was 

calculated by Maheras et al. (1997) to be 5.4 kg/d 

for beef cattle and 7.2 kg/d for milk cows. In 

addition, 25% of the intruder’s diet from 

vegetables was assumed to be contaminated. 

The human consumption diet developed by 

Yang and Nelson (1984, 1986), based on a 1977–

1978 United States Department of Agriculture 

survey (USDA 1983), was used in the assessment. 

The diet consists of 94 kg/yr of produce, 17 kg/yr 

of leafy vegetable, 89 L/yr of milk, and 55 kg/yr of 

beef products. 

The radionuclide intake from vegetables and 

resulting unit concentration dose factors (DFveg,i,k) 

assumes plant uptake by roots and foliar 

deposition. 

∑ × × [ ] + × ×

=

k veg,k veg,k iv,i foliar k

veg,i,k ing,i

f I B ( FA ASR FAR )

DF DCF (5-22) 
5-22 

where 

DFveg,i,k = unit concentration dose factor for 

radionuclide i from plant ingestion 

of vegetation type k for a unit soil 

concentration (mrem/yr per pCi/g) 

DCFing,i = ingestion CEDE conversion factor 

for radionuclide i (mrem/pCi) 

fveg,k = fraction of diet for vegetation type 

k derived from contaminated zone 

(dimensionless) 

Iveg,k = annual plant intake of vegetation 

type k (g/yr) 

Biv,i = plant-soil concentration ratio for 

radionuclide i (pCi/g plant per 

pCi/g soil) 

FA = area factor (dimensionless) 

ASRfoliar = mass loading factor (g/m3) 

FARk = foliar deposition factor for 

vegetation type k (m3

/g). 

Two vegetation types (k) were assumed: leafy 

vegetables and nonleafy vegetables such as fruits 

and grains. The fraction of the intruder’s diet 

derived from the contaminated zone was assumed 

to be 0.25 for each of these vegetation types. The 

plant-soil concentration ratios used for the 

assessment are provided in Table 5-4. The area 

factor was 0.94. The foliar deposition factor was 

0.054 m3

/g for fruits and nonleafy vegetables, 

0.26 m3

/g for leafy vegetables, and 0.28 m3/g for 

fodder (Maheras et al. 1997). 

The unit concentration dose factor (DFmilk,i) 

received through consumption of milk was 

calculated as follows: 

{ } [ ] k cow,soil ,i milk , fodder iv

milk ,i ing ,i milk milk milk

I B ( FA ASR FAR ) I

DF DCF f F I

+ × × +

= × × × ×

 (5-23) 

where 

DFmilk,i = unit concentration dose factor for 

radionuclide i from ingestion of 

milk for a unit soil concentration 

(mrem/yr per pCi/g) 

fmilk = fraction of milk derived from the 

contaminated site (dimensionless) 

Fmilk = milk transfer coefficient (d/L) 

Imilk = annual intake of milk (L/yr) 

Imilk, fodder = intake of fodder by livestock 

(kg/d) 

Icow,soil = intake of soil by livestock (kg/d). 

The fodder-to-milk transfer coefficients used 

for this assessment are provided in Table 5-4. The 

intake of fodder by livestock was 7.2 kg/d, 

assuming 9% of the pasture would be 

contaminated (Maheras et al. 1997). 

The unit concentration dose factor (DFbeef,i) 

received through consumption of beef products 

was determined as follows: 

{ } [ ] k cow,soil ,i beef , fodder iv

beef ,i ing ,i beef beef beef

I B ( FA ASR FAR ) I

DF DCF f F I

+ × × +

= × × × ×

 (5-24) 

where 

DFbeef,i = unit concentration dose factor for 

radionuclide i from ingestion of 

beef for a unit soil concentration 

(mrem/yr per pCi/g) 

fbeef = fraction of beef derived from 

contaminated site (dimensionless) 

Fbeef = beef transfer coefficient (d/kg) 

Ibeef = annual intake of beef (g/yr) 

Ibeef,fodder = intake of fodder by livestock 

(kg/d) 

Icow,soil = intake of soil by livestock (kg/d). 

The beef transfer coefficients used for this 

assessment are provided in Table 5-4. The intake 

of fodder by beef cows was 5.4 kg/d, assuming 9% 
5-23

Table 5-4. Plant, beef, and milk uptake factors used in the chronic post-drilling scenario. 

Nuclide 

Vegetable/soil 

(pCi/g plant 

per pCi/g soil) 

Beef 

(d/kg) 

Milk 

(d/L) 

10Be 4.00E−03 1.00E−03 2.00E−06 

14C 5.5 3.10E−02 1.20E−02 

113mCd 3.00E−01 4.00E−04 1.00E−03 

60Co 8.00E−02 2.00E−02 2.00E−03 

135Cs 4.00E−02 3.00E−02 8.00E−03 

154Eu 2.50E−03 2.00E−03 2.00E−05 

3

H 4.8 1.20E−02 1.00E−02 

166mHo 2.60E−03 2.00E−03 2.00E−05 

129I 2.00E−02 7.00E−03 1.00E−02 

115In 3.00E−03 4.00E−03 2.00E−04 

138La 2.50E−03 2.00E−03 2.00E−05 

94Nb 1.00E−02 3.00E−07 2.00E−06 

63Ni 5.00E−02 5.00E−03 2.00E−02 

107Pd 1.00E−01 1.00E−03 5.00E−03 

146Pm 2.50E−03 2.00E−03 2.00E−05 

87Rb 1.30E−01 1.50E−02 1.00E−02 

79Se 1.00E−01 1.00E−01 1.00E−02 

146Sm 2.50E−03 2.00E−03 2.00E−05 

151Sm 2.50E−03 2.00E−03 2.00E−05 

121mSn 2.50E−03 1.00E−02 1.00E−03 

98Tc 5.00E+00 1.00E−04 1.00E−03 

99Tc 5.00E+00 1.00E−04 1.00E−03 

123Te 6.00E−01 7.00E−03 5.00E−04 

Short Decay Chains 

108Ag 1.50E−01 3.00E−03 2.50E−02 

108mAg 1.50E−01 3.00E−03 2.50E−02 

137mBa 4.00E−02 3.00E−02 8.00E−03 

137Cs 4.00E−02 3.00E−02 8.00E−03 

152Eu 2.50E−03 2.00E−03 2.00E−05 

152Gd 2.50E−03 2.00E−03 2.00E−05 

93mNb 1.00E−02 3.00E−07 2.00E−06 

147Pm 2.50E−03 2.00E−03 2.00E−05 
Table 5-4. (continued). 

5-24 

Nuclide 

Vegetable/soil 

(pCi/g plant 

per pCi/g soil) 

Beef 

(d/kg) 

Milk 

(d/L) 

126Sb 1.00E−02 1.00E−03 1.00E−04 

126mSb 1.00E−02 1.00E−03 1.00E−04 

147Sm 2.50E−03 2.00E−03 2.00E−05 

126Sn 2.50E−03 1.00E−02 1.00E−03 

90Sr 3.00E−01 8.00E−03 2.00E−03 

90Y 3.00E−01 8.00E−03 2.00E−03 

93Zr 1.00E−03 1.00E−06 6.00E−07 

Decay Chain 1 

228Ac 2.50E−03 2.00E−05 2.00E−05 

212Bi 1.00E−01 2.00E−03 5.00E−04 

252Cf 1.00E−03 6.00E−05 7.50E−07 

244Cm 1.00E−03 2.00E−05 2.00E−06 

248Cm 1.00E−03 2.00E−05 2.00E−06 

240Np 2.00E−02 1.00E−03 5.00E−06 

240mNp 2.00E−02 1.00E−03 5.00E−06 

212Pb 1.00E−02 8.00E−04 3.00E−04 

212Po 1.00E−03 5.00E−03 3.40E−04 

216Po 1.00E−03 5.00E−03 3.40E−04 

236Pu 1.00E−03 1.00E−04 1.00E−06 

240Pu 1.00E−03 1.00E−04 1.00E−06 

244Pu 1.00E−03 1.00E−04 1.00E−06 

224Ra 4.00E−02 1.00E−03 1.00E−03 

228Ra 4.00E−02 1.00E−03 1.00E−03 

220Rn 0.00E+00 0.00E+00 0.00E+00 

228Th 1.00E−03 1.00E−04 5.00E−06 

232Th 1.00E−03 1.00E−04 5.00E−06 

208Tl 2.00E−01 2.00E−03 3.00E−03 

232U 2.50E−03 3.40E−04 6.00E−04 

236U 2.50E−03 3.40E−04 6.00E−04 

240U 2.50E−03 3.40E−04 6.00E−04 

Decay Chain 2 

225Ac 2.50E−03 2.00E−05 2.00E−05 
Table 5-4. (continued). 

5-25 

Nuclide 

Vegetable/soil 

(pCi/g plant 

per pCi/g soil) 

Beef 

(d/kg) 

Milk 

(d/L) 

241Am 1.00E−03 5.00E−05 2.00E−06 

217At 2.00E−01 1.00E−02 1.00E−02 

213Bi 1.00E−01 2.00E−03 5.00E−04 

249Cf 1.00E−03 6.00E−05 7.50E−07 

245Cm 1.00E−03 2.00E−05 2.00E−06 

221Fr 3.00E−02 3.00E−02 8.00E−03 

237Np 2.00E−02 1.00E−03 5.00E−06 

233Pa 1.00E−02 5.00E−03 5.00E−06 

209Pb 1.00E−02 8.00E−04 3.00E−04 

213Po 1.00E−03 5.00E−03 3.40E−04 

241Pu 1.00E−03 1.00E−04 1.00E−06 

225Ra 4.00E−02 1.00E−03 1.00E−03 

229Th 1.00E−03 1.00E−04 5.00E−06 

209Tl 2.00E−01 2.00E−03 3.00E−03 

233U 2.50E−03 3.40E−04 6.00E−04 

237U 2.50E−03 3.40E−04 6.00E−04 

Decay Chain 3 

227Ac 2.50E−03 2.00E−05 2.00E−05 

243Am 1.00E−03 5.00E−05 2.00E−06 

211Bi 1.00E−01 2.00E−03 5.00E−04 

251Cf 1.00E−03 6.00E−05 7.50E−07 

243Cm 1.00E−03 2.00E−05 2.00E−06 

247Cm 1.00E−03 2.00E−05 2.00E−06 

223Fr 3.00E−02 3.00E−02 8.00E−03 

239Np 2.00E−02 1.00E−03 5.00E−06 

231Pa 1.00E−02 5.00E−03 5.00E−06 

211Pb 1.00E−02 8.00E−04 3.00E−04 

211Po 1.00E−03 5.00E−03 3.40E−04 

215Po 1.00E−03 5.00E−03 3.40E−04 

239Pu 1.00E−03 1.00E−04 1.00E−06 

243Pu 1.00E−03 1.00E−04 1.00E−06 

223Ra 4.00E−02 1.00E−03 1.00E−03 
Table 5-4. (continued). 

5-26 

Nuclide 

Vegetable/soil 

(pCi/g plant 

per pCi/g soil) 

Beef 

(d/kg) 

Milk 

(d/L) 

219Rn 0.00E+00 0.00E+00 0.00E+00 

227Th 1.00E−03 1.00E−04 5.00E−06 

231Th 1.00E−03 1.00E−04 5.00E−06 

207Tl 2.00E−01 2.00E−03 3.00E−03 

235U 2.50E−03 3.40E−04 6.00E−04 

Decay Chain 4 

242Am 1.00E−03 5.00E−05 2.00E−06 

242mAm 1.00E−03 5.00E−05 2.00E−06 

210Bi 1.00E−01 2.00E−03 5.00E−04 

214Bi 1.00E−01 2.00E−03 5.00E−04 

250Cf 1.00E−03 6.00E−05 7.50E−07 

242Cm 1.00E−03 2.00E−05 2.00E−06 

246Cm 1.00E−03 2.00E−05 2.00E−06 

238Np 2.00E−02 1.00E−03 5.00E−06 

234Pa 1.00E−02 5.00E−03 5.00E−06 

234mPa 1.00E−02 5.00E−03 5.00E−06 

210Pb 1.00E−02 8.00E−04 3.00E−04 

214Pb 1.00E−02 8.00E−04 3.00E−04 

210Po 1.00E−03 5.00E−03 3.40E−04 

214Po 1.00E−03 5.00E−03 3.40E−04 

218Po 1.00E−03 5.00E−03 3.40E−04 

238Pu 1.00E−03 1.00E−04 1.00E−06 

242Pu 1.00E−03 1.00E−04 1.00E−06 

226Ra 4.00E−02 1.00E−03 1.00E−03 

222Rn 0.00E+00 0.00E+00 0.00E+00 

230Th 1.00E−03 1.00E−04 5.00E−06 

234Th 1.00E−03 1.00E−04 5.00E−06 

234U 2.50E−03 3.40E−04 6.00E−04 

238U 2.50E−03 3.40E−04 6.00E−04 
5-27 

of the pasture would be contaminated (Maheras et 

al. 1997). The intake of soil by beef cows was 

0.5 kg/d. 

A summary of the parameter values used in 

the intruder post-drilling scenario is listed in 

Table 5-5. Table 5-6 lists the unit concentration 

dose factors for the intruder post-drilling scenario. 

The unit concentration dose factors for 3

H and 14C 

were determined using the RESRAD computer 

code (Yu et al. 1993) and default parameters. 

The results of the intruder post-drilling 

scenario are provided in Section 5.6. 

5.5 Chronic Intruder PostConstruction Scenario 

Definition 

The chronic post-construction scenario 

assumes that an inadvertent intruder moves onto 

the TFF, excavates a basement, spreads the 

excavated material over the land, and 

subsequently farms the area. 

The excavation soil is assumed to be spread 

over 2,200 m2

 (24,000 ft2 or approximately onehalf acre) of land surface. The waste is assumed to 

be mixed to a depth of 0.61 m (24 in.). The 

mixing depth of 0.61 m is based on using a deep 

tilling plow to increase the depth of the root zone 

and to break up soil compaction. These plows are 

used in areas of southeast Idaho with highly 

erodible soils to minimize erosion. Deep tilling 

plows have shanks that till to a depth of 24 in. and 

are sold at Idaho Falls implement dealers 

(Maheras et al. 1997). 

The chronic post-construction scenario 

assumes that the intruder is exposed to the 

excavated soil during plowing and cultivation 

(i.e., dust inhalation). In addition, the intruder is 

assumed to ingest contaminated food products 

from the garden and from beef and milk cattle that 

consume contaminated forage. The intake of 

contaminated forage by cattle was adjusted 

according to the fraction of feed grown on 

contaminated cuttings and the necessary 

remaining feed obtained from uncontaminated 

ground. 

The soil activity concentration in the soils was 

then determined using the equation presented in 

Section 5.3. However, since the excavated soil is 

spread over a 2,200 m2

 area, this results in the 

600 m3

 of excavated soil being 0.27 m deep. The 

excavated waste soil is assumed to be mixed to a 

depth of 0.61 m. Therefore, the soil activity 

concentrations determined in Section 5.3 were 

reduced by a factor of 2.26 to account for the 

tilling and mixing of the soils to a depth of 

0.61 m. 

The determination of the unit concentration 

dose factors for each exposure pathway was 

determined using the equations in Section 5.4 for 

the intruder post-drilling scenario. 

The results of the intruder post-construction 

scenario are provided in Section 5.6. 

5.6 Intruder Analysis Results 

5.6.1 Acute Intruder-Drilling Scenario 

The soil concentrations and intruder doses 

resulting from the acute drilling scenario for each 

radionuclide for time periods of 100, 200, 300, 

400, 500, 600, 700, 800, 900, and 1,000 yr postclosure are provided in Appendix L. Please see 

this appendix for intermediate results. Figure 5-3 

shows the total intruder dose, dose from sand pad 

contamination, and dose from tank contamination. 

Figure 5-4 shows the doses from the major 

radionuclide contributors and the total from all 

nuclides. Figure 5-5 presents the intruder dose as 

a function of exposure pathway. 

The maximum dose after institutional control 

(i.e., 100 yr) for the acute intruder was 232 mrem. 

The major radionuclide contributors to the total 

acute intruder dose were 137Cs/137mBa with 

188 mrem, 238Pu with 15 mrem, 90Sr/90Y with 8.64 

mrem, 239Pu with 5.3 mrem, 241Am with 5.1 mrem, 

and 240Pu with 2.8 mrem. The acute drilling 

scenario dose results are less than the performance 

objective of 500 mrem for all times postinstitutional control.
5-28

Table 5-5. Parameter values for the chronic post-drilling scenario. 

Parameter Description Value 

Ac Area where contamination is spread (m2) 2,200 

ASR Mass loading factor (g/m3) 1.0E−03 

ASRfoliar Mass loading factor for garden (g/m3

) 1.0E−04 

Dc Final depth of contaminated zone after tilling (m) 0.61 

fin Occupancy factor indoors (unitless) 0.5 

fveg,leafy Fraction of leafy vegetation from site (unitless) 0.25 

fbeef Fraction of beef from site (unitless) 1 

fmilk Fraction of milk from site (unitless) 1 

fveg,nonleafy Fraction of nonleafy vegetation from site (unitless) 0.25 

foffsite Occupancy factor offsite (unitless) 0.25 

fout Occupancy factor outdoors (unitless) 0.25 

FAcow Area factor for resuspension on cow pathway (unitless) 0.94 

FAmilk Area factor for resuspension on milk pathway (unitless) 0.94 

FAveg Area factor for resuspension in garden (unitless) 0.94 

FARfodder Foliar deposition factor for fodder (m3

/g) 0.28 

FARleafy Foliar deposition factor for leafy vegetation (m3

/g) 0.26 

FARnonleafy Foliar deposition factor for nonleafy vegetation (m3

/g) 0.54 

Ibeef,fodder Beef animal intake of fodder (kg/d) 5.4 

Icow,soil Soil ingestion rate for cows (kg/d) 0.5 

Iinh,in Breathing rate—indoors (m3

/yr) 8,400 

Ibeef Intake of beef (g/yr) 55 × 104

Imilk Intake of milk (L/yr) 89 

Imilk,fodder Milk animal intake of fodder (kg/d) 7.2 

Iinh,out Breathing rate—outdoors (m3

/yr) 8,400 

Isoil Soil ingestion rate (g/yr) 3.65 

Iveg,leafy Human consumption of leafy vegetation (g/yr) 1.8 × 104

Iveg,nonleafy Human consumption of nonleafy vegetation (g/yr) 9.4E+04 

rcont Radius of the contaminated zone (m) 26.5 

rsand pad Radius of the sand pad (m) 7.8 

rTank Radius of the tank (m) 7.6 

rwell Radius of the well (m) 0.1 

ρ Bulk density of the contamination (g/m3

) 1.6E+06 

Sair Ratio of outdoor to indoor air (unitless) 0.4 

Sext,indoor Shielding factor for indoor exposures (unitless) 0.7b

tw Thickness of waste zone (m) 0.15 

 

a. Foliar deposition factors increased by a factor of ten for iodine. 

b. 137Cs/137mBa shielding factor is equal to 0.35 (see Appendix K). 
5-29

Table 5-6. Unit concentration dose factors for the chronic post-drilling scenario. 

Unit DCF 

Inhalation 

Unit DCF 

External 

Unit DCF 

Ingestion 

Vegetables 

Unit DCF 

Ingestion 

Beef 

Unit DCF 

Ingestion 

Milk 

Unit DCF 

Ingestion Soil 

Unit DCF 

Total 

Nuclide (mrem/yr per pCi/g) 

10Be 1.34E−03 6.35E−04 5.23E−04 1.34E−04 4.39E−07 1.28E−05 2.65E−03 

14C 1.77E−09 1.00E−09 7.60E−06 6.75E−07 5.59E−07 7.60E−10 8.84E−06 

113mCd 5.78E−03 3.83E−04 1.35E+00 7.51E−03 3.81E−02 4.41E−04 1.40E+00 

60Co 8.27E−04 8.13E+00 6.03E−02 2.76E−02 5.16E−03 7.37E−05 8.22E+00 

135Cs 1.72E−05 2.30E−05 7.92E−03 8.35E−03 3.97E−03 1.93E−05 2.03E−02 

154Eu 1.08E−03 3.95E+00 6.70E−04 5.39E−04 8.81E−06 2.61E−05 3.95E+00 

3

H 1.13E−09 0.00E+00 1.44E−07 1.17E−08 2.78E−08 1.53E−12 1.85E−07 

166mHo 2.92E−03 5.49E+00 5.89E−04 4.56E−04 7.45E−06 2.21E−05 5.50E+00 

129I 6.56E−04 7.77E−03 1.55E−01 6.46E−02 1.59E−01 7.56E−04 3.88E−01 

115In 1.41E−02 2.38E−04 1.33E−02 1.79E−02 1.46E−03 4.31E−04 4.74E−02 

138La 5.17E−03 3.98E+00 4.13E−04 3.32E−04 5.43E−06 1.61E−05 3.98E+00 

94Nb 1.68E−03 5.08E+00 2.00E−03 6.53E−08 7.27E−07 1.95E−05 5.08E+00 

63Ni 1.17E−05 0.00E+00 8.08E−04 1.22E−04 8.84E−04 1.58E−06 1.83E−03 

107Pd 4.83E−05 0.00E+00 4.19E−04 8.55E−06 8.12E−05 4.09E−07 5.57E−04 

146Pm 5.54E−04 2.36E+00 2.58E−04 2.07E−04 3.38E−06 1.00E−05 2.37E+00 

87Rb 1.22E−05 8.43E−05 1.79E−02 4.88E−03 6.29E−03 1.35E−05 2.92E−02 

79Se 3.72E−05 1.12E−05 2.43E−02 4.97E−02 9.44E−03 2.38E−05 8.36E−02 

146Sm 3.12E−01 0.00E+00 1.43E−02 1.15E−02 1.88E−04 5.58E−04 3.38E−01 

151Sm 1.13E−04 5.91E−07 2.73E−05 2.20E−05 3.58E−07 1.06E−06 1.65E−04 

121mSn 4.35E−05 1.18E−03 1.09E−04 4.38E−04 7.15E−05 4.24E−06 1.84E−03 

98Tc 8.64E−05 4.55E+00 6.84E−01 7.39E−04 1.59E−02 1.34E−05 5.25E+00 

99Tc 3.15E−05 7.51E−05 2.05E−01 2.21E−04 4.75E−03 4.00E−06 2.10E−01 

123Te 3.99E−05 3.93E−03 7.02E−02 6.02E−03 8.97E−04 1.14E−05 8.11E−02 

Short Decay Chains 

108Ag 0.00E+00 6.18E−02 0.00E+00 0.00E+00 0.00E+00 0.00E+00 6.18E−02 

108mAg 1.07E−03 5.17E+00 3.20E−02 1.65E−03 2.68E−02 2.09E−05 5.23E+00 

137mBa 0.00E+00 1.36E+00 0.00E+00 0.00E+00 0.00E+00 0.00E+00 1.36E+00 

137Cs 1.21E−04 3.13E−04 5.60E−02 5.90E−02 2.80E−02 1.37E−04 1.44E−01 

152Eu 8.35E−04 3.61E+00 4.55E−04 3.66E−04 5.97E−06 1.77E−05 3.61E+00 
Table 5-6. (continued). 

