Performance Assessment of Near-Surface Facilities for Disposal of Low-Level Radioactive Waste

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152

NCRP REPORT No. 152

PERFORMANCE ASSESSMENT OF NEAR-SURFACE FACILITIES FOR DISPOSAL OF LOW-LEVEL RADIOACTIVE WASTE

PERFORMANCE ASSESSMENT OF NEAR-SURFACE FACILITIES FOR DISPOSAL OF LOW-LEVEL RADIOACTIVE WASTE

N C R P

National Council on Radiation Protection and Measurements

NCRP REPORT No. 152

Performance Assessment of Near-Surface Facilities for Disposal of Low-Level Radioactive Waste

Recommendations of the NATIONAL COUNCIL ON RADIATION PROTECTION AND MEASUREMENTS

December 31, 2005

National Council on Radiation Protection and Measurements 7910 Woodmont Avenue, Suite 400 / Bethesda, MD 20814-3095

LEGAL NOTICE This Report was prepared by the National Council on Radiation Protection and Measurements (NCRP). The Council strives to provide accurate, complete and useful information in its documents. However, neither NCRP, the members of NCRP, other persons contributing to or assisting in the preparation of this Report, nor any person acting on the behalf of any of these parties: (a) makes any warranty or representation, express or implied, with respect to the accuracy, completeness or usefulness of the information contained in this Report, or that the use of any information, method or process disclosed in this Report may not infringe on privately owned rights; or (b) assumes any liability with respect to the use of, or for damages resulting from the use of any information, method or process disclosed in this Report, under the Civil Rights Act of 1964, Section 701 et seq. as amended 42 U.S.C. Section 2000e et seq. (Title VII) or any other statutory or common law theory governing liability.

Disclaimer Any mention of commercial products within NCRP publications is for information only; it does not imply recommendation or endorsement by NCRP.

Library of Congress Cataloging-in-Publication Data National Council on Radiation Protection and Measurements. Performance assessment of near-surface facilities for disposal of low-level radioactive waste. p. cm. — (NCRP report ; no. 152) Includes bibliographical references and index. ISBN-13: 978-0-929600-89-5 ISBN-10: 0-929600-89-4 1. Low level radioactive waste disposal facilities—United States—Evaluation. 2. Radioactive waste disposal in the ground—United States—Evaluation. I. National Council on Radiation Protection and Measurements. TD898.15.P47 2006 363.72'89--dc22 2006018391

Copyright © National Council on Radiation Protection and Measurements 2006 All rights reserved. This publication is protected by copyright. No part of this publication may be reproduced in any form or by any means, including photocopying, or utilized by any information storage and retrieval system without written permission from the copyright owner, except for brief quotation in critical articles or reviews.

[For detailed information on the availability of NCRP publications see page 448.]

Preface The search for solutions to the challenges posed by the need for long-term disposal and isolation of low-level radioactive waste has been long and complex. The Low-Level Radioactive Waste Policy Act, passed in 1980 and amended in 1985, specified that the disposal of most low-level waste not generated at U.S. Department of Energy sites is the responsibility of states or State Compacts. A critical factor in the process of determining acceptable disposal practices for low-level waste at any site is a demonstration of compliance with regulatory performance objectives. NCRP was asked to evaluate current approaches to performance assessment for near-surface disposal facilities for low-level radioactive waste, and Scientific Committee 87-3 was established to prepare a report on this subject. This Report provides a review of concepts underlying performance assessments of near-surface disposal facilities for low-level radioactive waste and approaches to conducting such assessments. This review includes discussions on the nature and scope of performance assessment, accepted approaches to conducting all aspects of a performance assessment, and unresolved issues in conducting performance assessments and applying the results. The Report also discusses a number of policy issues that affect conduct of performance assessment. Examples of these issues include the time period for complying with performance objectives, application of drinking water standards, and interpretation of performance objectives for compliance purposes. It is not the objective of this Report to present recommendations for resolution of policy issues, although the importance of such issues and other social, political and economic factors is recognized. Serving on the Committee were:

Chairmen David C. Kocher (1999–2006) SENES Oak Ridge, Inc. Oak Ridge, Tennessee

Matthew W. Kozak (1992–1999) Monitor Scientific LLC Richland, Washington

iii

iv / PREFACE Members William E. Kennedy, Jr. Dade Moeller & Associates, Inc. Richland, Washington

Roger R. Seitz Bechtel BWXT Idaho Scoville, Idaho

Vern Rogers* Rogers & Associates Engineering Corporation Salt Lake City, Utah

Terrence Sullivan Brookhaven National Laboratory Upton, New York

NCRP Secretariat E. Ivan White, Staff Consultant Cindy L. O’Brien, Managing Editor David A. Schauer, Executive Director

The Council is grateful for the financial support provided by the U.S. Department of Energy and the U.S. Nuclear Regulatory Commission at various times during the preparation of this Report. The Council also wishes to express its appreciation to the Committee members for the time and effort devoted to the preparation of this Report. Thomas S. Tenforde President

*deceased

Contents Preface . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . iii Executive Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1 1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11 1.1 Purpose of Report . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11 1.2 Scope of Report . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12 1.3 Related NCRP Recommendations . . . . . . . . . . . . . . . . 14 2. Definition and Principles of Performance Assessment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16 2.1 Nature of Performance Assessment . . . . . . . . . . . . . . . 16 2.2 Definition of Performance Assessment . . . . . . . . . . . . 18 2.3 General Principles of Performance Assessment . . . . . 20 2.3.1 Performance Assessment as an Iterative Process . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21 2.3.2 Performance Assessment as a Decision Tool . 22 2.3.3 Uncertainty in Results of Performance Assessment . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23 2.3.4 Integration and Interpretation of Results . . . . 23 2.3.5 Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24 2.4 Balance Between Conservatism and Realism in Performance Assessment . . . . . . . . . . . . . . . . . . . . . . . 25 3. Context for Performance Assessment . . . . . . . . . . . . . . 28 3.1 Definition of Low-Level Radioactive Waste . . . . . . . . . 28 3.1.1 Earliest Descriptions of Low-Level Waste . . . 28 3.1.2 Current Definition of Low-Level Waste . . . . . 29 3.2 Sources and Properties of Low-Level Waste . . . . . . . . 34 3.3 ICRP Recommendations on Disposal of Radioactive Waste . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 35 3.3.1 General Recommendations on Radiation Protection . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 35 3.3.2 General Policy on Application of Protection Principles to Radioactive Waste Disposal . . . . 36 3.3.3 Application of Protection Principles to Disposal of Solid Radioactive Wastes . . . . . . . 37 3.3.4 Discussion of ICRP Recommendations . . . . . . 45 v

vi / CONTENTS 3.4

3.5

Requirements for Near-Surface Disposal of Low-Level Waste . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 46 3.4.1 Authorized Disposal Systems . . . . . . . . . . . . . . 47 3.4.1.1 Legal and Regulatory Specifications . 47 3.4.1.2 Historical Development of Disposal Technologies . . . . . . . . . . . . . . . . . . . . 48 3.4.2 Requirements for Protection of the Public . . . . 49 3.4.2.1 Licensing Criteria Established by NRC . . . . . . . . . . . . . . . . . . . . . . . . 49 3.4.2.2 Requirements Established by DOE . . 52 3.4.2.3 Implementation of the ALARA Requirement . . . . . . . . . . . . . . . . . . . . 56 3.4.2.4 EPA Views on Requirements for Disposal of Low-Level Waste . . . . . . . 57 3.4.2.5 Implications of Performance Objectives . . . . . . . . . . . . . . . . . . . . . . 59 3.4.2.6 Requirements of States and State Compacts. . . . . . . . . . . . . . . . . . . . . . . 61 3.4.3 Unresolved Issues in Performance Objectives for Low-Level Waste Disposal . . . . . . . . . . . . . 61 3.4.3.1 Time Period for Compliance. . . . . . . . 62 3.4.3.2 Inclusion of Doses Due to Radon . . . . 63 3.4.3.3 Performance Objective for Protection of Groundwater. . . . . . . . . . . . . . . . . . 64 3.4.3.4 Interpretation of Performance Objectives for Compliance Purposes . 66 3.4.4 Other Approaches to Regulating Waste Disposal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 67 3.4.4.1 Approaches to Regulating Radioactive Waste Disposal . . . . . . . . 67 3.4.4.2 Approach to Regulating Disposal of Hazardous Chemical Waste . . . . . . . . 69 3.4.5 Requirements for Protection of the Environment . . . . . . . . . . . . . . . . . . . . . . . . . . . 73 Other Concepts in Performance Assessment . . . . . . . . 74 3.5.1 Institutional Controls . . . . . . . . . . . . . . . . . . . . 74 3.5.1.1 Active Institutional Controls . . . . . . . 74 3.5.1.2 Passive Institutional Controls . . . . . . 75 3.5.2 Model Validation and Confidence in Model Outcomes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 76

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3.5.2.1 Quality Assurance . . . . . . . . . . . . . . . 78 3.5.2.2 Model Calibration . . . . . . . . . . . . . . . 79 3.5.2.3 Evaluation of Conservative Bias . . . . 80 Concept of Reasonable Assurance . . . . . . . . . . 81 3.5.3.1 Description and Interpretation of Reasonable Assurance . . . . . . . . . . . . 83 3.5.3.2 An Approach to Achieving Reasonable Assurance of Compliance. . . . . . . . . . 84

4. Framework for Performance Assessment . . . . . . . . . . . 88 4.1 Data Collection, Conceptual Models, and Mathematical Models . . . . . . . . . . . . . . . . . . . . . . . . . . 89 4.1.1 Data Collection . . . . . . . . . . . . . . . . . . . . . . . . . 90 4.1.2 Development of Conceptual Models . . . . . . . . . 91 4.1.3 Selection and Implementation of Mathematical Models . . . . . . . . . . . . . . . . . . . . 93 4.2 Process for Conducting Performance Assessments . . . 95 4.2.1 Historical Perspective . . . . . . . . . . . . . . . . . . . 96 4.2.2 General Process for Conduct of Performance Assessment . . . . . . . . . . . . . . . . . . . . . . . . . . . 100 4.2.2.1 Description of Context for Performance Assessment. . . . . . . . . 101 4.2.2.2 Description of Disposal System. . . . 103 4.2.2.3 Development and Justification of Scenarios . . . . . . . . . . . . . . . . . . . . . 103 4.2.2.4 Formulation and Implementation of Models. . . . . . . . . . . . . . . . . . . . . . . . 104 4.2.2.5 Conduct of Calculations (Consequence Analysis) . . . . . . . . . . . . . . . . . . . . . . 105 4.2.2.6 Interpretation of Results . . . . . . . . . 106 4.2.2.7 Modifications of Assessment . . . . . . 107 4.2.2.8 Iterations of Performance Assessment . . . . . . . . . . . . . . . . . . . . 108 4.2.2.9 Summary . . . . . . . . . . . . . . . . . . . . . 109 5. Performance Assessment Models . . . . . . . . . . . . . . . . . 110 5.1 General Approach to Modeling of Disposal Systems . 110 5.1.1 Decoupling and Simplifying an Analysis . . . 113 5.1.2 Analysis by Modules . . . . . . . . . . . . . . . . . . . . 114 5.1.3 Analysis of Time Dependence . . . . . . . . . . . . 117 5.1.4 Organization of Section . . . . . . . . . . . . . . . . . 117

viii / CONTENTS 5.2

5.3

5.4

Cover Performance and Infiltration . . . . . . . . . . . . . . 117 5.2.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . 117 5.2.2 Types of Covers . . . . . . . . . . . . . . . . . . . . . . . . 121 5.2.3 Degradation of Covers . . . . . . . . . . . . . . . . . . 123 5.2.4 Approaches to Estimating Infiltration . . . . . . 125 5.2.5 Summary and Conclusions . . . . . . . . . . . . . . . 129 Performance of Concrete Barriers . . . . . . . . . . . . . . . . 130 5.3.1 General Approach to Modeling of Concrete Barriers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 132 5.3.2 Water Flow Through Concrete . . . . . . . . . . . . 132 5.3.3 Degradation of Concrete . . . . . . . . . . . . . . . . . 134 5.3.3.1 Sulfate Attack . . . . . . . . . . . . . . . . . . 135 5.3.3.2 Freeze/Thaw Cycling . . . . . . . . . . . . 136 5.3.3.3 Calcium Leaching . . . . . . . . . . . . . . . 136 5.3.3.4 Alkali-Aggregate Reaction . . . . . . . . 137 5.3.3.5 Corrosion of Reinforcing Steel . . . . . 137 5.3.3.6 Combination of Reactions . . . . . . . . 139 5.3.4 Application of Models . . . . . . . . . . . . . . . . . . . 140 5.3.5 Example Analyses of Long-Term Performance of Concrete Barriers . . . . . . . . . . . . . . . . . . . . 140 5.3.6 Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 143 Source Term . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 143 5.4.1 Inventories of Radionuclides . . . . . . . . . . . . . 144 5.4.2 Radionuclide Release Rates (Source Term) . . 147 5.4.3 Disposal Facility Concepts . . . . . . . . . . . . . . . 153 5.4.4 Waste Containers . . . . . . . . . . . . . . . . . . . . . . 155 5.4.5 Waste Forms . . . . . . . . . . . . . . . . . . . . . . . . . . 158 5.4.5.1 Waste-Form Performance: Aqueous Phase . . . . . . . . . . . . . . . . . . . . . . . . . 160 5.4.5.1.1 Surface Rinse with Partitioning. . . . . . . . . . . 160 5.4.5.1.2 Diffusion-Controlled Release . . . . . . . . . . . . . . 162 5.4.5.1.3 Dissolution (Constant) Release . . . . . . . . . . . . . . 165 5.4.5.1.4 Solubility-Limited Release . . . . . . . . . . . . . . 166 5.4.5.2 Waste-Form Performance: Gas Phase . . . . . . . . . . . . . . . . . . . . . . . . . 167 5.4.5.3 Ingrowth of Radionuclides . . . . . . . . 167

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Transport in Disposal Facility . . . . . . . . . . . . 168 5.4.6.1 Aqueous-Phase Transport . . . . . . . . 168 5.4.6.2 Gas-Phase Transport . . . . . . . . . . . . 170 5.4.7 Interfaces with Other Performance Assessment Models . . . . . . . . . . . . . . . . . . . . . 171 5.4.8 Source-Term Issues . . . . . . . . . . . . . . . . . . . . 171 5.4.8.1 Radionuclide Inventory Issues . . . . 172 5.4.8.1.1 Unit Source Term. . . . . . 172 5.4.8.1.2 Inaccurate Estimation of Inventories . . . . . . . . . . . 173 5.4.8.2 Waste-Container Issues . . . . . . . . . . 174 5.4.8.2.1 Insufficient Characterization of Containers . . . . . . . . . . . 175 5.4.8.2.2 Distributed Failure of Containers . . . . . . . . . . . 175 5.4.8.3 Waste-Form Issues . . . . . . . . . . . . . . 175 5.4.8.3.1 Changes in Waste Types and Characteristics . . . . 176 5.4.8.3.2 Insufficient Waste-Form Characterization . . . . . . 176 5.4.8.3.3 Insufficient Data on Release Rates . . . . . . . . . 177 5.4.8.3.4 Homogeneity of Wastes . 177 5.4.8.3.5 Issues of Geochemistry and Solubility . . . . . . . . . 178 5.4.8.4 Issues of Radionuclide Transport . . 179 5.4.8.4.1 Steady-State Flow . . . . . 179 5.4.8.4.2 Uniform Flow Fields . . . 179 5.4.8.4.3 Role of Geochemistry in Transport . . . . . . . . . . . . 180 5.4.8.4.4 Role of Microbial Processes . . . . . . . . . . . . 181 5.4.8.4.5 Role of Colloids. . . . . . . . 181 5.4.9 Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 182 Unsaturated Zone Flow and Transport . . . . . . . . . . . 183 5.5.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . 183 5.5.2 Interfaces with Other Performance Assessment Models . . . . . . . . . . . . . . . . . . . . . 185