5-30 

Unit DCF 

Inhalation 

Unit DCF 

External 

Unit DCF 

Ingestion 

Vegetables 

Unit DCF 

Ingestion 

Beef 

Unit DCF 

Ingestion 

Milk 

Unit DCF 

Ingestion Soil 

Unit DCF 

Total 

Nuclide (mrem/yr per pCi/g) 

152Gd 9.20E−01 0.00E+00 1.13E−02 9.07E−03 1.48E−04 4.40E−04 9.41E−01 

93mNb 1.10E−04 6.24E−05 1.46E−04 4.77E−09 5.31E−08 1.43E−06 3.21E−04 

147Pm 1.48E−04 2.99E−05 7.35E−05 5.92E−05 9.66E−07 2.87E−06 3.15E−04 

126Sb 4.43E−05 9.11E+00 3.00E−03 3.26E−04 5.45E−05 2.93E−05 9.12E+00 

126mSb 1.28E−07 4.98E+00 2.62E−05 2.85E−06 4.77E−07 2.56E−07 4.98E+00 

147Sm 2.83E−01 0.00E+00 1.30E−02 1.05E−02 1.71E−04 5.07E−04 3.07E−01 

126Sn 3.76E−04 8.85E−02 1.37E−03 5.51E−03 8.99E−04 5.34E−05 9.68E−02 

90Sr 4.91E−03 4.17E−04 1.20E+00 1.33E−01 6.75E−02 3.90E−04 1.40E+00 

90Y 3.19E−05 1.34E−02 9.04E−02 1.00E−02 5.10E−03 2.95E−05 1.19E−01 

93Zr 1.21E−03 0.00E+00 4.68E−05 4.61E−08 4.49E−08 4.54E−06 1.26E−03 

Decay Chain 1 

228Ac 1.17E−03 3.09E+00 1.52E−04 1.22E−06 2.00E−06 5.93E−06 3.09E+00 

212Bi 8.15E−05 6.01E−01 2.97E−03 1.21E−04 5.77E−05 2.91E−06 6.04E−01 

252Cf 5.93E−01 1.05E−04 3.06E−02 1.81E−03 3.67E−05 2.97E−03 6.29E−01 

244Cm 9.37E−01 7.55E−05 5.69E−02 1.12E−03 1.82E−04 5.52E−03 1.00E+00 

248Cm 6.25E+00 5.27E−05 3.84E−01 7.57E−03 1.23E−03 3.73E−02 6.68E+00 

240Np 3.08E−07 4.09E+00 1.33E−04 7.93E−06 6.79E−08 6.49E−07 4.09E+00 

240mNp 0.00E+00 1.06E+00 0.00E+00 0.00E+00 0.00E+00 0.00E+00 1.06E+00 

212Pb 6.38E−04 4.06E−01 1.28E−02 1.11E−03 6.95E−04 1.25E−04 4.21E−01 

212Po 0.00E+00 0.00E+00 0.00E+00 0.00E+00 0.00E+00 0.00E+00 0.00E+00 

216Po 0.00E+00 5.46E−05 0.00E+00 0.00E+00 0.00E+00 0.00E+00 5.46E−05 

236Pu 5.47E−01 1.34E−04 3.29E−02 3.24E−03 5.26E−05 3.19E−03 5.86E−01 

240Pu 1.62E+00 8.79E−05 9.99E−02 9.84E−03 1.60E−04 9.68E−03 1.74E+00 

244Pu 1.52E+00 4.53E−05 9.37E−02 9.23E−03 1.50E−04 9.09E−03 1.64E+00 

224Ra 1.19E−02 2.94E−02 4.10E−01 1.44E−02 2.57E−02 1.00E−03 4.92E−01 

228Ra 1.80E−02 0.00E+00 1.61E+00 5.65E−02 1.01E−01 3.93E−03 1.79E+00 

220Rn 0.00E+00 1.23E−03 0.00E+00 0.00E+00 0.00E+00 0.00E+00 1.23E−03 

228Th 1.29E+00 4.67E−03 1.12E−02 1.10E−03 8.94E−05 1.08E−03 1.31E+00 

232Th 6.20E+00 3.12E−04 7.71E−02 7.59E−03 6.17E−04 7.48E−03 6.29E+00 
Table 5-6. (continued). 

5-31 

Unit DCF 

Inhalation 

Unit DCF 

External 

Unit DCF 

Ingestion 

Vegetables 

Unit DCF 

Ingestion 

Beef 

Unit DCF 

Ingestion 

Milk 

Unit DCF 

Ingestion Soil 

Unit DCF 

Total 

Nuclide (mrem/yr per pCi/g) 

208Tl 0.00E+00 1.08E+01 0.00E+00 0.00E+00 0.00E+00 0.00E+00 1.08E+01 

232U 5.62E−02 5.35E−04 9.20E−02 1.26E−02 3.62E−02 3.59E−03 2.01E−01 

236U 4.74E−01 1.28E−04 1.89E−02 2.58E−03 7.43E−03 7.35E−04 5.04E−01 

240U 8.57E−06 8.54E−04 3.12E−04 4.26E−05 1.23E−04 1.22E−05 1.35E−03 

Decay Chain 2 

225Ac 4.08E−02 3.74E−02 7.80E−03 6.27E−05 1.02E−04 3.04E−04 8.65E−02 

241Am 1.68E+00 2.62E−02 1.03E−01 5.06E−03 3.29E−04 9.97E−03 1.82E+00 

217At 0.00E+00 9.65E−04 0.00E+00 0.00E+00 0.00E+00 0.00E+00 9.65E−04 

213Bi 6.48E−05 4.20E−01 2.02E−03 8.26E−05 3.92E−05 1.98E−06 4.23E−01 

249Cf 2.18E+00 1.03E+00 1.34E−01 7.90E−03 1.60E−04 1.30E−02 3.37E+00 

245Cm 1.72E+00 2.02E−01 1.06E−01 2.08E−03 3.38E−04 1.02E−02 2.04E+00 

221Fr 0.00E+00 8.85E−02 0.00E+00 0.00E+00 0.00E+00 0.00E+00 8.85E−02 

237Np 2.04E+00 4.66E−02 2.49E+00 1.49E−01 1.27E−03 1.22E−02 4.74E+00 

233Pa 3.61E−05 5.78E−01 1.02E−03 5.53E−04 9.24E−07 9.94E−06 5.80E−01 

209Pb 3.58E−07 4.57E−04 5.96E−05 5.19E−06 3.25E−06 5.82E−07 5.26E−04 

213Po 0.00E+00 0.00E+00 0.00E+00 0.00E+00 0.00E+00 0.00E+00 0.00E+00 

241Pu 3.12E−02 3.53E−06 1.93E−03 1.90E−04 3.09E−06 1.87E−04 3.35E−02 

225Ra 2.94E−02 6.61E−03 4.31E−01 1.52E−02 2.70E−02 1.05E−03 5.10E−01 

229Th 8.11E+00 1.91E−01 9.97E−02 9.81E−03 7.97E−04 9.66E−03 8.42E+00 

209Tl 0.00E+00 6.49E+00 0.00E+00 0.00E+00 0.00E+00 0.00E+00 6.49E+00 

233U 5.12E−01 8.11E−04 2.03E−02 2.78E−03 8.00E−03 7.91E−04 5.45E−01 

237U 1.33E−05 3.12E−01 2.23E−04 3.05E−05 8.77E−05 8.68E−06 3.12E−01 

Decay Chain 3 

227Ac 2.53E+01 2.94E−04 9.87E−01 7.94E−03 1.30E−02 3.85E−02 2.64E+01 

243Am 1.66E+00 8.52E−02 1.02E−01 5.04E−03 3.27E−04 9.92E−03 1.87E+00 

211Bi 0.00E+00 1.43E−01 0.00E+00 0.00E+00 0.00E+00 0.00E+00 1.43E−01 

251Cf 2.22E+00 3.09E−01 1.37E−01 8.09E−03 1.64E−04 1.33E−02 2.69E+00 

243Cm 1.16E+00 3.38E−01 7.09E−02 1.40E−03 2.27E−04 6.88E−03 1.58E+00 

247Cm 1.57E+00 9.86E−01 9.65E−02 1.90E−03 3.09E−04 9.36E−03 2.66E+00 

223Fr 2.97E−02 1.13E−01 7.24E−03 9.42E−03 4.40E−03 2.36E−05 1.64E−01 
Table 5-6. (continued). 

5-32 

Unit DCF 

Inhalation 

Unit DCF 

External 

Unit DCF 

Ingestion 

Vegetables 

Unit DCF 

Ingestion 

Beef 

Unit DCF 

Ingestion 

Milk 

Unit DCF 

Ingestion Soil 

Unit DCF 

Total 

Nuclide (mrem/yr per pCi/g) 

239Np 9.48E−06 4.37E−01 1.83E−03 1.09E−04 9.36E−07 8.93E−06 4.39E−01 

231Pa 4.85E+00 1.08E−01 2.97E+00 1.61E+00 2.69E−03 2.90E−02 9.57E+00 

211Pb 2.85E−05 1.64E−01 1.47E−04 1.28E−05 8.03E−06 1.44E−06 1.64E−01 

211Po 0.00E+00 2.51E−02 0.00E+00 0.00E+00 0.00E+00 0.00E+00 2.51E−02 

215Po 0.00E+00 5.58E−04 0.00E+00 0.00E+00 0.00E+00 0.00E+00 5.58E−04 

239Pu 1.62E+00 1.70E−04 9.99E−02 9.84E−03 1.60E−04 9.68E−03 1.74E+00 

243Pu 6.21E−07 4.71E−02 9.42E−06 9.28E−07 1.51E−08 9.14E−07 4.71E−02 

223Ra 2.97E−02 3.47E−01 7.38E−01 2.59E−02 4.62E−02 1.80E−03 1.19E+00 

219Rn 0.00E+00 1.73E−01 0.00E+00 0.00E+00 0.00E+00 0.00E+00 1.73E−01 

227Th 6.11E−02 2.97E−01 1.08E−03 1.06E−04 8.60E−06 1.04E−04 3.59E−01 

231Th 3.31E−06 2.17E−02 3.81E−05 3.76E−06 3.05E−07 3.70E−06 2.18E−02 

207Tl 0.00E+00 1.06E−02 0.00E+00 0.00E+00 0.00E+00 0.00E+00 1.06E−02 

235U 4.64E−01 4.24E−01 1.87E−02 2.56E−03 7.36E−03 7.28E−04 9.17E−01 

Decay Chain 4 

242Am 2.21E−04 2.99E−02 3.98E−05 1.96E−06 1.27E−07 3.86E−06 3.02E−02 

242mAm 1.61E+00 1.01E−03 9.92E−02 4.89E−03 3.17E−04 9.62E−03 1.72E+00 

210Bi 7.40E−04 2.08E−03 1.79E−02 7.32E−04 3.48E−04 1.75E−05 2.18E−02 

214Bi 2.49E−05 4.89E+00 7.92E−04 3.23E−05 1.53E−05 7.74E−07 4.89E+00 

250Cf 9.90E−01 7.11E−05 6.02E−02 3.56E−03 7.22E−05 5.83E−03 1.06E+00 

242Cm 6.53E−02 1.02E−04 3.24E−03 6.38E−05 1.04E−05 3.14E−04 6.90E−02 

246Cm 1.71E+00 6.97E−05 1.04E−01 2.06E−03 3.34E−04 1.01E−02 1.82E+00 

238Np 1.40E−04 1.77E+00 2.24E−03 1.34E−04 1.15E−06 1.09E−05 1.77E+00 

234Pa 3.08E−06 6.03E+00 6.06E−04 3.29E−04 5.50E−07 5.92E−06 6.03E+00 

234mPa 0.00E+00 4.71E−02 0.00E+00 0.00E+00 0.00E+00 0.00E+00 4.71E−02 

210Pb 5.13E−02 1.47E−03 1.50E+00 1.31E−01 8.20E−02 1.47E−02 1.78E+00 

214Pb 2.95E−05 7.51E−01 1.75E−04 1.52E−05 9.55E−06 1.71E−06 7.51E−01 

210Po 3.55E−02 2.75E−05 5.37E−02 2.64E−01 2.92E−02 5.21E−03 3.88E−01 

214Po 0.00E+00 2.69E−04 0.00E+00 0.00E+00 0.00E+00 0.00E+00 2.69E−04 

218Po 0.00E+00 2.95E−05 0.00E+00 0.00E+00 0.00E+00 0.00E+00 2.95E−05 
Table 5-6. (continued). 

5-33 

Unit DCF 

Inhalation 

Unit DCF 

External 

Unit DCF 

Ingestion 

Vegetables 

Unit DCF 

Ingestion 

Beef 

Unit DCF 

Ingestion 

Milk 

Unit DCF 

Ingestion Soil 

Unit DCF 

Total 

Nuclide (mrem/yr per pCi/g) 

238Pu 1.48E+00 9.04E−05 9.04E−02 8.90E−03 1.45E−04 8.76E−03 1.59E+00 

242Pu 1.55E+00 7.68E−05 9.48E−02 9.34E−03 1.52E−04 9.20E−03 1.67E+00 

226Ra 3.24E−02 1.85E−02 1.48E+00 5.22E−02 9.29E−02 3.63E−03 1.68E+00 

222Rn 0.00E+00 1.28E−03 0.00E+00 0.00E+00 0.00E+00 0.00E+00 1.28E−03 

230Th 1.23E+00 7.16E−04 1.55E−02 1.52E−03 1.24E−04 1.50E−03 1.25E+00 

234Th 1.32E−04 1.45E−02 3.85E−04 3.80E−05 3.08E−06 3.74E−05 1.51E−02 

234U 5.01E−01 2.40E−04 1.99E−02 2.72E−03 7.84E−03 7.76E−04 5.32E−01 

238U 4.48E−01 6.19E−05 1.79E−02 2.45E−03 7.04E−03 6.97E−04 4.76E−01 

1

10

100

1000

0 100 200 300 400 500 600 700 800 900 1000

Time (years post-closure)

Dose (mrem)

Tank

Sand Pad

All

Acute Dose Limit 500 mrem/yr

Figure 5-3. Results of the acute intruder-drilling scenario.
5-34

1.00E-03

1.00E-02

1.00E-01

1.00E+00

1.00E+01

1.00E+02

1.00E+03

1.00E+04

0 100 200 300 400 500 600 700 800 900 1000

Time (years post-closure)

Dose (mrem)

Nb-94

Sr-90+D

Cs-137+D

Pu-240

Am-241

Pu-239

Pu-238

Np-237

Total Dose

Acute Dose Limit 500 mrem

Figure 5-4. Results of the acute intruder-drilling scenario for individual radionuclides. 

1.00E-01

1.00E+00

1.00E+01

1.00E+02

1.00E+03

0 100 200 300 400 500 600 700 800 900 1000

Time (years post-closure)

Dose (mrem)

Inhalation

External

Ingestion

Total

Acute Dose Limit 500 mrem/yr

Figure 5-5. Results of the acute intruder-drilling scenario by exposure pathway.
5-35 

5.6.2 Acute Intruder-Construction 

Scenario 

The soil concentrations and intruder doses 

resulting from the acute drilling scenario for each 

radionuclide for time periods of 100, 200, 300, 

400, 500, 600, 700, 800, 900, and 1,000 yr postclosure are provided in Appendix L. Please see 

this appendix for intermediate results. Figure 5-6 

shows the doses from the major radionuclide 

contributors and the total from all nuclides. 

Figure 5-7 presents the intruder dose as a function 

of exposure pathway. 

The maximum dose after institutional control 

(i.e., 100 yr) for the acute intruder-construction 

scenario was 0.93 mrem. The major radionuclide 

contributors to the total acute intruder dose were 

137Cs/137mBa with 0.72 mrem, 94Nb with 

0.045 mrem, 90Sr/90Y with 0.017 mrem, and 238Pu 

with 0.013 mrem. The acute intruder-construction 

scenario dose results are less than the performance 

objective of 500 mrem for all times post-closure. 

5.6.3 Chronic Intruder Post-Drilling 

Scenario 

The soil concentrations and intruder doses 

resulting from the chronic post-drilling scenario 

for each radionuclide for time periods of 100, 200, 

300, 400, 500, 600, 700, 800, 900, and 1,000 yr 

post-closure are provided in Appendix L. Please 

see this appendix for intermediate results. 

Figure 5-8 shows the total intruder dose, dose 

from sand pad contamination, and dose from tank 

contamination. Figure 5-9 shows the doses from 

the major radionuclide contributors and the total 

from all nuclides. Figure 5-10 presents the intruder 

dose as a function of exposure pathway. 

The maximum dose after institutional control 

was 91.1 mrem/yr for the chronic intruder postdrilling scenario at 100 yr. The major radionuclide 

contributors to the total chronic intruder dose were 

90Sr/90Y with 51.5 mrem/yr and 137Cs/137mBa with 

36.5 mrem/yr. The chronic intruder post-drilling 

scenario dose results do not exceed the 

performance objective of 100 mrem/yr during the 

compliance period. 

5.6.4 Chronic Intruder PostConstruction Scenario 

The soil concentrations and intruder doses 

resulting from the chronic post-construction 

scenario for each radionuclide for time periods of 

100, 200, 300, 400, 500, 600, 700, 800, 900, and 

1,000 yr post-closure are provided in Appendix L. 

Please see this appendix for intermediate results. 

Figure 5-11 shows the doses from the major 

radionuclide contributors and the total from all 

nuclides. Figure 5-12 presents the intruder dose as 

a function of exposure pathway. 

The maximum dose was 26.1 mrem/yr for the 

chronic intruder post-construction scenario at 

100 yr post-closure. The major radionuclide 

contributors to the total chronic intruder dose 

were90Sr/90Y with 15.2 mrem/yr, 137Cs/137mBa with 

10.1 mrem/yr, 94Nb with 0.58 mrem/yr, and 238Pu 

with 0.15 mrem/yr. The chronic intruder postconstruction scenario dose results do not exceed 

the performance objective of 100 mrem/yr for the 

1,000-yr compliance period. 

5.7 Intruder Sensitivity/ 

Uncertainty Analysis 

DOE guidance for PAs (DOE 1999b) states 

that the sensitivity/uncertainty analysis for 

hypothetical inadvertent intruder analyses should 

be limited to qualitative arguments (e.g., 

explanation of the rationale for selected scenarios 

and parameters). Therefore, the discussion in this 

section is limited to qualitative rationale. 

The preceding sections describe the TFF as a 

unique environment for the evaluation of intruder 

scenarios. Many of the standard scenarios were not 

considered applicable to the TFF because the 

depth of the waste in the tanks is 10 m (30 ft). 

Therefore, only the acute intruder-drilling and 

chronic post-drilling scenarios were considered 

applicable. 

It is difficult to predict the future actions of 

humans, especially over a period of 1,000 yr. The 

fact that the intruder is assumed to drill a well 

directly through a tank is considered a very 

conservative assumption. It can be argued that the 
5-36 

1.00E-05

1.00E-04

1.00E-03

1.00E-02

1.00E-01

1.00E+00

1.00E+01

1.00E+02

1.00E+03

1.00E+04

0 100 200 300 400 500 600 700 800 900 1000

Time (years post-closure)

Dose (mrem)

Nb-94

Sr-90+D

Cs-137+D

Pu-240

Am-241

Pu-239

Pu-238

Np-237

Total Dose

Acute Dose Limit 500 mrem

Figure 5-6. Results of the acute intruder-construction scenario for individual radionuclides. 

1.00E-04

1.00E-03

1.00E-02

1.00E-01

1.00E+00

1.00E+01

1.00E+02

1.00E+03

0 100 200 300 400 500 600 700 800 900 1000

Time (years post-closure)

Dose (mrem)

Inhalation

External

Ingestion

Total

Acute Dose Limit 500 mrem/yr

Figure 5-7. Results of the acute intruder-construction scenario by exposure pathway. 
5-37 

0.1

1

10

100

1000

0 100 200 300 400 500 600 700 800 900 1000

Time (years post-closure)

Dose (mrem/yr)

Tank

Sand Pad

All

Chronic Dose Limit 100 mrem/yr

Figure 5-8. Results of the chronic intruder post-drilling scenario. 