x / CONTENTS 5.5.3

5.6

5.7

General Discussion of Unsaturated Zone Flow and Transport . . . . . . . . . . . . . . . . . . . . . . . . . 186 5.5.4 Data Requirements . . . . . . . . . . . . . . . . . . . . . 191 5.5.5 Modeling of Unsaturated Flow . . . . . . . . . . . . 194 5.5.6 Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 196 Aquifer Flow . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 197 5.6.1 Modeling of Aquifer Flow . . . . . . . . . . . . . . . . 197 5.6.2 Issues in Solving Flow Equation . . . . . . . . . . 200 5.6.2.1 Development of Steady-State Conditions . . . . . . . . . . . . . . . . . . . . . 200 5.6.2.2 Scale and Heterogeneity . . . . . . . . . 200 5.6.2.3 Boundary Conditions . . . . . . . . . . . . 202 5.6.2.4 Flow in Fractured Media . . . . . . . . . 203 5.6.3 Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 204 Radionuclide Transport in Groundwater and Surface Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 205 5.7.1 Phenomena That Influence Transport in Groundwater . . . . . . . . . . . . . . . . . . . . . . . . . . 206 5.7.1.1 Sorption. . . . . . . . . . . . . . . . . . . . . . . 206 5.7.1.2 Advection . . . . . . . . . . . . . . . . . . . . . 208 5.7.1.3 Diffusion . . . . . . . . . . . . . . . . . . . . . . 209 5.7.1.4 Dispersion . . . . . . . . . . . . . . . . . . . . . 210 5.7.2 Combination of Phenomena That Influence Transport in Groundwater . . . . . . . . . . . . . . . 213 5.7.3 Methods of Solution of Groundwater Transport Equation . . . . . . . . . . . . . . . . . . . . . 215 5.7.3.1 Analytical Solutions . . . . . . . . . . . . . 216 5.7.3.2 Green's Function (Semi-Analytical) Solutions . . . . . . . . . . . . . . . . . . . . . . 216 5.7.3.3 Finite-Element and Finite-Difference Solutions . . . . . . . . . . . . . . . . . . . . . . 217 5.7.3.4 Stream-Tube Solutions. . . . . . . . . . . 218 5.7.4 Boundary Conditions . . . . . . . . . . . . . . . . . . . 219 5.7.5 Modeling of Transport in Surface Water . . . . 221 5.7.5.1 Modeling of Discharges to Surface Water . . . . . . . . . . . . . . . . . . . . . . . . . 222 5.7.5.2 Modeling of Transport in Rivers and Streams . . . . . . . . . . . . . . . . . . . . . . . 223 5.7.5.3 Modeling of Transport in Lakes. . . . 224 5.7.5.4 Modeling of Transport in Sediment. 225

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5.7.6 Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 225 Atmospheric Transport Analysis . . . . . . . . . . . . . . . . 226 5.8.1 Models for Estimating Suspension of Particulates . . . . . . . . . . . . . . . . . . . . . . . . . . . 227 5.8.1.1 Modified Mass Loading Model. . . . . 227 5.8.1.2 Resuspension Factor Model . . . . . . . 229 5.8.1.3 Resuspension Rate Model . . . . . . . . 233 5.8.2 Release of Gases by Diffusion . . . . . . . . . . . . 235 5.8.3 Advective Transport . . . . . . . . . . . . . . . . . . . . 236 5.8.4 Atmospheric Transport Models . . . . . . . . . . . 237 5.8.5 Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 240 5.9 Biotic Transport . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 240 5.9.1 Background . . . . . . . . . . . . . . . . . . . . . . . . . . . 241 5.9.2 Biotic Transport Processes . . . . . . . . . . . . . . . 242 5.9.2.1 Transport Enhancement . . . . . . . . . 243 5.9.2.2 Intrusion and Active Transport . . . 243 5.9.2.3 Secondary Transport . . . . . . . . . . . . 244 5.9.3 Pathways of Human Exposure . . . . . . . . . . . 244 5.9.4 Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 245 5.10 Exposure Pathways and Radiological Impacts . . . . . 245 5.10.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . 245 5.10.2 General Recommendations . . . . . . . . . . . . . . 249 5.10.3 Exposure Scenarios for Off-Site Members of the Public . . . . . . . . . . . . . . . . . . . . . . . . . . 250 5.10.3.1 Definition of Environmental Conditions and Living Habits . . . . . 250 5.10.3.2 Exposure Scenarios for Different Release and Transport Pathways . . 251 5.10.4 Exposure Pathway Models . . . . . . . . . . . . . . . 252 5.10.4.1 General Considerations . . . . . . . . . . 252 5.10.4.1.1 Components of Exposure Pathway Models . . . . . . . 253 5.10.4.1.2 Multiplicative-Chain Models. . . . . . . . . . . . . . . 255 5.10.4.1.3 Specific-Activity Models 256 5.10.4.2 Models of Foodchain Pathways . . . . 258 5.10.4.2.1 Terrestrial Foodchain Pathways . . . . . . . . . . . . 258 5.10.4.2.2 Aquatic Foodchain Pathways . . . . . . . . . . . . 258 5.8

xii / CONTENTS 5.10.5 Selection of Model Parameter Values . . . . . . 259 5.10.6 Sources of Generic Data on Model Parameter Values . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 260 5.10.6.1 Dose Coefficients . . . . . . . . . . . . . . . 260 5.10.6.1.1 Internal Exposure . . . . . 261 5.10.6.1.2 External Exposure . . . . . 268 5.10.6.1.3 Summary of Dose Coefficients . . . . . . . . . . . 272 5.10.6.2 Usage Factors . . . . . . . . . . . . . . . . . . 272 5.10.6.3 Transfer Factors for Foodchain Pathways . . . . . . . . . . . . . . . . . . . . . . 274 5.10.6.3.1 Terrestrial Foodchain Pathways. . . . . . . . . . . . . 274 5.10.6.3.2 Aquatic Foodchain Pathways. . . . . . . . . . . . . 278 5.10.6.3.3 Summary of Transfer Factors. . . . . . . . . . . . . . . 279 5.10.7 Uncertainties in Dose Assessment Models . . 279 5.10.7.1 Uncertainties in Transfer and Usage Factors. . . . . . . . . . . . . . . . . . . . . . . . 279 5.10.7.2 Significance of Uncertainties in Transfer and Usage Factors. . . . . . . 280 5.10.7.3 Sources of Uncertainty in Dose Coefficients . . . . . . . . . . . . . . . . . . . . 282 5.10.7.4 Significance of Uncertainties in Dose Coefficients . . . . . . . . . . . . . . . . . . . . 282 5.10.8 Approaches to Estimating Risk . . . . . . . . . . . 283 5.10.9 Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 285 6. Inadvertent Human Intrusion . . . . . . . . . . . . . . . . . . . . 288 6.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 288 6.2 Role of Inadvertent Human Intrusion in Radioactive Waste Disposal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 291 6.2.1 Historical Perspective . . . . . . . . . . . . . . . . . . . 291 6.2.2 Regulatory Requirements for Near-Surface Disposal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 292 6.2.2.1 NRC Requirements. . . . . . . . . . . . . . 292 6.2.2.2 DOE Requirements . . . . . . . . . . . . . 294 6.3 Widely Used Scenarios for Inadvertent Intrusion . . . 295 6.3.1 Scenarios for Acute Exposure . . . . . . . . . . . . . 295

CONTENTS

6.4

6.5

6.6

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6.3.1.1 Construction Scenario . . . . . . . . . . . 295 6.3.1.2 Discovery Scenario . . . . . . . . . . . . . . 296 6.3.1.3 Drilling Scenario . . . . . . . . . . . . . . . 297 6.3.2 Scenarios for Chronic Exposure . . . . . . . . . . . 297 6.3.2.1 Agriculture Scenario . . . . . . . . . . . . 297 6.3.2.2 Resident, Nonagriculture Scenario . 299 6.3.2.3 Postdrilling Scenario . . . . . . . . . . . . 299 6.3.2.4 Groundwater Pathway for Chronic Intrusion Scenarios . . . . . . . . . . . . . 300 6.3.3 Comparison of Standard Scenarios for Inadvertent Intrusion . . . . . . . . . . . . . . . . . . . 301 6.3.4 Other Scenarios for Inadvertent Intrusion . . 302 Selection of Site-Specific Scenarios for Inadvertent Intrusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 304 6.4.1 Application of Widely Used Scenarios to Site-Specific Assessments . . . . . . . . . . . . . . . 304 6.4.2 Judgmental Factors in Selecting Exposure Scenarios . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 305 6.4.3 Summary of Principles of Scenario Selection 308 Inputs to Dose Analyses for Inadvertent Intruders . 308 6.5.1 Time of Occurrence of Intrusion . . . . . . . . . . 309 6.5.2 Radioactive Decay . . . . . . . . . . . . . . . . . . . . . . 309 6.5.3 Waste Dilution Following Disposal . . . . . . . . 310 6.5.4 Consideration of Radionuclide Transport . . . 312 Outputs of Dose Analyses for Inadvertent Intruders 313 6.6.1 Scenario Dose Conversion Factors . . . . . . . . 313 6.6.2 Waste Acceptance Criteria Based on Intruder Dose Assessment . . . . . . . . . . . . . . . . . . . . . . 314 Effects of Inadvertent Intrusion on Off-Site Releases of Radionuclides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 315 Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 317

7. Uncertainty, Sensitivity and Importance Analysis . . 320 7.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 320 7.2 Description of Importance Analysis . . . . . . . . . . . . . . 321 7.3 Purpose of Importance Analysis . . . . . . . . . . . . . . . . . 322 7.4 Nature of Uncertainties in Performance Assessment 323 7.4.1 Characteristics of Uncertainties . . . . . . . . . . 323 7.4.1.1 Type-A and Type-B Uncertainties . 324 7.4.1.2 Classification of Model Uncertainties . . . . . . . . . . . . . . . . . . 325 7.4.2 Uncertainty in Models . . . . . . . . . . . . . . . . . . 326

xiv / CONTENTS 7.4.3 7.4.4

7.5

7.6

7.7

7.8

Uncertainty in Future Site Conditions . . . . . 327 Uncertainty in Model Parameters . . . . . . . . . 328 7.4.4.1 Measurement Errors . . . . . . . . . . . . 328 7.4.4.2 Insufficient Data. . . . . . . . . . . . . . . . 329 7.4.4.3 Dependence of Measurements on Scale . . . . . . . . . . . . . . . . . . . . . . . 330 Mathematical Methods of Treating Uncertainty . . . . 330 7.5.1 Introduction to Mathematical Methods . . . . . 331 7.5.2 Propagation of Model Uncertainty . . . . . . . . . 332 7.5.3 Propagation of Future Uncertainty . . . . . . . . 334 7.5.4 Propagation of Parameter Uncertainty . . . . . 338 7.5.4.1 Deterministic Methods. . . . . . . . . . . 338 7.5.4.2 Probabilistic Methods. . . . . . . . . . . . 340 7.5.4.2.1 Monte-Carlo Analysis. . . 340 7.5.4.2.2 Perturbation Analysis . . 343 7.5.4.2.3 Possibilistic Analysis . . . 344 Role of Sensitivity Analysis in Importance Analysis . 344 7.6.1 Need for Sensitivity Analysis . . . . . . . . . . . . . 344 7.6.2 Methods of Parameter Sensitivity Analysis . 345 Application of Uncertainty Analysis to Importance Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 346 7.7.1 General Structure of Uncertainty Analysis . . 346 7.7.2 Evaluation of Different Methods of Uncertainty Analysis . . . . . . . . . . . . . . . . . . . 347 7.7.3 Evaluation of Approaches to Importance Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 350 Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 351

8. Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 353 8.1 Purpose and Scope of Performance Assessment . . . . 353 8.2 Basic Elements of Performance Assessment . . . . . . . 354 8.2.1 Development of Conceptual Models . . . . . . . . 355 8.2.2 Development and Application of Mathematical and Physical Models . . . . . . . . 356 8.2.3 Integration and Interpretation of Results . . . 357 8.3 Components of Performance Assessment Modeling . . 358 8.3.1 Cover Performance and Infiltration . . . . . . . . 358 8.3.2 Performance of Concrete Barriers . . . . . . . . . 359 8.3.3 Source Term . . . . . . . . . . . . . . . . . . . . . . . . . . 360 8.3.4 Unsaturated Zone Flow and Transport . . . . . 361 8.3.5 Aquifer Flow . . . . . . . . . . . . . . . . . . . . . . . . . . 363

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8.3.6

8.4 8.5 8.6 8.7

Radionuclide Transport in Groundwater and Surface Water . . . . . . . . . . . . . . . . . . . . . . . . . 364 8.3.7 Atmospheric Transport . . . . . . . . . . . . . . . . . 366 8.3.8 Biotic Transport . . . . . . . . . . . . . . . . . . . . . . . 366 8.3.9 Exposure Pathways and Radiological Impacts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 367 8.3.10 Overview of Components of Performance Assessment . . . . . . . . . . . . . . . . . . . . . . . . . . . 368 Inadvertent Human Intrusion . . . . . . . . . . . . . . . . . . 369 Uncertainty, Sensitivity and Importance Analysis . . 370 Iterative Nature of Performance Assessment . . . . . . 371 Reasonable Assurance of Compliance with Performance Objectives . . . . . . . . . . . . . . . . . . . . . . . . 372

Glossary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 373 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 396 The NCRP . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 439 NCRP Publications . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 448 Index . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 458

Executive Summary In the United States, low-level radioactive waste is defined as any radioactive waste arising from operations of the nuclear fuel cycle that is not classified as high-level waste (including spent fuel when it is declared to be waste), transuranic waste, or uranium or thorium mill tailings. Low-level waste is generated in many commercial, defense-related, medical, and research activities. Owing to its definition only by exclusion and its many sources, low-level waste occurs in a wide variety of physical and chemical forms, and it contains a wide range of concentrations of many different radionuclides. Most low-level waste, except relatively small volumes that contain high concentrations of radionuclides with half-lives on the order of 30 y or longer, is intended for disposal in facilities located on or near the ground surface. Decisions about acceptable nearsurface disposals of low-level waste are based in large part on the need to comply with regulatory requirements that have been established by the U.S. Nuclear Regulatory Commission (NRC) and the U.S. Department of Energy (DOE). These regulatory requirements include performance objectives that define allowable radiation exposures of the public at future times due to releases of radionuclides to the environment. In order to determine whether a particular facility complies with regulatory performance objectives for long-term protection of the public, a performance assessment of the facility must be conducted. As used in this Report: Performance assessment is an iterative process involving site-specific, prospective modeling evaluations of the postclosure time phase of near-surface disposal systems for low-level waste with two primary objectives: • to determine whether reasonable assurance of compliance with quantitative performance objectives can be demonstrated; and • to identify critical data, facility design, and model development needs for defensible and cost-effective licensing decisions and to develop and maintain operating limits (i.e., waste acceptance criteria). 1