1.00E-02

1.00E-01

1.00E+00

1.00E+01

1.00E+02

1.00E+03

0 100 200 300 400 500 600 700 800 900 1000

Time (years post-closure)

Dose (mrem/yr)

Nb-94

Sr-90+D

Cs-137+D

Pu-240

Am-241

Pu-239

Pu-238

Np-237

total

Chronic Dose Limit 100 mrem/yr

Figure 5-9. Results of the chronic intruder post-drilling scenario for individual radionuclides. 
5-38 

1.00E-04

1.00E-03

1.00E-02

1.00E-01

1.00E+00

1.00E+01

1.00E+02

1.00E+03

0 100 200 300 400 500 600 700 800 900 1000

Time (years post-closure)

Dose (mrem/yr)

Inhalation

External

Ingestion Vegetables

Ingestion Beef

Ingestion Milk

Ingestion Soil

Total

Chronic Dose Limit 100 mrem/yr

Figure 5-10. Results of the chronic intruder post-drilling scenario by exposure pathway. 

1.00E-04

1.00E-03

1.00E-02

1.00E-01

1.00E+00

1.00E+01

1.00E+02

1.00E+03

0 100 200 300 400 500 600 700 800 900 1000

Time (years post-closure)

Dose (mrem/yr)

Nb-94

Sr-90+D

Cs-137+D

Pu-240

Am-241

Pu-239

Pu-238

Np-237

total

Chronic Dose Limit 100 mrem/yr

Figure 5-11. Results of the chronic intruder post-construction scenario for individual radionuclides. 
5-39 

1.00E-06

1.00E-05

1.00E-04

1.00E-03

1.00E-02

1.00E-01

1.00E+00

1.00E+01

1.00E+02

1.00E+03

0 100 200 300 400 500 600 700 800 900 1000

Time (years post-closure)

Dose (mrem/yr)

Inhalation

External

Ingestion Vegetables

Ingestion Beef

Ingestion Milk

Ingestion Soil

Total

Chronic Dose Limit 100 mrem/yr

Figure 5-12. Results of the chronic intruder post-construction scenario by exposure pathway. 

person drilling the well would not continue to drill 

through the concrete and stainless-steel tanks and 

grout unless it was completely degraded. 

However, to be conservative, it was assumed the 

intruder would drill the well through the waste at 

100 yr post-closure. 

Uncertainty exists in the state of the concrete 

systems over time. In this analysis, credit was 

taken for the concrete vault providing a barrier to 

intrusion. Oztunali and Roles (1986) state that “if 

resistance is encountered during drilling (as would 

be the case if the drill bit hit solid rocks or intact 

metal), the crew would simply move a few yards 

horizontally and drill a new hole. Drilling 

equipment and bits that can drill through solid 

rock and intact metal are available; however, they 

are more expensive than the equipment and bits 

that are normally required and used for water well 

installation. Similarly, it is anticipated that 

standard drill bits would have difficulty 

penetrating reinforced concrete structures, 

although little difficulty is expected for waste 

stabilized using grout backfill.” Therefore, credit 

could be taken for the reinforced concrete vaults 

and the stainless-steel tanks, further reducing the 

doses for the intruder. 

Intruder dose assessments include parameter 

uncertainty in radionuclide dose conversion 

factors and human intrusion and terrestrial 

transport models. The uncertainty in the 

radionuclide dose conversion factors is the subject 

of considerable research and is beyond the scope 

of these analyses. The dose conversion factors 

used were taken from Federal Guidance Reports 

11 and 12 (EPA 1988 and 1993). The most 

conservative values were chosen for use in the 

intruder analysis. This approach is considered 

reasonable since there is no information available 

on the chemical and physical form of the waste 

constituents. 

In Equation (5-4) for the acute intruderdrilling scenario, the radius of the well is not 

present in the equation. Note, however, that the 

concentration in the soil will be the same for any 

given well radius but the total amount of 

contaminated soil will vary with the well radius. 

This equation assumes that the radionuclide 

activity is completely mixed with the drill cuttings. 
5-40 

The radionuclide activity would not likely be 

completely mixed with the well cuttings. 

However, when considering the different 

possibilities such as the cuttings from the tank 

being covered with cuttings from below the tank 

and the varying radiation fields that would be 

encountered as the drilling commenced with time, 

it seems appropriate to make this assumption of 

complete mixing. 

A 22-in.-diameter well provides 30 m3

(1,000 ft3

) of contaminated zone (i.e., over the 

depth of the well borehole of 122 m [400 ft]). 

When the soil and waste are spread out over the 

assumed contaminated zone of 2,200 m2

(24,000 ft2

), it results in a contaminated zone 

0.014 m (0.54 in.) thick. As the well diameter is 

increased, the thickness of the contaminated zone 

also increases, thus affecting the external dose 

rate. However, in this intruder analysis, a very 

large diameter well was used for purposes of 

conservatism. Additionally, the assumption that 

the drill cuttings are spread out over the ground 

and are not located in a mud pit adds more 

conservatism and site-specific application that 

increases the external dose for this scenario. 

Other conservative parameters were used in 

the acute intruder-drilling scenario as well. These 

include the assumptions that the intruder spends 

160 hr drilling the well, respires at a rate of 

8,400 m3

/yr, and is exposed to a dust loading of 

1E+03 g/m3

. Standard drilling scenarios typically 

use an exposure time of 6 hr (Oztunali and Roles 

1986). 

In Equation (5-15) for the chronic post-drilling 

scenario, two parameters of importance are present 

in the form of the radius of the well (rwell) and the 

final depth of the contaminated zone after tilling 

(Dc). Variations in the values used for these 

parameters will result in proportional changes in 

the predicted radionuclide concentrations in the 

contaminated zone. Therefore, if the depth of the 

contaminated zone after tilling (Dc) is reduced by a 

factor of 2, the contaminated zone concentrations 

increase by a factor of 2. If the radius of the well 

(rwell) is increased by a factor of 2, the 

contaminated zone concentrations increase by a 

factor of 4. The selection of parameter values for 

Dc and rwell are considered to be appropriate and 

based on site-specific information. 

The uncertainty in the intruder analyses is 

large and includes both the assumptions made 

during the development of the scenarios and the 

parameter values selected in the formulation of the 

mathematical representations. However, in this 

PA, every attempt was made to consider the sitespecific environment and habits of the people 

currently in the region. Therefore, the analyses are 

considered to be representative of the INEEL 

region. The predicted doses are considered to be 

conservative, and there is a reasonable assurance 

that the performance objectives will not be 

exceeded for 1,000 yr post-closure. 

Section 7 presents additional analysis of the 

uncertainty in the intruder dose results.
6-1 

6. INTERPRETATION OF RESULTS 

This section gives an interpretation of the 

results given in Sections 4 and 5. Section 7 

provides additional discussion on the sensitivity 

and uncertainty in the analyses. 

6.1 All-Pathways Dose 

The all-pathways dose calculations were 

dominated by the groundwater pathway with a 

dose of 1.4 mrem/yr. The atmospheric dose 

contributes and additional 0.51 mrem/yr to the 

member of the public. 

According to the results presented in 

Section 4, the doses to an individual of 

1.4 mrem/yr at 890 yr for the all-pathways 

groundwater scenario did not exceed the 

performance objective of 25 mrem/yr for 1,000 yr 

after project completion. The long-term analysis in 

Section 4 also indicates that peak doses do not 

exceed the performance objective during the 

period of closure to 1 million years post-closure. 

The atmospheric all-pathways dose of 

0.51 mrem/yr, when added to the groundwater 

pathway dose of 1.4 mrem/yr, is still below the 

performace objective of 25 mrem/yr. 

Uncertainty in the results is discussed in 

Section 7. 

6.2 Radon Flux Results 

Calculations were performed for a onedimensional bare slab resting on the ground 

surface to determine the radon flux at the surface 

of the slab (Section 3). These calculations are 

highly conservative with respect to the actual 

performance of a grouted heel at the TFF because 

the slab is modeled in open air. This means that no 

credit is taken for the vault or tank. Also, the 

radium inventory at the end of the 1,000-yr 

analysis period is used for the source in the 

calculations, which is conservative for all times 

before the end of the performance period. This fact 

means that more radon can diffuse out of the slab 

than would otherwise be expected. 

For 222Rn (the dominant isotope in terms of 

atomic fraction), the flux of radon at the surface of 

the slab was 0.39 pCi/m2

/s, which is more than one 

order of magnitude below the performance 

objective of 20 pCi/m2

/s. Given the conservative 

nature of the calculation, there are two reasons that 

this value is so low: cementitious materials do not 

transport radon very well, and the radium content 

of the grouted heel is relatively low. Although the 

bare slab model could not be shown to meet the 

standard for the shorter-lived isotopes of radon 

(

220Rn, half-life of 55.6 s, and 219Rn, half-life of 

3.96 s), Section 3.3.3 shows that 0.5 m (2 ft) of 

overburden is capable of reducing these radon 

fluxes significantly. Therefore, the shorter-lived 

isotopes are quite incapable of penetrating even a 

modest overburden. Because of their shorter halflives, the isotopes do not survive long enough to 

traverse the barrier. 

6.3 Volatile Radionuclide 

Releases 

The annual dose from volatile releases 

predicted by CAP88-PC atmospheric transport 

runs resulted in a dose for 3

H releases of 

0.008 mrem/yr and 14C doses of 0.5 mrem/yr, with 

a total dose of 0.5 mrem/yr. These doses are well 

below the atmospheric pathway performance 

objective of 10 mrem/yr. This was a very 

conservative analysis because flux rates were 

based on simplified models and the receptor was 

located at 100 m (300 ft) from the facility at all 

times (see Section 3.3.4). Also, the receptor would 

not be located near the TFF for the first 100 yr 

after closure because the site is assumed to be 

maintained and patrolled. 

6.4 Groundwater Protection 

The fundamental approach used in the 

numerical simulation of the unsaturated zone was 

to rely on the USGS definition of the site 

hydrostratigraphy, estimates of precipitation, and 

Big Lost River and percolation ponds seepage 

values, which were confirmed by several sources 

from the available literature (Anderson, Kuntz, 

and Davis 1999; Anderson 1991; Rodriguez et al. 
6-2 

1997). The hydraulic conductivity of the lowpermeable interbed was varied within reasonable 

ranges, based on lithologic descriptions, to predict 

the location and extent of the numerous perched 

water bodies underlying the INTEC facility. 

A two-dimensional groundwater flow and 

transport model was used to simulate contaminant 

releases from the tanks. The model domain 

dissected the center of two tanks, one sand pad, 

and included the releases from a break in the 

piping, providing the maximum release of 

contaminants to the underlying soil and 

groundwater. Since the dose limit calculation was 

applied to the two-dimensional slice through the 

center of the tanks, this approach ensures that all 

of the tanks will meet the groundwater 

performance objectives. In addition, the twodimensional discretization of the problem domain 

allowed a high degree of resolution of the 

hydrostratigraphy, including the discontinuous 

interbeds that form the perched zones underlying 

the facility. By assuming that the USGS cross 

sections were accurate, the resulting numerical 

model predicted contaminant transport around the 

perched zones and through breaks in the interbeds. 

Dilution by perched water was minimized and the 

point where the maximum contaminant 

concentrations entered the regional aquifer was 

identified and selected as the compliance point. 

According to the results presented in 

Section 4, the doses to an individual of 

0.77 mrem/yr at 890 yr for the all-pathways 

scenario did not exceed the performance objective 

of 4 mrem/yr for 1,000 yr after project completion. 

The long-term analysis in Section 4 also indicates 

that peak doses do not exceed the performance 

objective during the period of closure to 

1 million years post-closure. 

Uncertainty in the results is discussed in 

Section 7. 

6.5 Intruder Analysis Results 

The maximum dose after institutional control 

(i.e., 100 yr) for the acute intruder-drilling 

scenario was 232 mrem. The major radionuclide 

contributors to the total acute intruder dose were 

137Cs/137mBa with 188 mrem, 238Pu with 15 mrem, 

90Sr/90Y with 8.64 mrem, 239Pu with 5.3 mrem, 

241Am with 5.1 mrem, and 240Pu with 2.8 mrem. 

The acute drilling scenario dose results are less 

than the performance objective of 500 mrem for 

all times post-institutional control. 

The maximum dose post-institutional control 

(i.e., 100 yr) for the acute intruder-construction 

scenario was 0.93 mrem. The major radionuclide 

contributors to the total acute intruder dose were 

137Cs/137mBa with 0.72 mrem, 94Nb with 

0.045 mrem, 90Sr/90Y with 0.017 mrem, and 238Pu 

with 0.013 mrem. The acute intruder-construction 

scenario dose results are less than the performance 

objective of 500 mrem for all times post-closure. 

The maximum dose after loss of the concrete 

vault integrity was 91.1 mrem/yr for the chronic 

intruder post-drilling scenario at 100 yr postclosure. The major radionuclide contributors to the 

total chronic intruder dose were 90Sr/90Y with 

51.5 mrem/yr and 137Cs/137mBa with 36.5 mrem/yr. 

The chronic intruder post-drilling scenario dose 

results do not exceed the performance objective of 

100 mrem/yr during the compliance period. 

The maximum dose was 26.1 mrem/yr for the 

chronic intruder post-construction scenario at 

100 years post-closure. The major radionuclide 

contributors to the total chronic intruder dose were 

90Sr/90Y with 15.2 mrem/yr, 137Cs/137mBa with 

10.1 mrem/yr, 94Nb with 0.58 mrem/yr, and 238Pu 

with 0.15 mrem/yr. The chronic intruder postconstruction scenario dose results do not exceed 

the performance objective of 100 mrem/yr for the 

1,000-yr compliance period. 

It is difficult to predict the future actions of 

humans, especially over a 1,000-yr period. The 

fact that the intruder is assumed to drill a well 

directly through a tank is considered a very 

conservative assumption. It can be argued that the 

person drilling the well would not continue to drill 

through the concrete and stainless-steel tanks and 

grout unless it was completely degraded. The 

analysis is very sensitive to the time of intrusion 

because the short-lived fission products are the 

main contributor to the intruder doses at 100 yr 

post-closure. 
6-3 

Uncertainty exists in the state of the concrete 

systems over time. In this analysis, credit was not 

taken for the concrete vault providing a barrier to 

intrusion. Oztunali and Roles (1986) state that “if 

resistance is encountered during drilling (as would 

be the case if the drill bit hit solid rocks or intact 

metal), the crew would simply move a few yards 

horizontally and drill a new hole. Drilling 

equipment and bits that can drill through solid 

rock and intact metal are available; however, they 

are more expensive than the equipment and bits 

that are normally required and used for water well 

installation. Similarly, it is anticipated that 

standard drill bits would have difficulty 

penetrating reinforced concrete structures, 

although little difficulty is expected for waste 

stabilized using grout backfill.” Therefore, credit 

could be taken for the reinforced concrete vaults 

and the stainless-steel tanks, further reducing the 

doses for the intruder. 
 

7-1

7. PERFORMANCE EVALUATION 

A PA is the systematic analysis of the risks 

posed by a waste management system to the 

general public and the environment, and a 

comparison of those risks to the performance 

objectives. Section 7.1 provides an evaluation of 

the PA results with respect to the performance 

objectives listed in Section 1.4. A sensitivity/ 

uncertainty analysis is presented in Section 7.3. 

The results of the groundwater pathway analyses 

are discussed in comparison to the results 

presented in Section 7.1 to provide confidence 

that the projected performance and resulting doses 

are bounding. Selection criteria are discussed for 

the radionuclide inventory, radionuclide screening 

analysis, TFF degradation analysis, DUST-MS 

releases, and groundwater analysis. 

The criteria for the selection and 

quantification of four radionuclide inventory and 

transport scenarios are presented. Section 7.3 

contains a sensitivity/uncertainty analysis based 

on several scenarios involving the sensitive 

parameter for the transport of radionuclides. 

Different inventories, sorption coefficients, (Kd

values) hydraulic diffusivity values, and 

infiltration rates are varied in numerical 

simulations for four different scenarios and the 

results are discussed in Section 7.4. The four 

scenarios include a best case, realistic case, 

conservative case, and worst case. 

The final portions of this section present the 

uncertainties in the radon, volatile radionuclides, 

and intruder analyses. A discussion of future work 

completes this section. 

7.1 Comparison of Results to 

Performance Objectives 

This PA assesses the risk to two populations: 

the general public and hypothetical intruders. The 

risk to the general public has been assessed 

through the analysis of several conservative 

scenarios that reflect site-specific conditions at the 

TFF. The risk to hypothetical intruders was 

assessed through the analysis of four conservative 

scenarios. Tables 7-1 and 7-2 provide a summary 

of the PA results. The PA results indicate that all 

of the performance objectives are met for the TFF 

closure for all times post-closure. 

7.1.1 Analysis of Airborne Emissions 

The conservative case inventory, presented in 

Section 2.3, was used for the evaluation of the 

airborne emissions. 

Table 7-1. Comparison of conservative results to the all-pathways, air emissions, and radon performance 

objectives. 

Performance Objective PA Result 

Airborne emissions excluding radon 

10 mrem/yr 0.51 mrem/yr 

Average annual radon flux 

20 pCi/m2

/s 0.39 pCi/m2

/s 

Protection of groundwatera

4 mrem/yr 0.77 mrem/yr 

All-pathwaysa

25 mrem/yr 1.86 mrem/yr 

a. The all-pathways and protection of groundwater doses were never exceeded for the 1 million-yr analysis. 
 

7-2

Table 7-2. Comparison of intruder results to the performance measures. 

PA Results 

Performance Objective 100 yr 200 yr 500 yr 1,000 yr 

Acute drilling scenario 

500 mrem 232 mrem 46.5 mrem 18.8 mrem 16.3 mrem 

Acute construction scenario 

500 mrem 0.80 mrem 0.13 mrem 0.05 mrem 0.05 mrem 

Chronic post-drilling scenario 

100 mrem/yr 91.1 mrem/yr 10.9 mrem/yr 2.3 mrem/yr 2.2 mrem/yr 

Chronic post-construction 

scenario 

100 mrem/yr 

26.1 mrem/yr 3.1 mrem/yr 0.64 mrem/yr 0.61 mrem/yr 

The annual dose from volatile releases 

predicted by CAP88-PC atmospheric transport 

runs resulted in a dose for 3

H releases of 

0.008 mrem/yr and 14C doses of 0.50 mrem/yr, 

with a total dose of 0.51 mrem/yr. These doses are 

well below the atmospheric pathway performance 

objective of 10 mrem/yr and also indicate that the 

atmospheric pathway will not contribute 

significantly to the all-pathways dose. This was a 

very conservative analysis because flux rates were 

based on simplified models and the receptor was 

located 100 m (300 ft) from the facility at all 

times (see Section 3.3.4). Also, the receptor would 

not be located near the TFF for the first 100 yr 

after closure because the site is assumed to be 

maintained and patrolled. 

7.1.2 Analysis of Radon Flux 

The conservative case inventory, presented in 

Section 2.3, was used for the evaluation of the 

radon flux rates. 

Calculations were performed for a onedimensional bare slab resting on the ground 

surface to determine the radon flux at the surface 

of the slab (Section 3). These calculations are 

highly conservative with respect to the actual 

performance of a grouted heel at the TFF because 

the slab is modeled in open air. This means that 

no credit is taken for the vault or tank. Also, the 

radium inventory at the end of the 1,000-yr 

analysis period is used for the source in the 

calculations, which is conservative for all times 

before the end of the performance period; 

therefore, more radon can diffuse out of the slab 

than would otherwise be expected. 

For 222Rn (the dominant isotope in terms of 

atomic fraction), the flux of radon at the surface of 

the slab was 0.39 pCi/m2

/s, which is more than 

one order of magnitude below the performance 

objective of 20 pCi/m2

/s. Given the conservative 

nature of the calculation, there are two reasons 

that this value is so low: cementitious materials do 

not transport radon very well and the radium 

content of the grouted heel is relatively low. 

Although the bare slab model could not be shown 

government to meet the standard for the shorterlived isotopes of radon (220Rn, half-life of 55.6 s 

and 219Rn, half-life of 3.96 s), Section 3.3.3 shows 

that 0.5 m (2 ft) of overburden is capable of 

reducing these radon fluxes by at least a factor of 

1096. Therefore, the shorter-lived isotopes are 

quite incapable of penetrating even a modest 

overburden. Because of their shorter half-lives, 

the isotopes do not survive long enough to 

traverse the barrier. 

7.1.3 Analysis of Groundwater Pathway 

Doses 

The groundwater protection and all-pathways 

dose analyses were based on the conservative case 

inventory presented in Section 2.3. The results of 

the analyses presented in Sections 4 and 6 show 

that the performance objectives were not 

exceeded. 
7-3

According to the results presented in 

Section 4, the realized doses to a receptor 

individual of 0.77 mrem/yr at 890 yr for the 

groundwater protection scenario (i.e., drinking 

water) did not exceed the performance objective 

of 4 mrem/yr for 1,000 yr after project 

completion. The long-term analysis in Section 4 

also indicates that peak doses do not exceed the 

performance objective during the period of 

closure to one million years post-closure. 

According to the results presented in 

Section 4, the realized doses to a receptor 

individual of 1.4 mrem/yr at 890 yr for the allpathways scenario did not exceed the performance 

objective of 25 mrem/yr for 1,000 yr after project 

completion. The long-term analysis in Section 4 

also indicates that peak doses do not exceed the 

performance objective during the period of 

closure to one million years post-closure. 