2 / EXECUTIVE SUMMARY This definition emphasizes that performance assessment focuses primarily on a decision about compliance with performance objectives, rather than the much more difficult problem of predicting actual radiological impacts on the public at far future times. The purpose of this Report is provide a review of concepts underlying performance assessments of near-surface disposal facilities for low-level waste and approaches to conducting such assessments. This review includes discussions on the nature and scope of performance assessment, accepted approaches to conducting all aspects of a performance assessment, and unresolved issues in conducting performance assessments and applying the results. Challenges in conducting and defending performance assessments at specific sites also are emphasized. An understanding of general principles of performance assessment is important, because such understanding can lead to an efficient process and defensible product and can reduce the potential for misinterpretation of results. General principles of performance assessment are summarized as follows: • Performance assessment should be an iterative, flexible process of integrating modeling, data collection, and design activities in a manner that identifies those aspects of engineered and natural barriers in a disposal system of importance to a decision about compliance with regulatory performance objectives. The performance assessment process is important during all time phases of a facility from site selection and facility design through operations and postclosure monitoring and surveillance. • Performance assessment is a process that is intended to provide reasonable assurance of compliance with performance objectives; absolute assurance of compliance generally is not attainable by any means unless disposal of only very small quantities of radionuclides is allowed. • Since there is substantial uncertainty in models and important parameters used in performance assessment and some physical, chemical, and biological processes that affect the long-term performance of a disposal system may not be well understood, use of subjective scientific judgment is an essential aspect of performance assessment. Therefore, a variety of results that investigate the consequences of different plausible assumptions should be presented, rather than a single projected outcome. • An integration and interpretation of assumptions and results, in which conceptual models of the performance of a disposal

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system and their bases and the results of calculations are presented in a manner that reflects the many judgments involved and the importance of different aspects of an assessment to a demonstration of compliance with regulatory performance objectives, is a critical aspect of performance assessment. Quantitative performance objectives for near-surface disposal of low-level waste in NRC and DOE regulations are expressed in terms of limits on annual dose to members of the public, either equivalent dose to specific organs or tissues or effective dose equivalent. Therefore, projections of maximum concentrations of radionuclides in the environment at assumed locations of exposure are required. Although regulations are well established, a number of issues regarding performance objectives and their implementation are not yet fully resolved or are controversial. These include (1) the time period for compliance and the weight, if any, that should be given to projections of performance beyond the compliance period in determining acceptable disposals, (2) whether projected doses due to radon are included in performance objectives, which is potentially important in determining acceptable disposals of waste that contains radium, thorium, or uranium, (3) whether performance objectives should include a separate requirement for protection of groundwater resources in accordance with drinking water standards, which is important when drinking water standards would be more restrictive in determining allowable releases of many radionuclides than the existing performance objectives that apply to all exposure pathways combined, and (4) interpretation of performance objectives for compliance purposes—i.e., how highly uncertain results of performance assessments should be compared with fixed performance objectives in judging compliance. At most sites, movement of water is considered to be the most important means by which radionuclides may be released from a disposal facility and transported to locations where exposures of the public could occur. Even for the simplest types of near-surface facilities (e.g., an unlined trench with backfill and cap consisting of earthen materials), performance assessment requires an integration of results of a number of different types of models to provide an overall description of the performance of a disposal system. The usual approach to performance assessment is to model various components of the system separately and then link the components in a sequential fashion, with appropriate boundary and continuity conditions, to describe overall system performance. The different components that generally must be considered in a performance assessment are the following:

4 / EXECUTIVE SUMMARY • an analysis of cover performance and infiltration, the primary purpose of which is to estimate the flux of water (i.e., incident precipitation) that infiltrates through a natural or engineered cover system to locations of disposed waste or an engineered barrier (e.g., a concrete structure) above the waste; the performance of a cover system in inhibiting atmospheric releases of radionuclides in gaseous form also can be important at some sites and for some wastes; • an analysis of the performance of concrete barriers (e.g., vaults, modular canisters, or bunkers) that are used in many disposal facilities to enhance containment of low-level waste by (1) providing structural support for an earthen or engineered cover system, (2) delaying and inhibiting inflow of water to locations of disposed waste, (3) supplying additional adsorbing materials to retard movement of radionuclides into the surrounding environment, and (4) delaying and inhibiting release of radionuclides in leachate from a facility; • an analysis of the source term, the purpose of which is to estimate the rate of release of radionuclides from a disposal facility into the surrounding environment, either the vadose zone when releases are in liquid form or the atmosphere when releases are in gaseous form, by considering rates of release from waste forms and waste containers, transport within a disposal facility, and transport through any engineered barriers used in constructing the facility; • an analysis of flow and transport in the unsaturated (vadose) zone, which uses results of source-term modeling of liquid releases as input and provides estimates of releases into groundwater (zone of saturation); • an analysis of flow and transport in groundwater (zone of saturation), which uses results of modeling of flow and transport in the vadose zone, as well as any additional direct recharge to an aquifer that may occur due to runoff from a cover, as input and provides estimates of concentrations of radionuclides in groundwater at assumed locations of exposure; an analysis of flow and transport in surface water also may be required when groundwater into which releases occur discharges to the surface at locations close to a facility; • an analysis of atmospheric transport, which uses results of source-term modeling of gaseous releases or other means of transport of buried waste to the ground surface and release to the atmosphere as input and provides estimates

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of airborne concentrations of radionuclides at assumed locations of exposure; • an analysis of biotic transport, which considers actions of plants and animals that could affect transport of buried waste to assumed locations of exposure; biotic transport is not often considered explicitly in performance assessment, but it can serve to enhance release and transport of radionuclides and can be important at some sites; and • an analysis of exposure pathways and radiological impacts, which uses results of environmental transport models (i.e., groundwater or surface water flow and transport, atmospheric transport, or biotic transport models) that estimate concentrations of radionuclides in environmental media at assumed locations of exposure as input and provides estimates of transport through various exposure pathways to human receptors and the resulting radiation doses. Atmospheric and biotic transport often are considered to be unimportant compared with transport in water. However, these processes can be important in some environments and for some facility designs, and their importance generally should be evaluated in site-specific analyses. NRC and DOE regulations also require that near-surface disposal facilities provide protection of hypothetical inadvertent intruders who are assumed to come onto a disposal site after loss of institutional control and access disposed waste by such means as excavating to construct a foundation for a home or drilling. Protection of inadvertent intruders at sites licensed under NRC regulations is provided by the NRC’s waste classification system, which specifies limits on concentrations of radionuclides in Class-A, -B, and -C waste and technical requirements on disposal of waste in each class. At DOE sites, a site-specific assessment of potential impacts on inadvertent intruders is required for the purpose of establishing limits on concentrations of radionuclides; these limits can vary greatly depending on site conditions and the design of a facility. Under either regulations, compliance with a requirement to protect inadvertent intruders is based on analyses of potential radiological impacts in assumed intrusion scenarios. Standard scenarios that are often used are reviewed in this Report. Given that there are separate requirements to protect members of the public and inadvertent intruders, determinations of acceptable near-surface disposals of low-level waste essentially involve achieving a balance between acceptable releases of radionuclides beyond the site boundary and acceptable residual concentrations in a disposal

6 / EXECUTIVE SUMMARY facility after loss of institutional control. If excavation into waste and residence on a disposal site in a homesteader scenario are considered to be credible occurrences, criteria that define adequate protection of inadvertent intruders usually are more restrictive in determining acceptable disposals than performance objectives for protection of the public or the environment. Therefore, selection of credible intrusion scenarios can be very important in determining acceptable disposals in near-surface facilities. Performance assessments generally must consider uncertainties in models and parameter values, which result in uncertainty in results of modeling (e.g., projected doses to the public), and the sensitivity of model outputs to changes in assumptions and variations in parameters. Since the primary purpose of performance assessment is to provide a demonstration of compliance with regulatory performance objectives, a particular kind of uncertainty and sensitivity analysis, which is termed importance analysis, is emphasized in this Report. Importance analysis is an integration and interpretation of results obtained from the performance assessment process for the purpose of identifying assumptions and parameters which, when changed within credible bounds, can affect a decision about regulatory compliance. This type of analysis, which focuses on uncertainties and sensitivities that are important to a decision about compliance with performance objectives, is different from more traditional uncertainty and sensitivity analyses, which are concerned with representing uncertainty in the actual behavior of a disposal system and outcomes of waste disposal. An understanding of the distinction between importance analysis and traditional sensitivity and uncertainty analysis is important in conducting performance assessments efficiently and defending the results. Models of varying degrees of sophistication and complexity have been developed for all components of a performance assessment. However, detailed modeling of all aspects of the performance of a disposal system is beyond current capabilities and, indeed, is not required to achieve defensible results and robust decisions about the acceptability of waste disposals. Since there are a large number of radionuclides in low-level waste and a large number of potential pathways for transport and exposure, simple screening analyses to select for further analysis only those radionuclides and pathways that contribute significantly to projected doses to the public are an important initial step in making a performance assessment tractable. Once radionuclides and pathways are selected for further analysis, many stylized and simplifying assumptions normally are used in performance assessment in the interest of expediency. Examples of such assumptions include the following:

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• The potential importance of future climate change on infiltration and release and transport in water either is not considered or is modeled by assuming an abrupt change to an expected climate at some future time. • Infiltration through natural or engineered cover systems usually is modeled by assuming steady-state conditions, and transients that might occur as a result of episodic precipitation events are ignored. Failure of a cover system usually is modeled by assuming an instantaneous change or a series of instantaneous changes to natural conditions at some future time. • Degradation or failure of engineered barriers, either physical or chemical, usually is modeled by assuming an instantaneous change from an initial condition to a failed state at some future time or a constant rate of failure. • When there are highly heterogeneous distributions of radionuclides and a multiplicity of waste forms in a disposal facility, source terms usually are modeled by averaging radionuclide distributions over individual disposal units and assuming no more than a few idealized representations of waste forms. • The physical structure of unsaturated and saturated geologic media and their geochemical properties usually are assumed to be homogeneous and isotropic. • A graded approach to modeling flow and transport in the unsaturated (vadose) zone may be taken in which all or parts of the unsaturated zone are ignored (e.g., releases from a disposal facility are assumed to directly enter an underlying aquifer) or, less conservatively, a unit-gradient model that assumes steady-state flow and a flow rate equal to the infiltration rate is used. • The interface between models of flow in the vadose and saturated zones often is represented by simple boundary conditions (e.g., zero pressure head). • A linear sorption isotherm, described by the equilibrium solid/solution distribution coefficient (Kd), usually is assumed to represent all geochemical effects on transport in the disposal facility following release from a waste form and waste package and transport in the vadose and saturated zones. • All physical and chemical processes that affect release and transport of radionuclides often are assumed to be independent of radionuclide concentrations in waste, and the performance of the disposal system (e.g., projected dose to the public) is assumed to depend linearly on those concentrations.

8 / EXECUTIVE SUMMARY Challenges in conducting performance assessments and defending the results in a regulatory setting, which is tantamount to defending important assumptions, increase as the quantities of radionuclides that are intended for disposal in a facility increase and compliance with performance objectives cannot be demonstrated by using clearly conservative (pessimistic) models for highly uncertain components of performance. The lack of relevant site-specific data or data over time and spatial scales of importance is a frequent concern. Some of the challenges in modeling the performance of a disposal system when water is the medium in which releases are assumed to occur and a high level of performance of natural and engineered barriers is required to demonstrate compliance with performance objectives are summarized as follows: • Cover performance and infiltration: It can be difficult to justify that an engineered cover will perform as designed and constructed to control infiltration over time periods much beyond the period of institutional control. • Concrete barriers: There is little relevant data to predict the structural integrity and load-bearing capabilities of concrete over long time periods, so it is difficult to justify assumptions about structural integrity at times beyond several hundred years. Assumptions about infiltration through degraded concrete structures, including the relative importance of flow in fractures and pores, can be an important issue. • Source term: Inventories of radionuclides that often are expected to be the most important in releases to groundwater (e.g., 14C, 99Tc, and 129I) can be difficult to estimate. Modeling of releases can be challenging when radionuclide distributions are heterogenous and several waste forms are used. Although grout provides a homogeneous waste form for liquid wastes, modeling of long-term changes in hydrologic and geochemical properties of grout waste forms can be difficult when there are little relevant data and judgment must be relied on. • Unsaturated (vadose) zone flow and transport: Modeling at a detailed level is data-intensive and difficult to defend at specific sites owing to the complex, nonlinear relationships between moisture content, pressure (suction) head, and hydraulic conductivity and their dependence on soil type. A defensible conceptual model of flow and transport in unsaturated fractured rock is not yet available.

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• Saturated zone flow and transport: A groundwater velocity field is not directly measurable but must be generated using a model that is based on data on hydraulic head from monitoring wells and data on pump or core hydraulic conductivity tests. A velocity field so generated is non-unique, and multiple interpretations of data, with different effects on performance, may be reasonable at any site. Other issues of potential concern in modeling saturated zone flow include assumptions about boundary conditions, use of transient data to model steady-state conditions, the spatial scale and heterogeneity of a velocity field, modeling of flow in fractured rock, and the applicability (scaling) of laboratory data to field conditions. Issues in modeling transport include the simplistic nature of the Kd concept, justification of assumed Kd values at specific sites, treatments of diffusion and dispersion, and a lack of site-specific data on dispersivities. In contrast, modeling of atmospheric transport and exposure pathways and radiological impacts rarely is difficult or controversial, in part because the relevant processes are well understood and there are extensive studies to validate models normally used in performance assessments. A central issue that must be confronted in all performance assessments is whether simple and clearly conservative models should be used in demonstrating compliance with performance objectives or whether more complex and rigorous modeling should be undertaken in an effort to provide more realistic projections of outcomes at times far into the future. Either approach may be desirable for many reasons, and both have their difficulties. The point of view taken in this Report is that an appropriate balance between conservatism and more realistic approaches to performance assessment is largely a matter of judgment that should be applied on a site-specific basis. All performance assessments should attempt to incorporate some degree of realism to demonstrate an appropriate level of understanding of the long-term performance of disposal systems. The goal should be to provide a cost-effective and defensible assessment that is commensurate with the hazards posed by wastes that are intended for disposal at a specific site. At some sites with highly desirable characteristics, use of simple and conservative models for some aspects of system performance may not affect a decision about regulatory compliance. At other sites, however, efforts at more realistic modeling may be required. At any site, it is important to recognize that performance assessment is conducted to inform decisions about the

10 / EXECUTIVE SUMMARY acceptability of waste disposals, and that it is not necessary to obtain realistic projections of outcomes to render such decisions in a defensible manner. Although performance assessment involves a significant amount of subjective scientific judgment and there are important limitations in regard to predicting actual outcomes, these factors do not compromise the essential role of performance assessment in regulatory decision making.