A sensitivity/uncertainty analysis is presented 

in Section 7.3 for the groundwater pathway. The 

results of the conservative dose analysis are 

compared to the results of the uncertainty analysis 

to ensure that the projected doses are bounding. 

7.1.4 Analysis of Chronic Intruder PostDrilling Scenario 

The conservative case inventory, presented in 

Section 2.3, was used for the evaluation of the 

intruder doses. 

The maximum dose after institutional control 

(i.e., 100 yr) for the acute intruder-drilling 

scenario was 232 mrem. The major radionuclide 

contributors to the total acute intruder dose were 

137Cs/137mBa with 188 mrem, 238Pu with 

15.0 mrem, 90Sr/90Y with 8.64 mrem, 239Pu with 

15.3 mrem, 241Am with 5.1 mrem, and 240Pu with 

2.8 mrem. The acute drilling scenario dose results 

are less than the performance objective of 

500 mrem for all times post-institutional control. 

The maximum dose post-institutional control 

(i.e., 100 yr) for the acute intruder-construction 

scenario was 0.80 mrem. The major radionuclide 

contributors to the total acute intruder dose were 

137Cs/137mBa with 0.72 mrem, 94Nb with 

0.04 mrem, 90Sr/90Y with 0.02 mrem, and 238Pu 

with 0.01 mrem. The acute intruder-construction 

scenario dose results are less than the performance 

objective of 500 mrem for all times post-closure.

The maximum dose for the chronic intruder 

post-drilling scenario was 91.1 mrem/yr at 100 yr 

post-closure. The major radionuclide contributors 

to the total chronic intruder post-drilling scenario 

dose were 90Sr/90Y with 51.5 mrem/yr, and 

137Cs/137mBa with 36.5 mrem/yr. The chronic 

intruder post-drilling scenario dose results do not 

exceed the performance objective of 100 mrem/yr 

during the compliance period. 

The maximum dose was 26.1 mrem/yr for the 

chronic intruder post-construction scenario at 

100 yr post-closure. The major radionuclide 

contributors to the total chronic intruder dose 

were 90Sr/90Y with 15.2 mrem/yr, 137Cs/137mBa 

with 10.1 mrem/yr, 94Nb with 0.58 mrem/yr, and 

238Pu with 0.15 mrem/yr. The chronic intruder 

post-construction scenario dose results do not 

exceed the performance objective of 100 mrem/yr 

for the 1,000-yr compliance period

Uncertainty exists in the state of the concrete 

systems over time. In this analysis, credit was not 

taken for the concrete vault providing a barrier to 

intrusion. Oztunali and Roles (1986) state that “if 

resistance is encountered during drilling (as would 

be the case if the drill bit hit solid rocks or intact 

metal), the crew would simply move a few yards 

horizontally and drill a new hole. Drilling 

equipment and bits that can drill through solid 

rock and intact metal are available; however, they 

are more expensive than the equipment and bits 

that are normally required and used for water well

installation. Similarly, it is anticipated that 

standard drill bits would have difficulty 

penetrating reinforced concrete structures, 

although little difficulty is expected for waste 

stabilized using grout backfill.” Taking credit for 

both the reinforced concrete vault and the 

stainless-steel tank for additional years beyond 

100 yr would reduce the intruder doses further 

because the dose is due to shorter-lived 

radionuclides. 

The uncertainty in the results of the PA 

intruder analyses and potential effects on the 

projected doses is discussed in Section 7.10. 
7-4

7.2 Use of Performance 

Assessment Results 

The analyses presented in this PA were based 

on conservative assumptions and a conservative 

inventory. Based on the conservative analyses, all 

of the PA performance objectives were met. 

The performance objective budget was 

allocated to the entire TFF closure. This is 

considered appropriate because all tanks to be 

closed are located within close proximity to one 

another. In other words, closure of facilities 

located in separate hydraulic flow regimes does 

not apply to the TFF closure. As demonstrated in 

the Savannah River Site Tank Closure study,a

several facilities were considered during closure 

that were not within the same groundwater 

transport segments. The TFF is located within a 

flow regime in which all contaminants are 

considered within the specified groundwater 

transport segment modeled in this PA. Therefore, 

different performance objective budget allocations 

for each tank are not necessary. 

The two-dimensional groundwater transport 

model considered a 1-m (3-ft) slice through the 

major axis of two tanks in a north-south cross 

section. This ensures that the maximum unit area 

of waste is input in the model and the 

downgradient groundwater concentrations are 

representative of the centroid of the contaminant 

plume that would be realized in a threedimensional model. 

The sensitivity/uncertainty analysis presented 

in Sections 7.3 through 7.11 provide additional 

discussion of the PA results in light of the 

uncertainties in the analyses. 

 

a. Savannah River Site, 2000, Industrial Wastewater Closure 

Plan for F- and H-Area High-Level Waste Tank Systems, 

Preliminary Draft Revision 2, March 1. 

7.3 Groundwater Pathway 

Sensitivity/Uncertainty 

Analysis 

Several models and assumptions were used to 

assess the performance of the TFF, which results 

in several areas of the PA contributing to the 

uncertainty in the final results. Important areas of 

the PA analyses are discussed below and the 

sensitivity and uncertainty in the assumptions and 

parameter values are presented. A quantitative 

sensitivity/uncertainty analysis is provided for the 

groundwater pathway. 

Four scenarios were developed for the 

groundwater pathway: 1) worst case, 2) 

conservative case, 3) realistic case, and 4) best 

case. The development of these four scenarios, 

with varying parameter values, was suggested by 

NRC. The four scenarios provide a means of 

evaluating the sensitivity and uncertainty of the 

analysis presented in Section 4. The bounding 

analysis presented in Section 4 was based upon 

the conservative scenario and associated 

parameter values. The parameter values chosen 

for each scenario are presented in Table 7-3. The 

following sections provide discussions on the 

selection of these parameter values. 

7.3.1 Radionuclide Inventories 

As recommended by NRC in the Request for 

Additional Information dated June 5, 2002,b

 the 

proposed changes for evaluating the TFF 

radionuclide inventory includes best, realistic, 

conservative, and worst-case scenarios. NRC 

suggested that four scenarios be developed to 

include radionuclide inventory and transport 

parameters. NRC recommended “best case, 

expected behavior, a reasonably conservative 

case, and a worst case.”

 

b. Letter from Thomas H. Essig, Chief—Environmental and 

Performance Assessment Branch, Division of Waste 

Management, Office of Nuclear Safety and Safeguards, NRC, 

Joel T. Case, Director—INTEC Waste Programs, DOE-ID, 

2002, “Request for Additional Information on the Idaho 

National Engineering and Environmental Laboratory Draft 

Waste Incidental to Reprocessing Determination for the Tank 

Farm Facility Residuals (DOE/ID-10777),” June 5. 
7-5

Table 7-3. Overview of the parameter values for the sensitivity/uncertainty analysis. 

Nuclide Best 

Scenario Realistic Scenario 

Conservative 

Scenario 

Worst-Case 

Scenario 

Solid Radionuclide 

Inventory 

50% reduction 

from worst case 

25% reduction 

from worst case 

10% reduction 129I 

and 99Tc inventories 

reduced 

Depicts sodiumbearing waste 

(undiluted tankheel residual) 

Liquid Radionuclide 

Inventory 

95% reduction 

from worst case 

80% reduction 

from worst case 

50% reduction 129I 

and 99Tc inventories 

reduced 

Depicts sodiumbearing waste 

(undiluted tankheel residual) 

Infiltration 1.1 cm/yr 1.1 cm/yr 4.1 cm/yr 12.4 cm/yr 

Sr 0.006 0.006 0.003 0.001 

Tc 5 5 2.5 1 

I 0.03 0.03 0.008 0.002 

Grout Sorption 

Coefficients 

(m3

/kg) 

C 10 10 5.0 1.0 

Unsaturated Zone 

Longitudinal 

Dispersivities (m) 

0.52 (sediment) 

3.36 (basalt) 

0.52 (sediment) 

3.36 (basalt) 

0.29 (sediment) 

1.85 (basalt) 

0.052 (sediment) 

0.34 (basalt) 

Unsaturated Zone 

Transverse Dispersivities 

(m) 

0.26 (sediment) 

1.7 (basalt) 

0.26 (sediment) 

1.7 (basalt) 

0.14 (sediment) 

0.94 (basalt) 

0.026 (sediment) 

0.17 (basalt) 

Sr 24 24 18 12 

Tc 0.1 0.1 0.01 0 

I 5 5 0.1 0.01 

Interbed 

Sediment 

Sorption 

Coefficients 

(mL/g) C 20 20 10 2 

Sr 13 13 6 1 

Tc 0.24 0.24 0.01 0 

I 1 1 0.1 0 

Basalt Sorption 

Coefficients 

(mL/g) 

C 7.1 7.1 5.0 1.7 

For conservatism, a conservative waste 

inventory was used to bound the worst condition 

that 100% sodium-bearing waste remained in the 

tank after emptying it to expected levels. The 

worst condition implies higher waste 

concentrations in the tanks than shown in recent 

sampling and historical data and shows the 

impacts if no tank cleaning was done. 

The following discussions provide an 

overview of the development of the four inventory 

scenarios. A complete discussion on the 

development of the inventories is provided in 

Appendix A. 

The sand pad inventory was held constant 

throughout the sensitivity/uncertainty analysis 

because the sand pad is not proposed to be washed 

like the tanks and piping. 

7.3.1.1 Worst-Case Inventory. The worstcase radionuclide inventory initially was prepared 

to predict the greatest waste concentrations and 

masses that could remain at closure; it represents 

an upper estimate based on analytical data, 

numerical modeling, and knowledge of INTEC 
7-6

operations (See Appendix F). This case considers 

that tank internal surfaces have been cleaned and 

the residual waste has been pumped out until there 

is about a 1.25-in. waste heel remaining in the 

tank.

The piping inventory for the worst-case 

scenario was based upon 15.5 kg of solid waste in 

the piping. The total piping inventory was 

assumed to be placed where the 11 tanks are 

located, resulting in 1/6 of the piping inventory 

being placed in the modeling domain for the twodimensional groundwater model domain. 

7.3.1.2 Conservative Inventory. The 

conservative case inventory differs from the worst 

case because it is assumed that limited tank 

cleaning has occurred. The basic difference is the 

liquid radionuclide concentration is assumed to be 

reduced by 50% and solids mass is assumed to be 

reduced by 10%. Like the worst case, this case 

considers that tank internal surfaces have been 

cleaned and the residual waste has been pumped 

out until there is about a 1.25-in waste heel 

remaining in the tank. The conservative case was 

prepared to predict the greatest waste 

concentrations and volumes that could remain at 

closure with only limited cleaning and represents a 

very conservative estimate based on analytical 

data, numerical modeling, and knowledge of 

INTEC operations.

The piping inventory for the conservative case 

was based upon 15.5 kg of solid waste in the 

piping, reduced by 10%. 

7.3.1.3 Realistic Inventory. The realistic 

case inventory assumes that significant tank 

cleaning has occurred. The liquid radionuclide 

concentration is assumed to be reduced by 80% 

and solids mass is assumed to be reduced by 25% 

from the worst case. The realistic case was 

prepared to predict waste volumes that would 

remain at closure with significant cleaning and 

represents a conservatively realistic estimate based 

on analytical data, numerical modeling, and 

knowledge of INTEC operations. This scenario 

does not include the cleaning efficiencies shown in 

the mock-up of the tank cleaning but 

conservatively estimates the observed cleaning 

efficiencies of Tank WM-182. This case considers 

that tank internal surfaces have been cleaned and 

the residual waste has been pumped out until there 

is about a 1-in. waste heel remaining in the tank.

The piping inventory for the realistic case was 

based upon 15.5 kg of solid waste in the piping, 

reduced by 80%. 

7.3.1.4 Best-Case Inventory. The bestcase inventory of radionuclides differs 

significantly from the worst-case scenario. This 

scenario assumes a 95% reduction in liquid waste 

radionuclide concentrations and a 50% reduction 

in solid residual from the worst case. This case 

considers that tank internal surfaces have been 

cleaned and the residual waste has been pumped 

out until there is slightly over a 1-in. waste heel 

remaining in the tank. The best case was prepared 

to predict waste volumes that would remain at 

closure with thorough cleaning and represents a 

realistic but optimistic estimate based on analytical 

data, numerical modeling, and knowledge of 

INTEC operations. This scenario assumes that 

solid residual removal is about the same as that 

achieved in tank mock-up testing (INEEL 2001c). 

This conservatism allows for variability in the 

tank-to-tank solid inventory and structural and 

component differences; it also accounts for any 

reduced efficiency from cold mock-up conditions 

to field conditions.

The piping inventory for the best case was 

based upon 15.5 kg of solid waste in the piping, 

reduced by 95%. 

7.3.2 Sorption Coefficients 

A key area of uncertainty in the groundwater 

transport analyses is the sorption coefficients (Kd

values). Kd values have been found to range over 

several orders of magnitude for selected 

radionuclides. 

The following sections discuss the selection of 

distribution coefficients for the four sensitivity/ 

uncertainty scenarios. 

7.3.2.1 Sediment and Basalt Kd Values. 

Knowledge of the sorptive properties of 

contaminants is key to understanding contaminant 

movement at the TFF. The subsurface comprises 
7-7

surficial sediments, sedimentary interbeds, and 

basalt. There is limited site-specific adsorption 

information for contaminants in the subsurface 

environment at the INEEL (Del Debbio and 

Thomas 1989; Schmalz 1972). Distribution 

coefficients are available for cobalt, chromium, 

strontium, cesium, cadmium, mercury, and 

selenium for alluvium. For interbed sediments and 

basalt, distribution coefficients are available for 

cadmium, mercury, selenium, and strontium. 

Limited availability of site-specific adsorption 

information for other radionuclides for sediments 

and basalts at the INEEL has resulted in the use of 

adsorption parameters measured for sediments and 

basalts from other sites. Table 7-4 summarizes the 

measured Kd values for the INEEL. 

Several investigators have published 

compendia of soil and sediment distribution 

coefficient data (Baes and Sharp 1983; Coughtrey, 

Jackson, and Thorne 1985; Sheppard and Thibault 

1990). The most thorough of these compendia is 

Sheppard and Thibault (1990), which contains a 

breakout of the distribution coefficients by major 

soil types (i.e., sand, loam, clay, and organic). Past 

compendia grouped all soil types together. 

Distribution coefficient data for basalts are not 

as readily available as data for soils and sediments. 

The main source of Kd values is basalts from the 

Nuclear Energy Agencies sorption database 

(Ticknor and Ruegger 1989). These basalt sorption 

values are provided in Table 7-5. 

In addition, a survey of the literature for the 

WAG 3 modeling study (Rodriquez et al. 1997) is 

provided in Table 7-6. 

Kd values are considered sensitive parameters 

in predictive modeling of radionuclide transport in 

the subsurface. There are a limited number of 

actual measurements available for the 

radionuclides that are specific to the INEEL soils 

and basalt as indicated by Table 7-4. 

Consequently, it was necessary to survey the 

literature for similar sites where actual 

radionuclide sorption studies have been conducted. 

Table 7-3 provides a list of Kd values for key 

radionuclides used in the predictive transport 

modeling. The rationale for the selection criteria is 

presented in Table 7-7. The sensitivity/uncertainty 

analysis consisted of the use of three separate 

combinations of the sorption coefficients: (1) 

worst-case transport, (2) conservative transport, 

and (3) realistic/best transport. The realistic/best 

scenarios used the same Kd values. This method 

provided a means of comparing the conservative 

case analysis, presented in Sections 3 and 4, with 

the range of values presented in the literature. 

The appropriateness of using these Kd values 

at the INEEL is difficult to assess. The 

mineralogical differences have not been addressed, 

and the secondary minerals (e.g., clays and iron 

oxides) have not been quantified. The secondary 

minerals form in the fractures and may have a 

major impact on the adsorption of many 

radionuclides. Also, it is unclear if equilibrium is 

maintained during flow through the unsaturated 

zone. If local equilibrium is not valid, then the Kd

value will overestimate the retardation of 

contaminants. 

Column studies conducted on basalts at 

various moisture contents from saturation to 20% 

consistently resulted in essentially the same Kd

values (Porro, Newman, and Dunnivant 2000). 

This evidence indicates that Kd values for the 

saturated zone are applicable to the unsaturated 

zone. While the reliability of Kd values obtained 

from batch tests is still being debated, using 

saturated Kd values for unsaturated transport 

simulation is not a specific concern. 

The modeling study conducted for the Remedial 

Investigation/Baseline Risk Assessment (RI/BRA) 

for INTEC (Rodriguez et al. 1997) used the 

sediment distribution coefficient for the alluvium 

and the interbed sediments, a zero distribution 

coefficient for the vadose zone basalts, and 1/25 of 

the sediment distribution coefficient for the aquifer 

basalts. The vadose zone basalts were assumed to 

have a zero distribution coefficient in the RI/BRA 

based on the assumption that the water moves 

more quickly through vertical fractures in the 

vadose zone. This allows very little time for the 

contaminant to sorb to the basalt. In the aquifer, 

the RI/BRA modeling assumed that the water 

moves more slowly through the basalt fractures 

and matrix and there is sufficient time for the 

contaminants to sorb in the basalts. 
7-8

Table 7-4. Summary of sorption coefficients measured at the INEEL. 

Kd Values (mL/g)a

Element Alluvium 

Interbed 

Sediment Basalt Reference 

Cd 4891-2864 10,115-8622 2319-785 Del Debbio and Thomas (1989) 

Co 56 Not measured Not measured Schmalz (1972) 

Cr 1.2 Not measured Not measured Schmalz (1972) 

Cs 950 Not measured Not measured Schmalz (1972) 

Hg 1921-236 673-72 87-9.2 Del Debbio and Thomas (1989) 

Se 63-5.8 17-4.9 3.4-0.29 Del Debbio and Thomas (1989) 

Sr 35-52 110-186 1.1-2.7 Del Debbio and Thomas (1989) 

24 Not measured Not measured Schmalz (1972) 

a. Kd values reported by Del Debbio and Thomas (1989) were measured using batch equilibrium sorption methods on crushed 

materials that were passed through a 2-mm sieve. 

Table 7-5. Basalt Kd values from the NEA database. 

Element 

 NEA Basalt Kd Ranges 

(mL/g) 

 I 5.0E−02 

 Sr 2.0 to 1000 

 Tc 1.5E−01 to 5.5 

Table 7-6. Sorption coefficient ranges identified in WAG 3. 

Element Sediment Kd (Range) mL/g Basalt Kd (Range) mL/g 

C 5 (2-20)b — 

I 0.1 (0.02-5) — 

Sr 60 (35-186)c,d 6 (1-13)c,d

Tc 0d

 0d

 

a. Source: Rodriquez et al. 1997. 

b. Range from NEA (Ticknor and Ruegger 1989) and Sheppard and Thibault (1990). 

c. Source: Newman et al., in preparation, Evaluation of the Mobility of Am, Cs, Co, Pu, Sr, and U through INEEL Basalt and 

Interbed Materials: Summary Report of the INEL/Clemson University Laboratory Studies, ER-WAG7-82, INEEL internal 

report. 

d. Source: Del Debbio and Thomas 1989. 
7-9

Table 7-7. Selection rationale for the range of Kd values (mL/g) used in the sensitivity analysis. 

Radionuclide 

Range for Kd

values (mL/g) 

used in 

transport model Selection Rationalea

C (sediment) 2–20 Default: Ref. #1 loam soil Kd = 20 

Site Values: WAG 3 Ref. #2 and #3 Kd = 2–20 

Literature Confirmation: Sand Ref. #1 Kd = 5 

C (basalt) 1.7–7.1 Site Values: WAG 3 Kd = not reported 

Literature Confirmation: The lowest reported Ref. #4 Kd = 1.7 

I (sediment) 0.01–5 Default: Ref .#1 loam soil Kd = 5 

Site Values: WAG 3 Ref. #2 Kd = 5 

Literature Confirmation: No referenced value as low as Kd = 0.01 

I (basalt) 0–1 Site Values: WAG 3 Ref. #2 Kd = 0 

Literature Confirmation: Ref. #3 Kd = 0.05, Ref. #1 for sand Kd = 1 

Sr (sediment) 12–24 Ref. #1 loam soil Kd = 20 

Site Values: WAG 3 Ref.#2 Kd = 12–24 

Literature Confirmation: Ref. #5 reports a range of 35–52 alluvium 

and 110–186 for interbed sediments 

Sr (basalt) 1–13 Site Values: WAG 3 Ref. #2 Kd = 1–13 

Literature Confirmation: Ref. #5 reports a range of 1.1–2.7 for 

crushed basalt column tests, Ref. #4 Kd = 2–1000 

Tc (sediment) 0–0.1 Default: Ref. #1 loam soil Kd = 0.1 

Site Values: WAG 3 Ref. #2 Kd = 0 

Literature Confirmation: Ref. #3 lowest measured sand Kd = 0.02 

Tc (basalt) 0–0.24 Site Values: WAG 3 Ref. #2 Kd = 0 

Literature Confirmation: Ref. #5 measured Kd = 0.016 for crushed 

basalt column tests; Ref. #3 Kd = 0.15–5.5 

a. References: 

1. Sheppard and Thibault 1990. 

2. Rodriquez et al. 1997. 

3. NEA database (Ticknor and Ruegger 1989). 

4. Thibault, Sheppard, and Smith 1990. 

5. Del Debbio and Thomas 1989. 
7-10

The RI/BRA modeling also based the aquifer 

basalt distribution coefficients on an evaluation of 

the ratio of the distribution coefficient for 

sediments to that for basalts from the Del Debbio 

and Thomas (1989) report. Here they found the 

ratio to the 1/25. The Del Debbio and Thomas 

(1989) report, however, provides conflicting 

information on the differences in sediment and 

basalt distribution coefficients. This report 

provides for higher distribution coefficients in the 

basalts (based on breakthrough curve studies) in 

comparison to the batch distribution coefficient 

study. Because of this, the appropriateness of the 

assignment of vadose and aquifer basalt 

distribution coefficient values is difficult to 

assess. Therefore, the distribution coefficients for 

basalts were treated as uncertain as well as the 

sediment distribution coefficients. 