1. Introduction Low-level radioactive wastes are generated in a variety of commercial, defense-related, medical, and research activities. Most low-level waste1 generated in the United States, except relatively small volumes that contain the highest concentrations of radionuclides, is intended for disposal in facilities located on or near the ground surface. This intention is based largely on generic analyses, such as those performed by the U.S. Nuclear Regulatory Commission (NRC, 1981a; 1982a), which indicated that near-surface facilities should be capable of providing adequate protection of public health and the environment at times far into the future. Decisions about acceptable near-surface disposals of low-level waste at specific sites are made on the basis of applicable laws and regulations. Such decisions take into account many scientific, technical, economic and social factors. Regulatory requirements that apply to disposal of low-level waste in near-surface facilities include performance objectives that define allowable radiation exposures of the public at future times. In order to determine whether a particular disposal facility complies with performance objectives, a performance assessment of the facility must be conducted. A performance assessment essentially is a prospective evaluation of potential radiation exposures of the public at times after a disposal facility is closed. 1.1 Purpose of Report The purpose of this Report is to provide a review of concepts underlying performance assessments of near-surface disposal facilities for low-level waste and approaches to conducting such assessments. This review includes discussions on the nature and scope of performance assessment, accepted approaches to conducting all aspects of a performance assessment, and unresolved issues in conducting performance assessments and applying the results. 1 In this Report, the term “low-level waste” is used to refer to low-level radioactive waste, and similarly with “high-level waste.” Since the terms “low-level waste” and “high-level waste” are not used to describe wastes that contain hazardous chemicals, their use to describe different radioactive wastes should not cause confusion.

11

12 / 1. INTRODUCTION Many discussions in this Report also apply to deep geologic repositories, which are intended for disposal of high-level and transuranic radioactive wastes. In particular, the conceptual foundations and general principles of performance assessment described in this Report are applicable to geologic repositories, even though models used in performance assessments for geologic repositories and applicable regulatory criteria may differ from those for nearsurface facilities. Discussions in this Report generally do not apply to disposal of large volumes of uranium mill tailings, since this activity is conducted in accordance with a different set of legal and regulatory requirements that do not call for prospective evaluations of long-term radiological impacts on the public in determining acceptable disposal practices at specific sites.2 This Report focuses on performance assessments of proposed or currently operating facilities for near-surface disposal of low-level waste. Assessments to evaluate potential radiological impacts of past disposal practices at inactive or abandoned sites for the purpose of determining whether there is a need for site remediation to mitigate those impacts are outside the scope of this Report. Although technical aspects of such assessments may have much in common with performance assessments discussed in this Report, remediation of past disposal practices is addressed under a different legal and regulatory framework, and the need to assess potential impacts of past disposal practices at times far into the future is not yet established. 1.2 Scope of Report This Report is intended to provide a general overview of performance assessment, and to identify sources of current information about performance assessment and the different scientific disciplines involved in conducting performance assessments. An important emphasis is the level of detail that has been shown by past experience to be required in performance assessments of specific disposal facilities. It is not the intent of this Report to provide a comprehensive review of all aspects of performance assessment that may be important at a particular facility. 2

Performance assessments to evaluate potential radiation exposures of the public at mill-tailings disposal sites were conducted by the U.S. Environmental Protection Agency in developing regulatory requirements in the form of design objectives that apply at all sites (EPA, 1982; 1983). The acceptability of mill-tailings disposal at any site then is demonstrated by meeting the design objectives.

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Discussions in this Report also consider a number of policy issues that affect conduct of performance assessment. Examples include the time period for complying with performance objectives, application of drinking water standards to protection of groundwater resources, and whether impacts of inadvertent human intrusion on the normal performance of a near-surface disposal facility are evaluated in demonstrating compliance with performance objectives. It is not the purpose of this Report, however, to present recommendations for resolution of any such issues. Similarly, although the importance of social, political and economic factors is recognized and discussed to a limited extent, this Report is not concerned with addressing those factors and integrating them into the process of developing and licensing new facilities. This Report is organized as follows. Sections 2 and 3 provide background information of importance to understanding discussions on technical approaches to performance assessment, including a definition of performance assessment and a discussion of goals and principles of performance assessment (Section 2) and discussion of the broader context for performance assessment (Section 3). Sections 4 and 5 present current views on suitable technical approaches to conducting performance assessments. Section 4 presents a conceptual framework for conducting performance assessments, including general discussions of a recommended approach to performance assessment, principles for treatment of uncertainty, and development of confidence in results as part of the performance assessment process. These discussions emphasize advantages of an iterative approach to performance assessment involving interactions among site characterization and data collection, facility design, and modeling activities. Section 5 discusses models and databases that are considered suitable for use in performance assessments of near-surface disposal facilities for low-level waste, as well as important sources of uncertainty in those models. The approach taken in this Report is to divide performance assessment into several modules, each of which represents a particular aspect of a disposal system (e.g., longevity of engineered barriers, release of radionuclides from a disposal facility, transport of radionuclides in the environment, pathways of human exposure). Specific computer codes that may be used in performance assessment are rarely discussed in this Report. A code is no more than a particular numerical representation of models embodied in it, and it is far more important to focus on selection of credible models than on particular implementations of those models. Sections 6 and 7 present discussions on topics that are important to performance assessment but do not fit in discussions in

14 / 1. INTRODUCTION Section 5 on modeling of particular aspects of a disposal system. These topics include assessments of inadvertent human intrusion (Section 6) and treatment of uncertainty and sensitivity by means of importance analysis (Section 7). Finally, Section 8 provides summary comments on performance assessment, with particular emphasis on technical challenges in conducting performance assessments. 1.3 Related NCRP Recommendations The National Council on Radiation Protection and Measurements (NCRP) previously issued three reports that are relevant to performance assessments of near-surface disposal facilities for low-level waste. NCRP Report No. 76 (NCRP, 1984a) presents recommendations on models and databases for use in assessing radiation doses to the public following release of radionuclides to the atmosphere, surface water, or groundwater. Models of transport in the environment and exposure pathways discussed in Report No. 76 are relevant to performance assessment. However, some approaches to modeling environmental transport and exposure pathways discussed in that report may be more detailed than necessary or useful in performance assessment. For example, modeling of episodic releases and seasonal dependencies of exposures, which can be important in assessing doses resulting from actual or expected releases from operating facilities and in reconstructing doses from past releases, are not needed in prospective assessments of the performance of disposal facilities at far future times. NCRP Report No. 123 (NCRP, 1996a; 1996b) presents recommendations on screening models to assess impacts of releases of radionuclides to the environment, including screening models of atmospheric transport, transport in surface water, disposal of radionuclides in the ground, and transport in terrestrial and aquatic food chains. Those models are generic and are intended to be applied at any site. Screening involves use of simple models that employ clearly conservative (pessimistic) assumptions. Such models can be used to demonstrate that operating facilities comply with regulatory requirements or to eliminate unimportant radionuclides and exposure pathways from further consideration in a dose assessment. Screening models described in Report No. 123 may, in many cases, provide a suitable first iteration for related aspects of performance assessments of low-level waste disposal facilities, to be followed by use of more site-specific models of increasing sophistication. However, when generic screening models are used in performance assessment, including models recommended by NCRP,