7.3.2.1.1 Strontium—The basalt 

values were based on the range presented in 

Table 7-7 with the worst-case scenario being 

assigned the lowest reported values in the range 

and the best/realistic case being assigned the 

highest value in the reported range. The 

conservative values were assigned the mid-point 

of the reported basalt Kd range.

The sediment (i.e., interbed sediments in the 

unsaturated zone) were assigned Kd values based 

upon the modeling conducted for WAG 3 

CERCLA groundwater modeling at INTEC 

(Rodriquez et al., 1997). This modeling showed 

that a Kd value of 12 for the sediment resulted in 

90Sr groundwater concentrations higher then 

measured. A sediment value of 24 also was 

investigated and showed a closer match of the 

modeled concentrations with the measured 

concentrations. Therefore, the best/realistic 

scenario was assigned a Kd value of 24 for the 

interbed sediments, while the worst case was 

assigned a value of 12. The conservative case was 

assigned the mid-point value of the range. 

7.3.2.1.2 Carbon—The basalt values 

were based on the range presented in Table 7-7. 

The worst-case scenario was assigned the lowest 

reported value, while the best/realistic scenario 

was assigned the high value for the range. 

The sediment Kd values were taken from the 

range reported in Table 7-7 with the worst-case 

scenario being assigned the low value of the range 

and the best/realistic scenario being assigned the 

high value of the range. The conservative case 

was assigned a value of 5 mL/g, which was closer 

to the low end of the reported range. 

As noted above, sand Kd values were selected 

for 14C for transport simulations in the basalts 

where site-specific basalt Kd values were not 

available (Table 7-4). The published sand Kd

values were obtained from Sheppard and Thibault 

(1990). It was assumed that the lower sand Kd

values are more representative of the sorptive 

behavior of 14C in basalt. This assumption is based 

on the coating of weathered mineral or chemical 

precipitates on sand grains and the irregular 

fracture surfaces that can react with the 

radionucludes. While the surface area to the voidspace ratio for fractured media is less than a 

porous media (thereby reducing the sorptive 

capability for a fractured media), the presence of 

abundant iron oxides in basalt fractures capable of 

sorbing radionuclides offset the surface area 

effects. The degree of this offset is uncertain and 

difficult to apply to natural systems. 

This approach allows a degree of retardation 

of 14C in the basalts. The lower sand Kd values, 

compared to the higher Kd values for finer-grain 

sediment, are conservative on a relative basis. 

However, there is an acknowledged degree of 

uncertainty because of the lack of direct Kd

measurements available for 14C in basalt. 

7.3.2.1.3 Technetium—The basalt 

value for the worst-case scenario was based on the 

value of zero in Table 7-7. The best/realistic case 

was assigned a value of 0.24, which is in the low 

end of the range. This was done to consider the 

fact that Tc has been found to be mobile at the 

INEEL Site. The conservative value was set to a 

very low value of 0.01, again due to reports that 

Tc is mobile at the INEEL.

The sediment Kd value for the worst case was 

taken as zero, while the realistic case was 

assigned a value of 0.1 based on values in 

Table 7-7 for loam soil. The conservative value 
7-11

was assigned a value one order of magnitude less 

than the loam soil value. 

7.3.2.1.4 Iodine—The basalt value for 

the worst case was assigned a value of zero, while 

the best/realistic case was assigned a value of 1.0 

based on a reasonable upper end for the values 

reported in Table 7-7. The sediment Kd value for 

the worst case was assigned a value of 0.01, while 

the best/realistic case was assigned a value of 5.0. 

The conservative case was assigned a value of 0.1.

7.3.2.2 Grout Kd Values. The grout Kd

values were evaluated for the four groundwater 

pathway scenarios. A review of the Bradbury and 

Sarott (1995) study on Sorption Databases for the 

Cementitious Near-Field of a LLW Repository for 

Performance Assessment indicates that the Kd

values presented in their study were based on the 

conservative end of the possible choices (i.e., 

lower Kd values). Therefore, a combination of 

grout Kd values were chosen based on Bradbury 

and Sarott (1995), Allard (1985), and other studies 

specific to the elements of interest.

Concrete Kd values for reducing conditions 

have been chosen. The concrete is expected to 

exhibit strongly reducing conditions (Eh from –

300 to –500 mV) as do most concrete systems. In 

addition, the closure system will also consist of a 

mix of concrete and fly ash, slag, or other 

substances to ensure reducing conditions in the 

grout. 

The concrete Kd values were chosen from 

Region II of the Bradbury and Sarott study 

(1995). In Region II, the pore water composition 

is dominated by portlandite Ca(OH)2, which fixes 

the pH at 12.5. The portlandite will be slowly 

removed by groundwater flow but the quantities 

contained in the cement are so large that this 

phase buffers the system over very long periods of 

time. 

Bradbury and Sarott (1995) also discuss a 

Region III (Figure 1) for concrete systems, where 

the removal of Ca(OH)2 has become significant 

and the pH falls continuously. The calciumsilicate-hydrate gel is no longer stable and begins 

to dissolve incongruently. However, investigation 

into the original paper on these concrete regions 

by Atkinson, Everitt, and Guppy (1989) reveals 

that concrete is not expected to reach this state 

until 100,000 yr. Since the PA compliance period 

is only 1,000 yr, the concrete Kd values are not 

expected to change over this period of time and 

therefore are held constant. This assumption is 

supported by the concrete degradation analysis 

presented in Appendix E. Graphs of the hydraulic 

conductivity for the tank/grout show no increase 

over the 1,000-yr time period. The concrete 

degradation sensitivity analysis also shows that 

the tank/grout hydraulic conductivity does not 

increase until 800 yr for the least conservative 

case with only a one order of magnitude change in 

the hydraulic conductivity. Thus, the amount of 

water available for loss of Ca(OH)2 is still 

minimal. For conservatism, the PA assumes that 

the tank fails at 500 yr; however, the degradation 

analysis indicates the tank/grout life is much 

longer. In addition, graphs of carbonation 

degradation and Ca(OH)2 leaching do not show 

any significant changes during the 1,000-yr 

compliance period for loss of Ca(OH)2. 

The selection of concrete Kd values for the 

three groundwater pathway scenarios are 

presented in Table 7-3. The following discussion 

provides justification for each set of sorption Kd

values for each element shown in Table 7-3. 

7.3.2.2.1 Strontium—Ewart, Terry, 

and Williams (1985) report Sr distribution ratios 

for hardened cement paste and concrete between 

0.001 and 0.004 m3

/kg, with very little 

dependency on concentration. Atkinson and 

Nickerson (1988) summarized the results from 

different types of tests and give a best-estimate 

range for Sr of 0.003 to 0.006 m3

/kg. Due to the 

small difference in Kd values reported for Sr, the 

worst-case scenario was assigned the low value of 

0.001 m3

/kg, and the value was scaled for the 

other scenarios to a high value of 0.006 m3

/kg for 

the best/realistic scenario.

7.3.2.2.2 Technetium—Sorption 

data for Tc on cementitious materials are sparse. 

Under reducing conditions, Tc is present as 

hydrolyzed Tc+4 species. In some recent work, 

using Tc+4 at trace levels, distribution ratios of 

5 m3

/kg have been reported (Bayliss et al. 1991). 

Tc may be expected to sorb strongly under 
7-12

reducing conditions at a high pH. Bradbury & 

Sarott (1995) chose a conservative value (i.e., low 

value) of 1 m3

/kg for reducing conditions. 

Therefore, for the TFF PA, the conservative value 

of 1 m3

/kg was chosen for the worst-case scenario, 

with the Kd values being scaled up to the 

measured value of 5 m3

/kg for the best/realistic Kd

value.

7.3.2.2.3 Iodine—Iodine is assumed 

to be present as I-

 under oxidizing and reducing 

conditions (Bradbury and Sarott 1995). Many 

studies on the sorption of I-

 on cement paste at a 

high pH indicate that its sorption is low but finite. 

Bradbury and Sarott (1995) selected a 

conservative value (i.e., lower value) of 

0.002 m3

/kg for iodine in reducing 

trichloroethylene (TCE) concrete systems. Allard, 

Persson, and Terstenfelt (1985) recommended a 

Kd value of 0.03 m3

/kg for I- in concrete systems. 

Therefore, for the TFF PA, the conservative value 

of 0.002 m3

/kg was chosen for the worst-case 

scenario, with the Kd value being scaled up to the 

recommend value of 0.03 m3

/kg by Allard, 

Persson, and Terstenfelt (1985) for the 

best/realistic case.

7.3.2.2.4 Carbon—Allard, Persson, 

and Terstenfelt (1985) recommend a Kd value for 

C of 5 m3

/kg. Studies by Allard, Torstenfelt, and 

Andersson (1981) and Bayliss et al. (1988) have 

investigated the sorption of 14C as 14CO3

2- in 

cement/concrete systems. Generally, very high 

sorption values (approximately 10 m3

/kg) have 

been reported. Therefore, for the TFF PA, the 

value of 10 m3

/kg has been chosen for the 

best/realistic case and the scaled-down value of 

1 m3

/kg for the worst-case scenario.

7.3.3 Hydraulic Dispersivity Variations 

Dispersivity is a scale-dependent term used in 

the advective-dispersion equation to compensate 

for the variation in transport of solute particles. 

Ideally, the transport of solute particles could be 

described by the processes of advection and 

diffusion if it were possible to define the 

velocities throughout the porous media. Because 

this is impractical, averaging the groundwater 

velocities over some representative scale results in 

a scale-dependency of the dispersion term. A rule 

of thumb for estimating longitudinal dispersivity 

is 1/10 of the transport distance and a transverse 

dispersivity of one-half of the longitudinal value 

(Mills et al. 1985). Mills et al. (1985) also state 

that the unsaturated zone dispersivities are 

normally observed to be lower than the 1/10 rule. 

Inverse modeling of the INEEL Test Area North’s 

TCE and tritium plume suggests the fracture 

basalt has a smaller value than rule-of-thumb 

dispersivity values. 

Pickens and Grisak (1981) provide a 

comparison of scale-dependent dispersion based 

on field and laboratory studies. The longitudinal 

dispersivity ranges from 12 to 61 m based on 

computer modeling studies in granular geologic 

material associated with larger contaminated 

zones. Results of field tracer tests yield a range in 

longitudinal dispersivities of 0.012 to 15.2 m. 

Small-scale laboratory tests using repacked 

granular material yield a range in longitudinal 

dispersivities on the order of 0.01 to 1 cm. Work 

conducted by Sudicky and Cherry (1979) and Lee, 

Cherry, and Pickens (1980) showed that point 

sampling compared to samples from large screen 

intervals was a large factor in dispersivities 

reported in the literature. These researchers 

showed a range of longitudinal dispersivities 

based on point sampling methods of 0.01 to 

0.22 m. 

In evaluating the best estimate of dispersivity 

values for the unsaturated zone modeling, several 

factors were incorporated into the analysis. The 

transport scale is on the order of 130 m for 

transport through the unsaturated zone to the 

underlying water table. There are 20 separate 

zones that are assigned a specific hydraulic 

conductivity value that differs from the adjacent 

zones. The variations in hydraulic conductivity 

inherently impart dispersion on the transport of 

solutes. The scale of transport in the individual 

geologic units or zones is on the order of a few to 

tens of meters. As shown in Figure 7-1, the scale 

of transport distances for the individual 

stratigraphic units results in a range of 

longitudinal dispersivity of 0.01 to 1.0 m. The 

INEEL WAG 3 modeling for the INTEC facility, 

where the TFF is located, used a dispersivity 

value of 5 m for their unsaturated zone modeling 

(Rodriquez et al. 1997). 
7-13

Figure 7-1. Longitudinal dispersivity graph (Mills et al. 1985). 

The results of an extensive review of the 

literature for measured dispersivity values in 

porous and fractured material are presented in 

Table 7-8. The range in measured dispersivity 

values is significantly less than the calculated 

site values for the Snake River Plain Aquifer. In 

addition, the measurement scale for literature 

values reported in Table 7-8 are closer to the 

scale of transport for the individual geologic 

units in the unsaturated zone underlying the site. 

Consequently, the best/realistic, conservative, 

and worst-case scenarios were chosen from the 

values listed below. 

For the realistic scenario, the highest values 

presented in Table 7-8 were selected. These 

values represented the highest measured or 

estimated values available in the literature. For 

the best/realistic scenarios, the interbed 

sediments will be assigned a value of 0.52 m for 

the longitudinal dispersivity and one-half that 

value, or 0.26 m for the transverse dispersivity. 

For basalt portions of the unsaturated zone, the 

selected values were 3.36 m for the longitudinal 

dispersivity and 1.7 m for the transverse 

dispersivity for the best/ realistic scenarios. The 

worst-case scenario was assigned 1/10 of the 

dispersivity values used for the best/realistic 

case. The conservative case was assigned values 

that are in the mid-point of the ranges of 

dispersivities used for the best/realistic and 

worst-case scenarios. 

7.3.4 Infiltration Variations 

A wide range in the infiltration rates at the 

site is reported in the literature and site-specific 

studies. These values range from a low of 

0.16 in./yr (0.4 cm/yr) to a high of 4.9 in./yr 

(12.4 cm/yr). Most of the reported values in the 

literature are estimates based on the amount of 

precipitation and best-guess estimates of the 

evapotranspiration rates for the area. 

One of the few actual field measurements 

conducted at the site is reported by Cecil et al. 

(1992). By measuring tritium and 36Cl profiles in 

the soil, test results yielded a range of 0.4 to 

1.1 cm/yr (0.2 to 0.43 in./yr) infiltration. 

A detailed investigation of water infiltration 

rates was conducted at the CFA Landfill by 
7-14

Table 7-8. Dispersivity values measured or calculated and reported in the literature (in meters). Boateng and Cawlfield (1999) Bresler and Dagan (1983) Broermann et al. (1997) Hinsby et al. (1996) Mallants et al. (2000) Pickens and Grisak (1981) Pistre et al. (2002) Pohll et al. (1999) Russo and Bresler (1981) 

Sudicky, Cherry, and Frind (1983) 

Unsaturated 

soil 

αL=0.005

αT=0.001 

αL=0.03 αL=0.08–

0.52 

Saturated soil αL=0.67–

6.51 

αL=0.09 

αT=0.0085 

αL=0.7 αL=0.01–

0.08 

αT=0.005–

0.03 

Unsaturated 

Fractures 

 

Saturated 

Fractures 

 αL=0 αL=0.01–

3.36 

αL=1 

αT=1 

 

Miller et al. (1990). The CFA Landfill is 

covered by an earthen-based material from the 

surrounding area. Based on soil analysis, 

precipitation, and evapotranspiration estimates, a 

range of infiltration rates from 1.0 in. (2.5 cm) to 

1.6 in./yr (4.1 cm/yr) was reported. This 

measured value is the same infiltration rate used 

in the WAG 3 modeling analysis (Rodriquez et 

al. 1997) for the Tank Farm area. 

The measured value of 1.6 in./yr (4.1 cm/yr) 

for the CFA Landfill is considered an 

appropriate value for the conservative case since 

the TFF area will be covered with an engineered 

barrier under the CERCLA program. Therefore, 

this infiltration rate was considered for the 

realistic and base-case modeling estimates. 

Three infiltration rates were considered in 

the uncertainty analysis, 12.4 cm/yr (4.88 in./yr) 

as the highest rate for the worst-case scenario, 

4.1 cm/yr (1.6 in./yr) as the base-case infiltration 

for the conservative scenario, and 1.1 cm/yr 

(0.43 in./yr)as the realistic/best case infiltration. 

The low infiltration rate of 1.1 cm/yr 

(0.43 in./yr) is considered applicable to a wellengineered landfill cover that is required for the 

TFF. 

7.3.5 Time of Degradation 

Degradation analyses were conducted for 

the TFF closure system, which includes the 

vaults, tanks, grout, and piping. The degradation 

analyses were conducted to provide support for 

the PA analyses that assumed degradation step 

changes of 100 yr for the vault and grout 

between the vault and tanks, 500 yr for the tanks 

and grout inside of the tanks, and 500 yr for the 

piping. 

Table 7-9 gives a summary of the degradation 

analyses using the assumption that a 50% loss of 

grout is the point of degradation failure of the 

system hydraulically. Table 7-9 shows the 

degradation times assumed in the PA, those for 

the base-case degradation analysis, and also the 

minimum and maximum degradation times from 

the sensitivity analysis (see Appendix E). 

As indicated by the data in Table 7-9, the 

assumptions for degradation step changes in the 

PA are well within the ranges predicted by the 

degradation analysis. The uncertainty in the 

degradation of the TFF system components is 

large, as indicated in Table 7-9. However, the 

degradation steps assumed for the PA are equal 

to or less than the minimum failure times 

predicted by the degradation analysis. The data 

indicate that additional credit for a system 

component failure times could be taken. 
7-15

Table 7-9. Summary of the degradation analysis, showing the time of failure for each component. 

TFF System 

Component 

PA Failure Time 

Assumption 

(yr) 

Base Case 

Degradation 

Failure Time 

(yr) 

Minimum 

Degradation 

Failure Time 

(yr) 

Maximum 

Degradation 

Failure Time 

(yr) 

Vault 100 175 100 > 10,000 

Grout (between vault 

and tank) 

100 3,500 500 > 10,000 

Pipinga

 500 8,000 1,750 > 10,000 

Tank and grout inside 

tank 

500 8,000 1,750 > 10,000 

 

a. Based on the degradation analysis, taking credit for the stainless-steel piping, and thus being equal to the tank degradation 

time. 

However, to ensure that the analysis was 

bounding, conservative failure times were used 

for the analyses. 

Corrosion rates from Palmer et al. (1999) are 

used to estimate the vault failure time from the 

expansion-corrosion reaction of the reinforcing 

steel and the times of the tank and the piping 

failure from corrosion. The corrosion rates are 

based on the maximum uniform corrosion rate 

observed in Tanks WM-180 through WM-189 as 

reported in Palmer et al. (1999) for stainless-steel 

coupons placed in these tanks for times ranging 

from 10 to 30 yr. From these studies, the 

maximum uniform corrosion rates ranged from 

2.0E−05 to 5.3 E−02 mpy, where mpy is mils per 

year and mils are 0.001 in. The degradation 

calculations used a conservative value of 10-5 m/yr 

(3.9 E−01 mpy), which is greater than the 

maximum observed corrosion rate and is 

considered reasonable and conservative given the 

available data are not consistent with the expected 

grouted vault and tank conditions and do not 

account for corrosion over the time period of 

interest (i.e., 1,000 yr). 

In the vault, tank, and piping degradation 

calculations, the thickness of the reinforcing steel, 

tank wall, and piping were assumed to be 0.5 in. 

(1 cm), 3/16 in. (0.5 cm), and 0.1 in. (0.3 cm), 

respectively. These values were selected based on 

design information and represent a conservative 

(i.e., minimum thickness) value. For example, the 

tank wall thickness was assumed to be 3/16 in. 

(0.5 cm), while the tank wall thickness varied 

from 3/16 in. (0.5 cm) on the dome, to 1/4 in. 

(0.6 cm) on the upper walls, to 5/16 in. (0.8 cm) 

on the lower walls. 

For the tank and piping, corrosion was 

modeled as occurring inside and outside of the 

walls (i.e., from the grout inside and outside the 

walls). Since corrosion was modeled from the 

inside and outside of the walls, complete 

corrosion of the tank and piping was assumed to 

occur when one-half of the wall thickness 

corroded. Corrosion of reinforcing steel was 

assumed to occur from the outside of the 

cylindrical reinforcing steel/concrete vault 

interface and complete corrosion was assumed 

once 50% of the reinforcing steel corroded. 

Using the extreme values from the range of 

maximum uniform corrosion rates in Palmer et al. 

(1999) and the assumptions of vault, tank, and 

piping stainless steel thickness and direction of 

corrosion attack, the failure times of the stainless 

steel in the vault, tank, and piping range from 

10 million to 5,000 yr, 5 million to 2,000 yr, and 3 

million to 1,000 yr for the minimum and 

maximum corrosion rates, respectively. In the 

DUST-MS calculations, the times for the vault 

and tank/piping failure of 100 and 500 yr were 

used. Additionally, the possible corrosion 
7-16

mechanisms for the stainless-steel coupons in the 

tanks (Palmer et al. 1999) are for conditions that 

are more aggressive than those expected in the 

grouted vault, tank, and piping. 

Thus, the failure times used to determine the 

onset of radionuclide release rates in DUST-MS 

for the vault, tank, and piping are conservative 

when compared to the available corrosion data. 

The corrosion model is considered to have 

provided bounding times for corrosion failure of 

the vault, tank, and piping. Furthermore, the 

release rate analysis used initial times for releases 

that were earlier than the times predicted by the 

most conservative of the corrosion rates. 

The assumed degradation times were not 

altered in the groundwater pathway sensitivity/ 

uncertainty analysis. 

7.3.6 Groundwater Pathway-Sensitivity 

and Uncertainty Analysis Results 

This section discusses the results of the 

sensitivity and uncertainty analysis. The results 

are divided into two categories: (1) sensitivity and 

uncertainty analysis based upon the four 

groundwater pathway scenarios in a matrix of 

combinations, and (2) sensitivity/uncertainty 

analysis based upon individual changes in a 

specified parameter. 

7.3.6.1 Groundwater Scenario 

Sensitivity/Uncertainty Results. The four 

groundwater scenarios: (1) worst-case, 

(2) conservative, (3) realistic, and (4) best, were 

analyzed for the groundwater protection scenario 

(i.e., drinking water) and the all-pathways 

scenario doses. These four scenarios were run for 

the matrix of possibilities: (1) inventory, 

(2) infiltration, and (3) transport parameters (i.e., 

distribution coefficients). This matrix resulted in 

36 groundwater run combinations.

The sensitivity/uncertainty analysis was 

conducted for radionuclides 129I, 99Tc, 90Sr/90Y, 

and 14C, which were determined to be the major 

dose drivers for the groundwater pathway in 

Section 4. 

The dose results of the sensitivity/uncertainty 

matrix for the four scenarios are provided in 

Tables 7-10 and 7-11. The conservative case (i.e., 

conservative grout and transport Kd values, 

4.1 cm/yr infiltration, and the conservative 

inventory) used in the dose compliance analysis in 

Section 4 is highlighted in yellow. 

The use of the matrix allows for an 

interpretation of the sensitivity and uncertainty in 

the parameters of interest in the groundwater 

analysis. 

All graphs of the radionuclide releases and 

groundwater doses for each matrix combination 

shown in Tables 7-10 and 7-11 are provided in 

Appendix M. 

The drinking water doses for the four main 

scenarios are shown in Figures 7-2 through 7-5 for 

the radionuclides of interest. The four main 

groundwater scenarios consisted of (refer to 

Table 7-3 for parameter values): 

• Worst-case scenario—consisting of the 

worst-case grout Kd values, worst-case 

transport Kd values, 12.4 cm/yr (4.88 in./yr) 

infiltration, and worst-case inventory 

• Conservative scenario—consisting of the 

conservative grout Kd values, conservative 

transport Kd values, 4.1 cm/yr (1.6 in./yr) 

infiltration, and conservative inventory 

• Realistic scenario—consisting of the 

realistic grout Kd values, realistic transport 

Kd values, 1.1 cm/yr (0.43 in./yr) 

infiltration, and realistic inventory 

• Best scenario—consisting of the same 

transport parameters as the realistic 

scenario with the best inventory. 

The locations of the time of peak drinking 

water dose and magnitude of the peak drinking 

water doses for all other cases would be located 

between the two extremes of the worst-case 

scenario and the best scenario for each 

radionuclide of interest shown in Figures 7-2 

through 7-5. 
7-17 

Table 7-10. Drinking water doses for sensitivity/uncertainty groundwater analyses.a

Parameterization Drinking Water Dose (yr post-closure) (mrem/y

Grout Kd Transport Kd Infiltration Inventory 129I 99Tc 90Sr/90Y 

Worst-Case 23.1 (538) 1.07 (2370) 17.3 (294) 0.008 (Conservative 8.7 0.99 17.3 0.004 Realistic 6.7 0.09 17.3 0.002 

Worst-Case Worst-Case 12.4 

Best 4.4 0.05 17.3 0.0004 Worst-Case 9.09 (607) 0.38 (5060) 3.02 (342) 1E−4 (Conservative 3.42 0.35 3.02 5E−05 Realistic 2.64 0.03 3.02 2E−05 

Worst-Case Worst-Case 4.1 

Best 1.75 0.02 3.02 5E−06 

Worst-Case 2.66 (844) 0.09 (1.75E+04) 0.04 (461) 9E−10 Conservative 1.0 0.09 0.04 4E−10 Realistic 0.77 0.008 0.04 2E−10 

Worst-Case Worst-Case 1.1 

Best 0.51 0.005 0.04 5E−11 Worst-Case 5.72 (635) 0.47 (4270) 0.02 (453) 4E−06 Conservative 2.15 0.43 0.02 2E−06 Realistic 1.66 0.04 0.02 9E−07 

Conservative Conservative 12.4 

Best 1.10 0.02 0.02 2E−07 

Worst-Case 2.05 (890) 0.13 (1.46E+04) 0.001 (551) 1E−11 Conservative 0.77 0.12 0.001 6E−12 Realistic 0.60 0.011 0.001 2E−12 

Conservative Conservative 4.1 

Best 0.39 0.007 0.001 6E−13 Worst-Case 0.49 (1890) 0.04 (4.13E+04) 3.4E−07 (891) 0 Conservative 0.18 0.03 3.4E−07 0 Realistic 0.14 0.003 3.4E−07 0 

Conservative Conservative 1.1 

Best 0.09 0.002 3.4E−07 0 

Worst-Case 1.49 (1060) 0.23 (8100) 4.77E−05 (856) 5E−09 Conservative 0.56 0.21 4.77E−05 2E−09 Realistic 0.43 0.02 4.77E−05 9E−10 

Realistic/Best Realistic/Best 12.4 

Best 0.29 0.01 4.77E−05 2E−10 Worst-Case 0.50 (1960) 0.07 (2.33E+04) 3.53E−07 (988) 2E−17 Conservative 0.19 0.07 3.53E−07 9E−18 Realistic 0.14 0.006 3.53E−07 4E−18 

Realistic/Best Realistic/Best 4.1 

Best 0.09 0.004 3.53E−07 9E−19 

Worst-Case 0.14 (5670) 0.02 (8.05E+04) 1.14E−12 (1310) 0 Conservative 0.05 0.01 1.14E−12 0 Realistic 0.04 0.001 1.14E−12 0 

Realistic/Best Realistic/Best 1.1 

Best 0.03 0.0008 1.14E−12 0 

a. Conservative-case used in dose analysis is highlighted in yellow; green shows the worst-case scenario with the conservative inventory.
7-18 

Table 7-11. All-pathway doses for sensitivity/uncertainty groundwater analyses.a

Parameterization All-Pathways Dose (yr post-closure) (mrem/y

Grout Kd Transport Kd Infiltration Inventory 129I 99Tc 90Sr/90Y 

Worst-Case 40.4 (538) 7.52 (2370) 85.8 (294) 0.04 (1Conservative 15.2 6.98 85.8 0.02 Realistic 11.7 0.630 85.8 0.008 

Worst-Case Worst-Case 12.4 

Best 7.76 0.38 85.8 0.002 Worst-Case 15.9 (607) 2.65 (5060) 15.0 (342) 4.95E−Conservative 5.97 2.46 15.0 2.48E−Realistic 4.61 0.22 15.0 9.94E−

Worst-Case Worst-Case 4.1 

Best 3.05 0.13 15.0 2.48E−

Worst-Case 4.65 (884) 0.685 (1.75E+04) 0.18 (461) 4.31E−Conservative 1.75 0.64 0.18 2.16E−Realistic 1.35 0.06 0.18 8.66E−

Worst-Case Worst-Case 1.1 

Best 0.89 0.03 0.18 2.16E−Worst-Case 9.98 (635) 3.29 (4270) 0.12 (453) 2.01E−Conservative 3.75 3.05 0.12 1.01E−Realistic 2.89 0.28 0.12 4.03E−

Conservative Conservative 12.4 

Best 1.92 0.16 0.12 1.01E−

Worst-Case 3.59 (890) 0.94 (1.46E+04) 0.006 (551) 5.75E−Conservative 1.35 0.87 0.006 2.88E−Realistic 1.04 0.08 0.006 1.15E−

Conservative Conservative 4.1 

Best 0.69 0.05 0.006 2.88E−Worst-Case 0.86 (1890) 0.25 (4.13E+04) 1.68E−06 (891) 0 Conservative 0.32 0.23 1.68E−06 0 Realistic 0.25 0.21 1.68E−06 0 

Conservative Conservative 1.1 

Best 0.17 0.01 1.68E−06 0 

Worst-Case 2.61 (1060) 1.62 (8100) 2.36E−04 (856) 2.23E−Conservative 0.98 1.50 2.36E−04 1.11E−Realistic 0.76 0.14 2.36E−04 4.47E−

Realistic/Best Realistic/Best 12.4 

Best 0.50 0.08 2.36E−04 1.11E−Worst-Case 0.87 (1960) 0.5 (2.33E+04) 1.75E−06 (988) 9.02E−Conservative 0.33 0.46 1.75E−06 4.51E−Realistic 0.25 0.04 1.75E−06 1.81E−

Realistic/Best Realistic/Best 4.1 

Best 0.17 0.02 1.75E−06 4.52E−

Worst-Case 0.24 (5670) 0.11 (8.05E+04) 5.65E−12 (1310) 0 Conservative 0.088 0.1 5.65E−12 0 Realistic 0.068 0.009 5.65E−12 0 

Realistic/Best Realistic/Best 1.1 

Best 0.045 0.005 5.65E−12 0 

a. Conservative-case used in dose analysis is highlighted in yellow; green shows the worst-case scenario with the conservative inventory. 
7-19

10-8

10-6

10-4

10-2

100

102

100 101 102 103 104 105 106

I-129 (worst-case)

I-129 (conservative)

I-129 (realistic)

I-129 (best)

Drinking Water Dose (mrem/yr)

Time (yr)

Figure 7-2. 129I drinking water doses for the four groundwater scenarios. 

10-16

10-14

10-12

10-10

10-8

10-6

10-4

10-2

100

102

100 101 102 103 104 105 106

Tc-99 (worst-case)

Tc-99 (conservative)

Tc-99 (realistic)

Tc-99 (best)

Drinking Water Dose (mrem/yr)

Time (yr)

Figure 7-3. 99Tc drinking water doses for the four groundwater scenarios. 
7-20

10-16

10-14

10-12

10-10

10-8

10-6

10-4

10-2

100

102

100 101 102 103 104 105 106

Sr-90/Y-90 (worst-case)

Sr-90/Y-90 (conservative)

Sr-90/Y-90 (realistic)

Sr-90/Y-90 (best)

Drinking Water Dose (mrem/yr)

Time (yr)

Figure 7-4. 90Sr/90Y drinking water doses for the four groundwater scenarios. 

10-15

10-13

10-11

10-9

10-7

10-5

10-3

10-1

101

100 101 102 103 104 105 106

C-14 (worst-case)

C-14 (conservative)

C-14 (realistic)

C-14 (best)

Drinking Water Dose (mrem/yr)

Time (yr)

Figure 7-5. 14C drinking water doses for the four groundwater scenarios.
7-21

The doses for the all-pathways scenarios are 

provided in Appendix M and are not shown here 

because they are similar in character to the 

drinking water dose graphs in terms of arrival 

time and location of the peak dose. The only 

difference is the magnitude of the doses, as can be 

seen from a comparison of Tables 7-10 and 7-11. 

The drinking water doses presented in 

Table 7-10 for the matrix of scenarios shows a 

maximum drinking water dose of 23.1 mrem/yr at 

538 yr post-closure for the worst-case scenario. 

This dose is the result of 129I releases from the 

tanks. The next highest contributor to the drinking 

water dose for the worst-case scenario is from 

90Sr/90Y with a drinking water dose of 

17.3 mrem/yr at 294 yr post-closure. This dose is 

due to releases from the sand pad. The 90Sr/90Y 

doses are the same for all inventory combinations 

because the sand pad inventory was held constant 

throughout the analyses. The sand pad releases 

dominate the 90Sr/90Y doses because the releases 

occur at 100 yr and the decay time is less than the 

releases from the tank inventory. 

99Tc and 14C doses are low for all scenarios 

and do not contribute significantly to the overall 

doses from the TFF. This is due to the grout Kd 

values, which limit the releases of these 

radionuclides from the tanks and vaults. 

The conservative case was used in the 

Section 4 dose analysis (highlighted in yellow in 

Tables 7-10 and 7-11). This case is considered the 

applicable set of parameter values for comparison 

to the performance objectives. The infiltration rate 

of 4.1 cm/yr (1.6 in./yr) was considered applicable 

to an earthen cover. Currently the TFF is covered 

by 10 ft (3 m) of earthen material. In addition, 

future closure work at the TFF for CERCLA 

actions will provide an additional engineered 

cover over the TFF, reducing the infiltration 

further than that assumed in the conservative case. 

The conservative inventory also is considered 

appropriate for the conservative case, because 

current knowledge of the cleaning process, 

completed on Tank WM-182, shows that the 

process is effective. The conservative case was 

based on the conservative Kd values for the grout 

and unsaturated zone interbeds and sediments. 

There is a large uncertainty in the distribution 

coefficients; therefore, the conservative case used 

the average values of the reported ranges in the 

literature. 

The large uncertainty in the grout and 

transport distribution coefficients is addressed in 

Tables 7-10 and 7-11. The uncertainty is shown 

by evaluation of the worst-case scenario grout and 

transport Kd values, in combination with an 

infiltration rate of 4.1 cm/yr (1.6 in./yr) and the 

conservative inventory (i.e., highlighted in green 

in Tables 7-10 and 7-11). In this case, it has been 

shown that the resulting doses are less than the 

performance objectives. In other words, even if 

the distribution coefficients for grout and transport 

are selected at the worst-case level, the doses do 

not exceed the performance objectives. 

The doses in Tables 7-10 and 7-11 for the 

worst-case combinations located above the 

highlighted green area (worst-case Kd values, 

4.1 cm/yr (1.6 in./yr) infiltration, and conservative 

inventory) are provided to show the worst 

possible combinations of parameters. However, 

these combinations of parameter values are not 

considered applicable to the TFF. 

Tables 7-10 and 7-11 can be used to evaluate 

the sensitivity and uncertainty of the parameter 

values. Based on the results of the sensitivity/ 

uncertainty analysis, the following observations 

were made: 

• Model runs are not sensitive to inventory 

changes due to the small changes in 

inventory in comparison to the ranges of 

possibilities for the other parameters in the 

analysis. 

• Grout Kd values have a broader range 

reported in the literature than the Kd values 

for the stratigraphic units. Therefore, the 

groundwater scenario doses are affected 

more by the selection of the grout Kd values 

than the transport Kd values. 

• Sand pad releases dominate the resultant 

dose calculations for 90Sr/90Y transport 

simulations due to the release time of 

100 yr, compared to the tanks at 500 yr and 

the subsequent decay. 
7-22

• Dose peaks shift in time (i.e., later arrival 

time for increasing Kd values) for variations 

in sorption coefficients in the unsaturated 

zone and cause a subsequent decrease in the 

doses. 

• Decreases in infiltration shift the dose peak 

to later arrival times and also cause a 

decrease in the radionuclide doses. 

• Changes in predicted radionuclide doses 

and peak arrival times are less sensitive to 

variations in hydraulic dispersivity values. 

7.3.6.2 Sensitivity and Uncertainty of 

Individual Parameters. The sensitivity and 

uncertainty in each parameter selection also was 

investigated. The sensitivity and uncertainty of the 

drinking water doses based on variations in 

dispersion, unsaturated zone Kd values, infiltration 

rates, grout Kd values, and inventory were 

investigated separately.

Figure 7-6 provides a comparison of the effect 

of changing the dispersion coefficient from the 

worst-case values to the best-case values (refer to 

Table 7-3 for parameter values). The doses are 

based on the worst-case parameter values for all 

remaining parameters; therefore, only the 

dispersion coefficients were changed. 

The use of the best-case dispersion 

coefficients (i.e., higher values of dispersion) 

results in lower doses than the worst-case 

dispersion coefficients (i.e., lower values of 

dispersion) due to additional dispersion of the 

contaminant plume during transport. The peak 

dose arrival time also decreases as expected. 

However, in comparison to the other parameter 

changes, the dispersion coefficient is a less 

sensitive parameter. 

Figure 7-7 provides a comparison of the effect 

of changing the unsaturated zone Kd values for 

basalts and interbed sediments. The evaluation of 

the Kd values was done using four cases: (1) zero 

Kd values for basalts and interbed sediments, (2) 

basalt and sediment Kd values assigned to each 

appropriate unit, (3) only the sediment Kd value 

assigned and the basalt assigned zero Kd values, 

and (4) basalt Kd value assigned and the sediment 

Kd assigned zero Kd values. 

10-4

10-3

10-2

10-1

100

101

102

0 200 400 600 800 1000

I-129 (worst dispersion)

I-129 (best dispersion)

Sr-90/Y-90 (high dispersion)

Sr-90/Y-90 (best dispersion)

Drinking Water Dose (mrem/yr)

Time (yr)

Radionuclide Best-Case Dispersion Worst-Case Dispersion 

I-129 13.7 mrem/yr (535 yr) 23.1 mrem/yr (538 yr) 

Sr-90/Y-90 10.8 mrem/yr (291 yr) 17.3 mrem/yr (294 yr) 

Figure 7-6. Sensitivity/uncertainty of dispersion coefficients on drinking water doses. 
7-23

10-4

10-3

10-2

10-1

100

101

102

0 200 400 600 800 1000

Zero Kd

Basalt Kd

Sediment Kd

Basalt/Sediment Kd

Drinking Water Dose (mrem/yr)

Time (yr)

Radionuclide Zero Kd Basalt/Sediment Kd Sediment Kd Basalt Kd 

I-129 23.1 mrem/yr (538 yr) 23.1 mrem/yr (538 yr) 23.1 mrem/yr (538 yr) 22.9 mrem/yr (539 yr) 

Sr-90/Y-90 21.7 mrem/yr (541 yr) 0.065 mrem/yr (706 yr) 0.29 mrem/yr (661 yr) 3.2 mrem/yr (579 yr) 

Figure 7-7. Sensitivity/uncertainty of unsaturated zone Kd values on 129I and 90Sr/90Y drinking water 

doses.

First, the importance of the interbed sediment 

Kd values and the basalt Kd values were 

investigated using 129I. The basalt and interbed 

sediments were independently assigned the same 

values in the sensitivity analysis, (i.e., 0.01 mL/g). 

This analysis shows that the basalts and interbeds 

are of equal importance in the transport analysis 

when assigned the same Kd value independently. 

The sensitivity and uncertainty in the 90Sr 

groundwater transport analysis was investigated 

by assigning different Kd values to the interbed 

sediments (12 mL/g) and the basalts (6 mL/g). 

The analysis shows that the interbed sediments 

with the higher Kd value were the most important 

parameter in the analysis as expected. This is 

important because the interbed sediments tend to 

have higher Kd values but are less uncertain in the 

analysis than the assignment of Kd values for the 

basalt units. 

The sensitivity and uncertainty in the 

infiltration rate also was investigated. Figures 7-8 

and 7-9 present the results of the infiltration 

sensitivity/uncertainty analysis. The infiltration 

rate is an important factor in the analysis as can be 

seen by these results. The infiltration rate affects 

not only the transport rate through the unsaturated 

zone but also the release rate from the waste form. 

The infiltration analysis for 129I, shown in 

Figure 7-8, shows the change in arrival time of the 

peak dose being increased in time for the lower 

infiltration rate, as expected. In addition, the 

resulting doses also decrease with the decreasing 

infiltration rate due to the reduction in the release 

of contaminants from the waste forms. 

The infiltration analysis for 90Sr/90Y, shown in 

Figure 7-9, again shows that the arrival time of the 

peak doses are shifted to an increased time for the 

lower infiltration rates. In addition, the doses are 

reduced for lower infiltration rates due to the 

lower infiltration releasing less contaminant from 

the waste forms. The 90Sr/90Y dose results are 

more sensitive to changes in infiltration than 129I 

because to the increased arrival times and 

subsequent decay. 

The sensitivity and uncertainty in the grout 

distribution coefficients (Kd values) also was 

investigated. Figures 7-10 and 7-11 show the 
7-24

10-4

10-3

10-2

10-1

100

101

102

0 200 400 600 800 1000

12.4 cm/yr

4.1 cm/yr

1.1 cm/yr

Drinking Water Dose (mrem/yr)

Time (yr)

 12.4 cm/yr 4.1 cm/yr 1.1 cm/yr 

I-129 23.1 mrem/yr (538 yr) 9.1 mrem/yr (607 yr) 2.7 mrem/yr (883 yr) 

Figure 7-8. Sensitivity/uncertainty of infiltration on 129I drinking water doses. 

10-4

10-3

10-2

10-1

100

101

102

0 200 400 600 800 1000

12.4 cm/yr

4.1 cm/yr

1.1 cm/yr

Drinking Water Dose (mrem/yr)

Time (yr)

 12.4 cm/yr 4.1 cm/yr 1.1 cm/yr 

Sr-90/Y-90 17.3 mrem/yr (294 yr) 3.0 mrem/yr (341 yr) 0.036 mrem/yr (457 yr) 

Figure 7-9. Sensitivity/uncertainty of infiltration on 90Sr/90Y drinking water doses. 
7-25

10-20

10-18

10-16

10-14

10-12

10-10

10-8

10-6

10-4

10-2

100

0 200 400 600 800 1000

Worst-Case

Conservative

Realistic/Best

Release Rate (Ci/yr)

Time (yr)

 Worst-Case Conservative Realistic/Best 

I-129 2.38E−05 Ci/yr (534 yr) 8.18E−06 (620 yr) 2.42E−06 (940 yr) 

Figure 7-10. Sensitivity/uncertainty of grout Kd values on 129I releases. 

10-20

10-18

10-16

10-14

10-12

10-10

10-8

10-6

10-4

10-2

100

0 200 400 600 800 1000

Worst-Case

Conservative

Realistic/Best

Release Rate (Ci/yr)

Time (yr)

 Worst-Case Conservative Realistic/Best 

Sr-90 0.031Ci/yr (124 yr) 0.012 Ci/yr (147 yr) 0.0034 Ci/yr (182 yr) 

Figure 7-11. Sensitivity/uncertainty of grout Kd values on 90Sr/90Y releases. 
7-26

sensitivity and uncertainty in the release rates for 

129I and 90Sr. Again, the result of increasing the Kd

value from the worst-case parameter value (i.e., 

lower grout Kd value) to the best case (i.e., higher 

grout Kd value) is a delay in the time of peak 

release and a decrease in the release rate. The 

groundwater pathway dose analyses are sensitive 

to the grout Kd value because the reported ranges 

in the grout Kd values are large. 

The sensitivity and uncertainty in the 

inventories also was investigated. Figures 7-12 

and 7-13 provide the change in the drinking water 

doses based on the four tank inventories. The peak 

dose arrival times are the same for each case 

because the transport parameters were held 

constant. The sensitivity and uncertainty in the 

inventories shows a reduction in the doses in 

direct proportion to the decreases in the 

radionuclide inventories. The groundwater 

pathway dose analyses are not very sensitive to 

the inventory changes because the range of 

inventories is not as great as the range of other 

parameter values.

7.4 Sensitivity/Uncertainty in 

Hydraulic Parameters 

The modeling approach discussed in detail in 

Section 3 attempts to minimize subsurface water 

transport uncertainties due to lack of data. By 

adopting a detailed stratigraphic model 

representation, using site data that was confirmed 

by two or more sources and using basalt 

permeabilities based on detailed lithologic 

analysis presented in Anderson, Kuntz, and Davis 

(1999), the magnitude of model uncertainty was 

significantly lowered. Model calibration with the 

existing perched zone data yielded a reasonable 

approximation of the system—further reducing 

model uncertainty. The following discussion 

provides an analysis of the sensitivity of the 

vertical conductivity parameter value on the 

formation of perched water in the model. 

7.4.1 Vertical Conductivity Sensitivity 

Analysis 

Figure 7-14 compares the distribution of 

perched zones as a function of vertical 

permeability variations in the D-DE2 interbed 

(Zone 9). This interbed is associated with the 

upper-most perched zone at several locations 

across the site and has been identified by borehole 

data as containing low permeable sediments 

capable of restricting infiltrating water. The 

simulations were conducted using recharge solely 

from the Big Lost River versus the model 

calibration that was conducted using the 

additional source of recharge from the percolation 

ponds. Before the 1000-yr simulations, the 

percolation ponds will be removed from service. 

Consequently, the sensitivity analysis is 

conducted on representative moisture conditions 

during future model simulation. 

For the simulation, a calibrated vertical 

permeability value of 0.0025 m/d was 

incorporated into the model for D-DE2 interbed. 

As shown in the Figure 7-14, this permeability 

value was varied by plus and minus one order of 

magnitude to evaluate the impact on perched zone 

distribution. The upper-perched zones vary from 

nearly nonexistent to relatively extensive in the 

Tank Farm area. These results show that the 

model is sensitive to permeability changes in the 

important stratigraphic units. The calibrated 

permeability is within the measured range for the 

type of interbed lithology. Therefore, the 

permeability values used in the model are 

reasonable and the model is dynamic in 

responding to input variations. 

7.4.2 Transport Resulting From Dam 

Failure Flood 

The worst-case flooding scenario reported in 

the site literature is a dam failure analysis 

conducted by Van Haaften et al. (1984). A PMF 

event above Mackay Dam results in a 1.1 m 

(3.7 ft) overtopping of the dam, with a 0.5-hr dam 

failure, 42.6-m (140-ft) breach in the dam, and no 

flow losses downstream. The result of this worstcase scenario is flooding of the INTEC facility 

with 1 to 2 m (3 to 7 ft) of water covering the site 

and a peak flow of 3292.5 m3

/s (116,272 ft3/s). 

The flood hydrograph for the base-case discharge 

shows a peak flow that lasts only 7.5 hr 

(Figure 2-17, Section 2). Since the INTEC facility 

is located at the margins of the flood, the area 

would be inundated for a lesser a period of time. 
7-27

10-5

10-4

10-3

10-2

10-1

100

101

102

500 550 600 650 700

Worst-Case

Conservative

Realistic

Best

Drinking Water Dose (mrem/yr)

Time (yr)

 Worst-Case Conservative Realistic Best 

I-129 23.1 mrem/y 8.7 mrem/yr 6.7 mrem/yr 4.4 mrem/yr 

Figure 7-12. Sensitivity/uncertainty of inventory on 129I drinking water doses. 

10-5

10-4

10-3

10-2

10-1

600 650 700 750 800 850 900 950 1000

Worst-Case

Conservative

Realistic

Best

Drinking Water Dose (mrem/yr)

Time (yr)

 Worst-Case Conservative Realistic Best 

Sr-90/Y-90 0.065 mrem/yr 0.056 mrem/yr 0.046 mrem/yr 0.030 mrem/yr 

Figure 7-13. Sensitivity/uncertainty of inventory on 90Sr/90Y drinking water doses. 
7-28

Figure 7-14. Sensitivity of the moisture content distribution in the unsaturated zones with variations in 

vertical hydraulic conductivity values in the D-DE−2 interbed (Zone 9) of 0.025, 0.0025, and 

0.00025 m/d. 
7-29

The impact on radionuclide transport resulting 

from such a flooding event was evaluated by 

simulating increased infiltration in the region of 

the flood. It was assumed that flood occurred at 

500 yr—the same time as when the tanks failed. 

Based on double-ring infiltrometer tests 

conducted on desert alluvial soils by Kearl (1981), 

the infiltration rate was measured at 2.5 cm/hr 

(0.98 in./hr). Assuming the flood at the facility 

and subsequent ponding water continues to 

infiltrate for a full day, a total of 61 cm (24 in.) of 

water was input in the simulation. Transport 

simulations indicate this large amount of 

infiltration forced radionuclide migrations 

laterally in the model domain due to the dominant 

influence of the Big Lost River boundary 

condition under higher saturation conditions. 

Consequently, in order to evaluate the impact of 

flooding and subsequent increases in perching 

water on present transport pathways, the 

infiltration due to flooding was lowered to a rate 

of 100 times the 12.4 cm/yr (4.88 in./yr) 

infiltration rate. The extent of the flooding as 

illustrated in Figure 2-18 shows that the entire 

modeling domain is covered with water. 

A transport comparison for a unit 

concentration of 129I under non-flooding and 

flooding conditions is presented in Figure 7-15. 

For non-flooding conditions, a peak concentration 

is reached during the early portion of the 

simulation with a continual decrease in 

concentration values. For the flooding conditions, 

a peak concentration is reached slightly earlier 

than non-flooding conditions. However, the peak 

concentration resulting for flooding conditions is 

less than for non-flooding conditions. The 

perching layers below the Tank Farm increase in 

thickness and extent during the flooding scenario 

and retard the movement of contaminants 

downward (Figure 7-16). As contaminants spread 

across the surface of the perched zones, dilution 

occurs from mixing with the perched water and 

infiltration from precipitation. The rate of decline 

in 129I concentrations in the flooding simulation is 

due to the amount of additional water that must 

move through the system. Once the water from 

the flooding event has passed, concentrations 

continue to decline to minimal levels. 

7.4.3 Transmissivity in the Regional 

Aquifer 

There is a large range of transmissivity values 

for the basalt aquifer reported in site literature. 

There are numerous investigations of magnitude 

and distribution of hydraulic conductivity in the 

regional aquifer. This includes work by Thomas 

(1988), Ackerman (1991), and Anderson, Kuntz, 

and Davis (1999). Based on the available data, the 

highest range of hydraulic conductivity resulting 

in the highest transport rates were used in the 

model simulations resulting in a conservative 

approach to the modeling. 

7.4.4 Big Lost River Seepage 

Seepage from the Big Lost River is a major 

driving force for water movement in the 

unsaturated zone underlying the INTEC facility. 

Detailed stream loss measurements were 

conducted by Bennett (1990) for the Big Lost 

River. These measurements showed seepage 

losses in the vicinity of the INTEC facility. The 

losses were used to estimate recharge to the 

unsaturated zone at the site. 

The modeling assumed that the bulk of the 

perched water zones underlying the site were the 

result of seepage from the Big Lost River. 

Resulting calibration simulations using existing 

perched zone data compared favorably with 

measured seepage rates, increasing confidence in 

the seepage values and reducing uncertainty. 

7.4.5 Basalt Porosity 

The value chosen for the basalt porosity 

controls the rate of water and contaminant 

transport, particularly in the aquifer. Transport 

rates are moderately affected by the selection of 

this parameter. Available porosity measurements 

show a range of 5 to 15% with a historical 

acceptable value of 10%. A value of 10% porosity 

was used in the modeling. 

7.4.6 Unsaturated Hydraulic 

Characteristic 

Unsaturated hydraulic characteristic curves 

are based on measurements for porous media and 
7-30

10-8

10-7

10-6

10-5

10-4

500 520 540 560 580 600

No Flood

Concentration (Ci/m3

)

Time (yr)

10-8

10-7

10-6

10-5

10-4

500 520 540 560 580 600

Flood (100 Times 12.4 cm/yr)

Concentration (Ci/m3

)

Time (yr)

Figure 7-15. Comparison of radionuclide concentrations at the water table for non-flooding and flooding 

conditions. 
7-31

Figure 7-16. Extent of perched water due to non-flooding conditions and flooding conditions. 

Flood

No Flood
7-32

are poorly defined for water movement in 

unsaturated basalt fractures. Based on work 

conducted by Magnuson (1995), characteristic 

curves for basalt units were adapted for use in the 

model. These curves assume the instantaneous 

movement of water in the basalts with only a 

minor decrease in capillary pressures. 

This parameter impacts transport rates in the 

unsaturated basalts. The rapid movement of water 

in response to small capillary pressure decreases 

actually predicts, conservatively, the movement of 

water and contaminants, which mitigates model 

uncertainty. 

7.5 Piping Inventory 

Uncertainty 

The inventory in the piping is presented in 

Section 2.3 and Appendix A. The piping was 

assumed to be released at the same time as the 

tank inventory at 500 yr. Previous TFF PA 

analyses, presented in Appendix F, were based on 

the inventory releases being controlled by 

diffusion to initial breaks in the pipes beginning at 

the time of closure. The analysis presented in this 

report is based on the fact that the piping is 

stainless steel and the insides of the pipes will be 

grouted along with the concrete trenches that 

provide secondary containment. Therefore, the 

piping degradation would be similar to the tank 

degradation. 

The uncertainty in the piping inventory can be 

evaluated by simple evaluation of the original 

mass assumed to be located in the piping (see 

Section 2.3 and Appendix A). The original mass 

assumed to be located in the piping was 15.5 kg. 

The piping drinking water doses are directly 

proportional to the initial mass of contaminants 

assumed to be in the piping. The drinking water 

doses from piping releases for the conservative 

case were 90Sr/90Y at 4.0E−09 mrem/yr, 99Tc at 

2.9E−04 mrem/yr, 14C at 3.6 E−19 mrem/yr, and 

129I at 3.7E−03 mrem/yr. As can be seen, the doses 

from the piping releases are low. An order of 

magnitude increase in the assumption of the 

original contaminant mass in the piping would not 

result in a significant increase in the drinking 

water or all-pathways doses due to piping 

contamination. 

7.6 Sand Pad Uncertainty 

The sand pad analysis presented in 

Section 2.3 was based on the assumption that the 

sand pad was saturated during the vault 

contamination event. Therefore, the contaminants 

in the vault liquid were assumed to reach the sand 

pad only by diffusion from the vault liquid to the 

sand pad pore space. The diffusion assumption 

was assessed against the assumption that the sand 

pad was at residual saturation during the 

contamination event. Therefore, the contaminants 

from the vault liquid would directly fill the sand 

pad pore volume. 

Two assumptions were made in the analysis: 

(1) sand pad residual saturation is 0.3 and (2) total 

sand pad porosity is 0.34. The contaminated vault 

liquid was assumed to fill the pore voids in the 

sand pad. The sand pad was then washed and 

decayed as discussed in Section 2.3 and 

Appendix A. Table 7-12 compares the two 

conceptual model results for the sand pad. 

Table 7-12 presents the increase of activity in 

the sand pad based on the pore-filling assumption 

in comparison to the original diffusion 

assumption. As can be seen, the change in the 

activity estimates are less than one order of 

magnitude. 

The results of this analysis show that the 

change in conceptualization of the sand pad 

contamination events does not change the results 

significantly. The differences in the drinking 

water doses due to the sand pad conceptual 

models are compared in Table 7-13 for the 

conservative case. 

Table 7-13 shows that the difference in the 

conceptualization of the original contamination 

event does not result in a significant increase in 

the drinking water doses due to the sand pad. The 

same can be said for the all-pathways doses. 

Therefore, the uncertainty in the sand pad 

inventories due to conceptual model differences 

does not result in a change in compliance with the 

performance objectives. 
7-33

Table 7-12. Comparison of the sand pad conceptual models. 

Radionuclide 

Diffusional Model Sand Pad 

Activity at 2016 

(Ci) 

Pore-Filled Model Sand Pad 

Activity at 2016 

(Ci) 

Activity Factor 

Increase 

90Sr 226 295 1.3 

99Tc 2.0E−12 1.9E−11 9.5 

14C 3.9E−07 8.7E−07 2.2 

129I 1.1E−06 5.0E−06 4.5 

Table 7-13. Comparison of the sand pad drinking water doses for each conceptual model. 

Radionuclide 

Diffusional Model Dose Results 

(mrem/yr) 

Pore-Filled Model Dose Results 

(mrem/yr) 

90Sr/90Y 1.2E−03 1.6E−03 

99Tc 9.93E−12 9.4E−11 

14C 5.7E−17 1.2E−16 

129I 1.8E−04 8.1E−04 

7.7 DUST-MS Release Model 

Uncertainty 

7.7.1 Heterogeneity in Tank Grout 

This section presents a qualitative discussion 

showing considerations in the release rate 

modeling approach and selection of input 

parameters followed by a quantitative analysis 

relating the implications to modeling 

heterogeneities in the source term conceptual 

model. 

The effects of heterogeneities in the tank 

grout (i.e., shrinkage cracks) on degradation and 

release rates are not directly modeled in the 

DUST-MS source term release rate calculations. 

However, because of the lack of available sitespecific information and uncertainty and 

variability in the available information, reasonable 

and conservative assumptions were made in the 

source term release rate calculations. When sitespecific data were not available, upper-limit 

assumptions were made (e.g., diffusion 

coefficient). These assumptions are described 

throughout this document. 

For example, the DUST-MS estimations of 

release rates from the tank and sand pad are based 

on a number of conservative assumptions. Piping, 

vault, and tank failure times used in the release 

rate determinations are 0, 100, and 500 yr. These 

values are earlier than the times predicted from 

the degradation analysis. Moreover, the 

degradation analyses used parameters that were 

considered reasonable and conservative based on 

the available data. 

A quantitative analysis was conducted to 

explore and compare the release rates using 

DUST-MS and also using a first-order release rate 

model. Releases were computed assuming matrix 

diffusion into the shrinkage cracks between the 

grout and tank wall. In both models, the bottom 

1 ft (0.3 m) of the grout tank was assumed to 

contain the radionuclides present in the tank. 

In the first-order model release rate, a mass 

balance equation based on a control volume 

approach was employed. This same approach was 

used in the hand calculation performed for the 

sensitivity analyses to verify DUST-MS release 

rates. Following this description, the approach to 
7-34

compute release rates based on diffusion from the 

source into concrete cracks when fracture flow 

occurs is presented. This approach is consistent 

with the analysis used to determine the mass of 

radionuclides in the sand pad. The radionuclide 

129I is modeled in these analyses using worst-case 

scenario values. 

7.7.1.1 Release Rates from Analytical 

Model. The rate of change of mass inside the 

control volume (i.e., the tank source) decreases 

from water infiltrating and mobilizing 

radionuclides in the tank source. This relationship 

is expressed in Equation (7-1). 

C Q SA

dt

dM

P il

T = × inf × (7-1) 

where 

MT = total mass in the tank source 

(Ci, assumed initially to be a 

unit source) 

t = time (yr) 

Cp = pore water concentration in 

the tank source (Ci/cm3

) 

Qinfil = flow rate of infiltrating water 

in the tank source (12.6 cm/yr) 

SA = source surface area normal to 

flow (cm2

, circle with a radius 

of 760 cm). 

The tank source is assumed to be cylindrical 

with a radius of 7.6 m (25 ft) and a height of 1 ft 

(0.3 m). In this vault/tank conceptualization, the 

radionuclides in the tank are completely mixed in 

the tank source volume. 

To determine the mass remaining in the tank 

source, MT, it was assumed that the radionuclides 

partitioned between the water in the pore space 

and the solid (i.e., degraded grout) and the 

radionuclide concentrations are at equilibrium 

according to the partition coefficient. The 

relationship between the pore water concentration 

and the total mass contained in the tank source is 

described by Equation (7-2). 

d soil W

T

p K M V

M C

× + = (7-2) 

where 

Msoil = mass of the source 

(5.56E+07 cm3

 at a bulk 

density of 2.12 g/cm3

) 

Vw = volume of source water (cm3

, 

assuming a moisture content 

of 0.242). 

Substituting Equation (7-2) into Equation 

(7-1) and solving for the total mass in the tank as 

a function of time, MT, yields Equation (7-3). 

exp[ ( 500)] MT = M0 −α t − (7-3) 

where 

M0 = initial mass in tank (Ci, unit 

initial source) 

d soil W

il

K M V

Q SA

× + = inf α

t-500 = time since the beginning of the 

release at 500 yr. 

The mass release rate from the tank is 

obtained by differentiating Equation (7-3) with 

respect to time, t and is expressed in 

Equation (7-4). 

exp[ ( 500)] = − M0 − t − dt

dMT α α (7-4) 

Equation (7-4) was evaluated for the same 

times used in the DUST-MS release rate 

calculations shown in Figure 7-17. The DUSTMS release rates from the tank source, together 

with the release rates computed using the 

analytical solution in Equation (7-4) are provided 

in Figure 7-17. 

7.7.1.2 Release Rates into Concrete 

Cracks. To determine the release rate into 

shrinkage cracks between the grout and tank, it 
7-35

10-6

10-4

10-2

100

0 200 400 600 800 1000

DUST Tank Source Release Rate

Analytical Leaching Model Source Release Rate

Diffusion Model Assuming Concrete Cracks

Release Rate (Ci/yr)

Time (yr)

DUST Peak

 0.39 Ci/yr

Analytical Peak

 0.088 Ci/yr

Diffusion Peak

 0.036 Ci/yr

Figure 7-17. Comparison of DUST-MS tank release rates and analytical calculations for the 129I Tank 

source using worst-case scenario parameters with release rates estimated assuming diffusion into concrete 

cracks with fracture flow. 

was conservatively assumed that the entire 

circumference of the grout/tank interface served 

as a preferential water flow pathway and 

radionuclides diffused from the grout to the 

grout/tank interface crack. At this interface, the 

radionuclide concentration was assumed to be 

maintained at zero. That is, the radionuclides were 

swept away by the water flowing at the grout/tank 

interface crack. The approach and parameters 

used to model diffusion from the grout to the 

shrinkage crack is consistent with the approach 

used to model diffusion into the sand pad. This 

approach is provided in Section 2.3.2 and is 

summarized below. 

To model diffusion, numerous published 

reference materials provided solutions to diffusion 

equations (e.g., Crank 1975; Choy and Reible 

2000). The following equation describes onedimensional diffusion from a semi-infinite media 

∈ [ ∞) ∂

∂

















= ∂

∂ 0, 2

2

( ) z

x

C

R

D

t

C A

fA

A A eff (7-5) 

where 

CA = concentration of radionuclide 

A (M/L3

) 

D A (eff)= diffusion coefficient of 

radionuclide A, assumed to be 

10-5 cm2

/s, which is 

conservative for 

sand/degraded grout (L2

/T) 

RfA = retardation factor of 

radionuclide A in grout 

(unitless) 

z = distance (L). 

where 

CA0 = initial concentration of 

radionuclide A (M/L3

). 

with uniform initial concentration and zero 

concentration at the surface. Equation (7-5) is 

presented in Choy and Reible (2000). 

The boundary conditions associated with 

Equation (7-5) and its solution include the 

following: 
7-36

( ) , 0 0 CA z t z=0 = t > (7-6) 

( ) , 0 CA z t z→∞ = CA0 t > (7-7) 

( ) = ∈ [ ∞) = , 0, 0 0 C z t C z A t A (7-8) 

The boundary and initial conditions Equations 

(7-6) through (7-8) for Equation (7-5) describe a 

semi-infinite system with a constant initial 

concentration of radionuclide A (CA0) and 

represent transport (diffusion) toward the surface 

where the concentration is maintained at zero. The 

zero concentration boundary condition implies 

that as soon as the contaminant diffuses to the 

surface it is immediately removed. For mass 

transport in the grout to the grout/tank interface 

shrinkage crack, this assumption is conservative 

(i.e., does not underestimate mass transfer at the 

grout/tank interface shrinkage crack). 

The analytical solution to Equation (7-5), 

using boundary and initial conditions in Equations 

(7-6) through (7-8) for the surface flux, jA(t), in 

units of M/L2

T is offered in Choy and Reible 

(2000). 

( ) 0 ( )

=0 = 0 t >

t

D R

j t C A eff fA

A z A

π

 (7-9) 

Equation (7-9) determines the amount of a 

contaminant that diffuses from the grout to the 

grout/tank interface shrinkage crack. The flux 

computed using Equation (7-9) is multiplied by 

the area normal to flow in the grout/tank interface 

shrinkage crack. This area is the product of the 

tank circumference (i.e., 2πr, where r is the tank 

radius, 7.6 m [25. ft]) and the height of grout at 

the bottom of the tank that contains radionuclides 

(i.e., 1 ft [0.3 m]). The release rates from the 

grout/tank interface shrinkage cracks computed 

using this approach and Equation (7-9) result in 

the values shown in Figure 7-17. 

From Figure 7-17, the peak release rates 

computed using DUST-MS and the first-order 

release rate model are 0.39 and 0.88 Ci/yr, 

respectively, while the peak release rate, 

determined assuming diffusion from the 

grout/tank interface shrinkage crack is 

0.036 Ci/yr. Thus, conservatively assuming that 

the entire grout/tank wall interface is a 

preferential water pathway and that radionuclides 

diffuse from the bottom 1 ft (0.3 m) of the grout, 

the peak DUST-MS release rate is an order of 

magnitude greater than the peak release rate 

computed assuming diffusion. Even if two or 

three cracks similar to the shrinkage crack at the 

grout/tank interface are assumed, the peak 

diffusion release rates would be less than the 

DUST-MS and first-order model release rates. 

Moreover, if the shrinkage cracks are assumed to 

occur earlier (e.g., at t = 0), the peak diffusion 

release rates are still less than the DUST-MS and 

first-order model release rates. 

Consequently, since the peak dose serves as 

the performance measure, the conservative 

assumptions described at the beginning of this 

section and the results shown in Figure (7-17) 

indicate that the DUST-MS peak release rates are 

reasonable when compared to analytical results. 

Additionally, assuming a large enough tank 

infiltration exists in the grout tank to cause 

fracture flow, peak release rates arising from 

heterogeneities in tank grout cracks are expected 

to be an order of magnitude smaller than the 

DUST-MS peak release rate. This analysis 

confirms that (1) modeling release rates using 

DUST-MS and assuming surface rinsing are 

reasonable and conservative and (2) even if 

conservative assumptions are made, the effects of 

heterogeneities on the peak release when 

compared to the DUST-MS peak release rate is 

small. 

7.7.2 Comparison to Analytical 

Calculations 

Release rates of radionuclides from the TFF 

vaults were estimated using the DUST-MS 

computer code (Sullivan 1993, 1996). Onedimensional DUST-MS transport simulations 

were conducted for radionuclide sources in the 

grouted tank and in the sand pad beneath the tank 

to estimate the radionuclide release rates. To 

evaluate the sensitivity of these release rates to the 

selected DUST-MS release rate model (i.e., 
7-37

surface rinsing), the following comparative 

analysis was conducted between release rates 

computed using DUST-MS and release rates from 

an analytical solution to a first-order mass balance 

equation. 

DUST-MS simulations were conducted for 

129I release rates from the tank assuming worstcase scenario parameters. Release rates were 

calculated from the source in the tank and from 

the concrete vault at the vault floor/basalt 

interface (i.e., following transport through the 

tank, sand pad, and vault floor). These results are 

provided in Figure 7-18. The release rates begin at 

500 yr when the tank is assumed to have totally 

corroded and grout inside the tank is completely 

degraded. The results in Figure 7-18 show the 

effects of transport through the tank, sand pad, 

and vault floor on the release rates from the 

source in the tank. 

For comparison with DUST-MS results, an 

analytical solution to a first-order mass balance 

equation based on a control volume approach was 

employed. The rate of change of mass inside the 

control volume (i.e., the tank source) decreases 

from water infiltrating and mobilizing 

radionuclides in the tank source. This relationship 

is expressed in Equation (7-10). 

C Q SA

dt

dM

P il

T = × inf × (7-10) 

where 

MT = total mass in the tank source 

(Ci, assumed initially to be a 

unit source) 

t = time (yr) 

Cp = pore water concentration in 

the tank source (Ci/cm3

) Qinfil =

 flow rate of infiltrating 

water in the tank source 

(12.6 cm/yr) 

SA = source surface area normal to 

flow (cm2

, circular with a 

radius of 760 cm). 

The tank source is assumed to be cylindrical 

with a radius of 7.6 m (25. ft) and the height of 

1 ft (0.3 m). In this vault/tank conceptualization, 

the radionuclides in the tank are completely mixed 

in the tank source volume. 

To determine the mass remaining in the tank 

source, MT, it was assumed that the radionuclides 

partitioned between the water in the pore space 

10-12

10-10

10-8

10-6

10-4

10-2

100

102

100 1000

DUST Source 

DUST Concrete Vault 

Analytical Leaching Model Source 

Release Rate (Ci/yr)

Time (yr)

Figure 7-18. Comparison of DUST-MS release rates with analytical calculations for the 129I tank source 

using worst-case scenario parameters. 
7-38

and the solid (i.e., degraded grout) is at 

equilibrium according to the partition coefficient, 

Kd. The relationship between the pore water 

concentration and the total mass contained in the 

tank source is described by Equation (7-11). 

d soil W

T

p K M V

M C

× + = (7-11) 

where 

Msoil = mass of the source 

(5.56E+07 cm3

 at a bulk 

density of 2.12 g/cm3

) 

Vw = volume of source water (cm3, 

assuming a moisture content 

of 0.242). 

Substituting Equation (7-11) into Equation 

(7-10) and solving for the total mass in the tank as 

a function of time, MT, yields Equation (7-12). 

exp[ ( 500)] MT = M0 −α t − (7-12) 

where 

M0 = initial mass in tank (Ci, unit 

initial source) 

 

d soil W

il

K M V

Q SA

× + = inf α

t-500 = time since the beginning of the 

release at 500 yr. 

The mass release rate from the tank is 

obtained by differentiating Equation (7-12) with 

respect to time, t, and is expressed in 

Equation (7-13). 

exp[ ( 500)] = − M0 − t − dt

dMT α α (7-13) 

Equation (7-13) was evaluated for the same 

times used in the DUST-MS release rate 

calculations shown in Figure 7-18. The DUSTMS release rates from the tank source and from 

the vault floor together with the release rates 

computed using the analytical solution in 

Equation (7-13) are provided in Figure 7-18. 

In Figure 7-18, the DUST-MS release rate 

peak is larger (about 0.4 versus 0.09) than the 

analytically-based release rate peak. The DUSTMS releases occur sooner than the analyticallybased releases as evidenced in the slopes of these 

release rates following the peak releases. Thus, it 

is concluded from the results in Figure 7-18 that 

the DUST-MS release rates computed with the 

surface rinsing model are consistent and 

conservative relative to release rates determined 

with a mass balance, equilibrium-partitioningbased analytical model. 

7.8 Radon Analysis 

The radon transport analysis was based on a 

very conservative diffusion model. Due to the 

very low flux results for radon, it is difficult to 

pose a situation in which a significant change in 

the result could occur. The most important 

parameter would be the radium concentrations in 

the waste. However, there is no conceivable way 

for the inventory to change enough to increase the 

radon fluxes and make them near the performance 

objective. Table 7-14 summarizes the radon flux 

results for the four inventories (i.e., worst-case, 

conservative, realistic, and best). 

7.9 Volatile Radionuclide 

Analysis 

The analysis of atmospheric doses for volatile 

radionuclides was conducted in a very 

conservative manner for 3

H and 14C. The 

inventory for the releases was based upon the 

conservative inventory. The total dose for the 

atmospheric pathway was 0.51 mrem/yr, which is 

very low in comparison to the performance 

objective of 10 mrem/yr. 

It is difficult to pose a situation in which a 

parameter could change significantly enough to 

cause the atmospheric doses to be near the 

performance objective. Table 7-14 summarizes 

the atmospheric dose results for the four 

inventories (i.e., worst-case, conservative, 

realistic, and best). 
7-39

Table 7-14. Doses for each inventory option for radon flux, atmospheric releases, and intruder scenarios. 

Worst-Case 

Inventory 

Conservative 

Inventory 

Realistic 

Inventory 

Best 

Inventory 

Acute Drilling Intruder (mrem) 276 232 193 144 

Chronic Drilling Intruder 

(mrem/yr) 

106 91.1 75.1 52.8 

Acute Construction Intruder 

(mrem) 

0.89 0.80 0.67 0.45 

Chronic Construction Intruder 

(mrem/yr) 

29.1 26.1 21.8 14.5 

Radon Flux (pCi/m2

/s) 0.48 0.39 0.29 0.19 

Atmospheric (mrem/yr) 1.1 0.51 0.22 0.05 

7.10 Intruder Analyses 

DOE guidance for PAs (DOE 1999b) states 

that the sensitivity/uncertainty analysis for 

hypothetical inadvertent intruder analyses should 

be limited to qualitative arguments (e.g., 

explanation of the rationale for selected scenarios 

and parameters). 

The discussions in Section 5 show that the 

TFF is a unique environment for the evaluation of 

intruder scenarios. Many of the standard scenarios 

were not considered applicable to the TFF 

because the depth of the waste in the tanks is 10 m 

(30 ft). Therefore, only the intruder-drilling and 

post-drilling scenarios were considered applicable 

for the tank and sand pad contamination. 

Additional scenarios were evaluated for the piping 

inventory located less than 3 m (10 ft) below the 

surface. These scenarios included the intruderconstruction scenario and the post intruderconstruction scenario. 

It is difficult to predict the future actions of 

humans, especially over a period of 1,000 yr. The 

fact that the intruder is assumed to drill a well 

directly through a tank is considered a very 

conservative assumption. It can be argued that the 

person drilling the well would not continue to drill 

through the concrete and stainless-steel tanks and 

grout unless it was completely degraded. 

However, to be conservative, the well was 

assumed to be drilled through the waste at 100 yr 

post-closure. The analysis is very sensitive to the 

time of intrusion since the short-lived fission 

products are the main contributor to the intruder 

doses at 100 yr post-closure. 

Uncertainty exists in the state of the concrete 

systems over time. In this analysis, credit was 

taken for the concrete vault providing a barrier to 

intrusion. Oztunali and Roles (1986) state that “if 

resistance is encountered during drilling (as would 

be the case if the drill bit hit solid rocks or intact 

metal), the crew would simply move a few yards 

horizontally and drill a new hole. Drilling 

equipment and bits that can drill through solid 

rock and intact metal are available; however, they 

are more expensive than the equipment and bits 

that are normally required and used for water well 

installation. Similarly, it is anticipated that 

standard drill bits would have difficulty 

penetrating reinforced concrete structures, 

although little difficulty is expected for waste 

stabilized using grout backfill.” Therefore, credit 

could be taken for the reinforced concrete vaults 

and the stainless-steel tanks, further reducing the 

doses for the intruder. 

In Equation (5-4) for the acute intruderdrilling scenario, the radius of the well is not 

present in the equation. Note, however, that the 

concentration in the soil will be the same for any

given well radius, but the total amount of 

contaminated soil will vary with the well radius. 

This equation assumes that the radionuclide 

activity is completely mixed with the drill 

cuttings. The radionuclide activity would not 
7-40

likely be completely mixed with the well cuttings. 

However, when considering the different 

possibilities such as the cuttings from the tank 

being covered with cuttings from below the tank 

and the varying radiation fields that would be 

encountered as the drilling commenced with time, 

it seems appropriate to make this assumption of 

complete mixing. 

A 22-in.-diameter well provides 30 m3

(1,000 ft3

) of contaminated zone (i.e., over the 

depth of the well borehole of 122 m [400. ft]). 

When the soil and waste are spread out over the 

assumed contaminated zone of 2,200 m2

(24,000 ft2

), it results in a contaminated zone 

0.014 m (0.54 in.) thick. As the well diameter is 

increased, the thickness of the contaminated zone 

also increases, thus affecting the external dose 

rate. However, in this intruder analysis, a very 

large-diameter well was used for purposes of 

conservatism. Additionally, the assumption that 

the drill cuttings are spread out over the ground 

and are not located in a mud pit adds more 

conservatism and site-specific application, which 

increases the external dose for this scenario. 

Other conservative parameters were used in 

the acute intruder-drilling scenario as well. These 

include the assumption that the intruder spends 

160 hr drilling the well. Standard drilling 

scenarios typically use an exposure time of 6 hr 

(Oztunali and Roles 1986). 

In Equation (5-14) for the chronic postdrilling scenario, two parameters of importance 

are present in the form of the radius of the well 

(rwell) and the final depth of the contaminated zone 

after tilling (Dc). Variations in the values used for 

these parameters will result in proportional 

changes in the predicted radionuclide 

concentrations in the contaminated zone. 

Therefore, if the depth of the contaminated zone 

after tilling (Dc) is reduced by a factor of 2, the 

contaminated zone concentrations increase by a 

factor of 2. If the radius of the well (rwell) is 

increased by a factor of 2, the contaminated zone

concentrations increase by a factor of 4. The 

selection of parameter values for rwell and Dc are 

considered to be appropriate and based on sitespecific information. 

For both the acute and chronic intruder 

scenarios, the conservative inventory was used. 

Table 7-14 summarizes the intruder dose results 

for the four inventories (i.e., worst-case, 

conservative, realistic, and best). 

The uncertainty in the intruder analyses is 

large and includes both the assumptions made 

during the development of the scenarios and the 

parameter values selected in the formulation of 

the mathematical representations. However, in 

this PA, every attempt was made to consider the 

site-specific environment and habits of the people 

currently in the region. Therefore, the analyses are 

considered to be representative of the INEEL 

region. The predicted doses are considered to be 

conservative, and there is a reasonable assurance 

that the performance objectives will not be 

exceeded for 1,000 yr post-closure. 

7.11 Dose Analyses 

The dose analyses for the groundwater 

pathway contain many parameters dealing with 

human exposure. For the groundwater protection 

pathway, standard drinking water intake rates (i.e., 

730 L/yr) were used to bound the doses. 

The all-pathways doses were calculated 

stochastically to account for the variations in 

human exposure parameters. Each parameter in 

the dose calculations was represented by a 

distribution of values. These distributions were 

then sampled, and the dose results were ranked 

from lowest to highest for 1,000 realizations. The 

950th dose value in the ranking was then used as 

the 95% confidence level. 

In certain circumstances, national average 

data have been used for formulating some of the 

biosphere dose parameter distributions; however, 

site-specific information has been used for others 

as referenced from Maheras et al. (1997), where 

appropriate. Other site-related information (i.e., 

from Yucca Mountain dose analysis, a similar 

semi-arid environment) also has been used where 

appropriate (LaPlante and Poor 1997). 

To the best knowledge of the authors, no 

reliable site-specific information is known to exist 

for the human consumption rate of locally-
7-41

generated water. For reference, Maheras et al. 

(1997) used a constant value of 258 L/yr from 

Yang and Nelson (1984) that was based on 

national data. This parameter was shown to be the 

most important for sensitivity/uncertainty 

analyses. In this case, national average data were 

used to generate the parameter distribution. Since 

national average data would include more data 

from warmer climates than data from cooler 

climates (compared to the INEEL), and because 

the drinking water rate is greater in warmer 

climates, the distribution used for this parameter 

tends toward greater consumption rates than those 

expected for locally-generated data. Hence, the 

rate is conservative. (The values used in the 

Maheras et al. (1997) report are provided in 

Section 3 along with the parameter distributions 

used in the TFF PA analyses for comparison.) 

7.12 Future Work 

Because this PA is considered a living 

document for closure projects such as the TFF, it 

will be reviewed as additional information and 

studies are conducted to determine if it still 

bounds the TFF environment. As additional data 

become available and the PA needs to be revised, 

additional analysis will be required. The following 

areas of future work are presented to facilitate 

discussion for improving the PA in future 

revisions. 

Further work should be conducted to refine 

and confirm the existing radionuclide inventories 

that will be present at the INTEC facility at site 

closure. This work includes additional sampling 

and analysis of existing waste and refinement of 

potential waste estimates for unsampled areas, 

such as the piping and sand pads. Sampling of the 

tanks before closure will be necessary to evaluate 

the 129I inventory to ensure that the groundwater 

protection performance objectives are met. 

Future research and development studies 

should be reviewed and incorporated into the PA 

as they become available to reduce the uncertainty 

and address data gaps identified in this PA. 

Specifically, work conducted in the areas of 

unsaturated hydraulic parameters, distribution 

coefficients, and the behavior of fracture flow in 

the basalt units would be useful for future 

iterations of the PA analyses. Unsaturated 

hydraulic properties needing refinement include 

the moisture characteristic curves, unsaturated 

hydraulic conductivities, and moisture retention 

values used for the interbed. In addition, 

unsaturated data that are representative of the 

behavior of fracture flow under unsaturated 

conditions would improve the analysis of the 

groundwater pathway in the PA. This is 

particularly important in the model simulations 

because the fracture flow is represented as a 

porous flow in the model. 

Future research or literature studies that refine 

the distribution coefficients used in the model 

simulations would be useful for the transport 

calculations in the PA. For example, similar 

investigations conducted at the Hanford facility 

could be evaluated to locate specific tests 

conducted to measure Kd values for basalt and 

interbed units. New information obtained from 

similar site studies should be examined to 

adequately address the sensitivity of this 

important parameter on transport rates of specific 

radionuclides in the subsurface at INTEC.
8-1 

8. PREPARERS 

Mark S. Jarzemba, Ph.D. 

Ph.D., Nuclear Engineering, The Ohio State 

University, 1993 

M.S., Nuclear Engineering, The Ohio State 

University, 1991 

B.S., Engineering Physics, The Ohio State 

University, 1988 

Dr. Jarzemba is a Nuclear Engineer/Health 

Physicist with 8 years of experience in radiation 

transport, dosimetry, and environmental 

radioactivity (man-made and naturally occurring). 

He is the author of 12 peer-reviewed journal 

articles, 10 conference proceedings, and numerous 

contractor reports on these subjects. Much of his 

professional experience is in both high- and lowlevel waste repository performance assessment. 

Dr. Jarzemba provided technical support for 

the evaluation of intruder doses, the all-pathways 

dose assessment, and the radon release assessment 

for the TFF PA. 

Peter M. Kearl, CGWP 

M.S., Hydrology/Hydrogeology, University of 

Nevada-Reno, 1982 

B.S., Geology, Mesa State College, 1979 

Mr. Kearl is a Certified Ground Water 

Professional with over 20 years of experience in 

environmental transport modeling and hydrologic 

assessments. He has been involved in technical 

aspects of performance assessments for the 

Savannah River Site. 

Mr. Kearl conducted the hydrologic 

assessment and groundwater modeling involving 

the contaminant transport simulations for this PA. 

Mr. Kearl has performed hydrologic studies in 

40 states and five foreign countries. 

Robert W. Rice 

M.S., Chemical Engineering, Arizona State 

University, 1987 

B.S., Chemical Engineering, University of 

Colorado, 1983 

B.S., Business, University of Colorado, 1983 

Mr. Rice has 16 years of experience in the 

fields of chemical and process engineering and 

modeling. He participated in the development of 

the Nuclear Regulatory Commission’s Total 

System Performance Assessment computer code 

that is used to conduct probabilistic performance 

assessments for the proposed high-level waste 

repository at Yucca Mountain. This code checks 

the results and assumptions in DOE’s Total 

System Performance Assessment for Yucca 

Mountain. His work experience includes 

unsaturated and saturated zone flow and transport 

modeling, code development and testing, behavior 

of the near-field environment surrounding the 

waste packages, waste package degradation, 

infiltration from the ground surface, and 

determination of dose at the receptor location. 

For this PA report, Mr. Rice was involved in 

the degradation analysis of the tank, vault, grout, 

and piping systems used to estimate the changes in 

permeability over time. 

Edward Kirk Roemer 

B.S., Environmental Management, Columbia 

Southern University, 2001 

A.S., Geology, Mesa State College, 1995 

Mr. Roemer has 17 years of experience in 

environmental field characterization, transport 

modeling, and GPS/GIS graphics. He has been 

involved in technical aspects of performance 

assessments for the Savannah River Site, Nevada 

Test Site, and Yucca Mountain. 

Mr. Roemer provided overall report 

preparation, graphics, and databases for the TFF 

PA. He was involved in the evaluation of the 

intruder doses, all-pathways dose assessment, and 

general development of the PA report. 

David J. Thorne, CHP 

M.S., Health Physics, Colorado State University, 

1988 

B.S., Geology, Mesa State College, 1985 

Mr. Thorne is a Certified Health Physicist 

with 17 years of experience in environmental 

transport modeling and dose assessment. He has 
8-2 

been involved in technical aspects of performance 

assessments for the Savannah River Site, Nevada 

Test Site, and Yucca Mountain. 

Mr. Thorne provided project coordination and 

technical oversight for the TFF PA. He was 

involved in the evaluation of the intruder doses, 

all-pathways dose assessment, and overall 

development of the PA report. 

John C. Walton, Ph. D. 

Ph.D., Chemical Engineering, University of Idaho, 

1991 

M.S., Chemical Engineering, University of 

Washington, 1988 

M.S., Environmental Science, University of 

Virginia, 1981 

B.S., Biology and Geology, Western Illinois 

University, 1979 

Dr. Walton has over 20 years of experience in 

the environmental field. His work has involved the 

development of predictive models for concrete 

degradation and the analysis of fluid flow and 

mass transport through concrete vaults and waste 

forms used to isolate nuclear waste. Dr. Walton 

also developed a model for leakage from flaws in 

geomembrane liners used in landfills and surface 

impoundments, and a mechanistic corrosion model 

for waste package degradation. Other work has 

included the development of a generalized reactive 

transport code used for release rate and corrosion 

models. This code has supported the Nuclear 

Regulatory Commission’s analysis of the 

engineered barriers performance for the proposed 

high-level waste repository at Yucca Mountain. 

Dr. Walton is currently an associate professor at 

the University of Texas at El Paso. 

For this PA report, Dr. Walton was involved 

in the application of concrete degradation and 

corrosion models to the tank, vault, grout, and 

piping systems that were used in the degradation 

analysis to estimate the changes in permeability 

over time.
9-1 

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