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proper justification for their use at a disposal site of concern must be provided. Screening models for disposal of radionuclides in the ground, which include models for releases to groundwater and exposures to buried waste by several pathways assuming loss of institutional control at 10 y after facility closure, and transport of radionuclides in terrestrial and aquatic food chains could be particularly useful in performance assessment. Models of transport in surface water, which are discussed in this Report, also could be useful at some disposal sites. NCRP Report No. 129 (NCRP, 1999) presents recommendations on screening levels (concentrations) of radionuclides in surface soil. For each of several assumptions about future uses of contaminated land and assumptions about exposure pathways and pathway models in each land-use scenario, radionuclide-specific screening factors expressed as annual effective doses per unit activity concentration in surface soil [Sv (Bq kg)–1] are derived. The review and evaluation of exposure pathway models used to derive screening levels in surface soil essentially updates recommendations in NCRP Report No. 76 (NCRP, 1984a). In contrast to the approach to screening in NCRP Report No. 123 (NCRP, 1996a; 1996b), which uses point values of all parameters, the screening analysis in Report No. 129 incorporates assumptions about uncertainties in all parameters used in estimating annual effective doses. Median values and 95th percentiles of uncertain screening factors are derived, and the 95th percentiles are used to obtain screening levels in surface soil that correspond to an annual effective dose of 0.25 mSv. Thus, the intent is that at any site at which a particular land-use scenario is considered appropriate, application of the derived screening levels should ensure that, with a high level of confidence, annual doses to members of the public who might occupy contaminated land would be less than the assumed dose criterion. Screening levels developed in Report No. 129 could be particularly useful in performance assessments in regard to assessing doses following dispersal of radionuclides over the land surface (e.g., by irrigation with contaminated groundwater) and in assessing doses from inadvertent intrusion into a facility.

2. Definition and Principles of Performance Assessment This Section presents a general discussion of objectives and principles of performance assessment, and the definition of performance assessment used in this Report. The role of performance assessment in radiation protection of the public is discussed further in Section 3.3. 2.1 Nature of Performance Assessment Performance assessment is concerned with prospective evaluations of waste disposal systems. A disposal system is comprised of multiple engineered and natural barriers (ICRP, 1998), which are intended to inhibit movement of radionuclides from locations of waste emplacement into the general environment beyond the boundary of a disposal site and to inhibit intrusion into waste by water, plants, burrowing animals, and humans. Examples of engineered barriers include concrete disposal vaults, impermeable caps above buried waste that are constructed with man-made or natural materials, and grouting of waste to control ingress and egress of water and release of radionuclides. An important natural barrier at any near-surface disposal site is the ability of native soils to sorb radionuclides and, thus, retard their migration. Performance assessment is an essential activity in developing disposal facilities and gaining approval by regulatory authorities, because it provides the only available link between measurable properties of a disposal facility or waste and potential long-term radiological impacts of waste disposal on the public. This link is crucial because, unlike many more familiar engineered systems, there usually are not intuitively evident relationships between measurable properties of a disposal system and consequences of waste disposal. For instance, one cannot draw general conclusions that a geologic stratum with high permeability is either favorable or unfavorable as a disposal site. Performance assessment is used to integrate available information about the long-term behavior of a disposal facility for the purpose of obtaining a defensible 16

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conclusion regarding the ability of the facility to protect the public in accordance with applicable regulatory criteria. Given the complexities associated with assessments of natural and engineered disposal systems, a critical part of this integration process involves iterative feedback to identify data, modeling and design needs for the process of gaining regulatory approval of a facility. Performance assessment also can be important after regulatory authorities have approved a disposal facility. Additional data may be gathered during disposal operations or after facility closure, and a performance assessment may need to be maintained and updated until institutional control is relinquished. Thus, performance assessment should be viewed as a tool for risk management, not just as a means of gaining approval of a facility by regulatory authorities. Unlike safety analyses of nuclear reactors, airplanes and bridges, for example, that address relatively short-term behavior of engineered systems on the basis of large amounts of empirical data, performance assessment addresses the long-term behavior of complex systems for which relatively little empirical data are available.3 It is, therefore, important to distinguish performance assessment from other, more standard engineering problems. The primary distinction is that performance assessment relies more heavily on judgments, which are necessitated by a lack of observations of the long-term performance of disposal facilities. Indeed, judgments are essential and form the basis of calculations of long-term performance. There needs to be a frank acknowledgment that required judgments do not always have a firm basis in relevant measurements on real systems. Some judgments will be based on established principles of science and engineering, but others will be little more than informed opinion. Performance assessments commonly involve a combination of simple screening models and more complex analyses. Screening models may be used, for example, to eliminate unimportant radionuclides and transport or exposure pathways from further consideration in an assessment or to address in a clearly conservative 3Experience

with low-level waste disposal facilities that have operated in the past can provide useful information for performance assessments of currently operating or planned facilities. For example, observations at historical facilities that did not perform adequately can provide information on designs and environmental conditions that are not likely to prove satisfactory. However, the number of such facilities is limited, and most of them did not include engineered barriers and waste forms of the type frequently used in currently operating or planned facilities.

18 / 2. DEFINITION AND PRINCIPLES OF PERFORMANCE ASSESSMENT manner aspects of the performance of a disposal system for which little information is available and modeling is difficult, such as transport of radionuclides in unsaturated soil. In conducting a performance assessment, considerable effort often is spent in developing an appropriate balance between use of realistic and conservative models; this important issue is discussed further in Section 2.4. The key to conducting performance assessments is to structure an analysis so that it is defensible on the basis of available information on the long-term performance of a disposal facility. 2.2 Definition of Performance Assessment The definition of performance assessment in this Report emphasizes several concepts that are important to the conduct, interpretation and use of performance assessments, and that distinguish performance assessment from typical engineering analyses. As used in this Report: Performance assessment is an iterative process involving site-specific, prospective modeling evaluations of the postclosure time phase of near-surface disposal systems for low-level waste with two primary objectives: • to determine whether reasonable assurance of compliance with quantitative performance objectives can be demonstrated; and • to identify critical data, facility design, and model development needs for defensible and cost-effective licensing decisions and to develop and maintain operating limits (i.e., waste acceptance criteria). This definition emphasizes that, for purposes of this Report, performance assessment is a process that focuses primarily on a decision about compliance with regulatory requirements, rather than the much more difficult problem of predicting actual outcomes of waste disposal (i.e., actual radiological impacts on the public and the environment). The following paragraphs summarize key aspects of this definition. Further discussions are contained in the following section. The term “iterative process” refers to an expectation that performance assessment probably will require two or more sequential sets of calculations as additional data are collected during site characterization activities and, perhaps, during facility operations and postclosure monitoring of a facility. In general, an iterative

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approach helps to minimize a tendency for performance assessment to become focused on details of models without considering the relative importance of specific aspects of an assessment to the primary concern of demonstrating compliance with regulatory requirements; it also emphasizes that performance assessment is more than a modeling exercise. Modeling of disposal systems must be conducted in coordination with data collection, facility design activities, and long-term monitoring. When applied efficiently, the performance assessment process provides input to management decisions regarding needs for additional data collection, design activities, and monitoring. In general, iteration of performance assessments is necessary when new information becomes available that could affect a licensing decision, but care should be taken to avoid unnecessary iterations that would impede the licensing process.4 The term “site specific” refers to the need to base performance assessments on relevant data for a facility being considered. Generic performance assessments have been useful in developing regulations, investigating the cost-effectiveness of various designs, and building confidence that near-surface disposal facilities at well-chosen sites should be protective of public health and the environment, but models used in generic assessments should not be applied to specific sites and facility designs without proper justification. The optimum approach to collecting site-specific data depends primarily on conditions at a site, the design of a facility, and the need for additional data in demonstrating compliance with performance objectives. The term “postclosure” emphasizes that performance assessment is concerned only with the consequences of waste disposal following emplacement of all waste and closure of a facility.5 Routine and accidental releases of radionuclides prior to closure are better addressed using standard safety assessments of site operations and monitoring. However, performance assessment can be helpful in guiding monitoring activities during the preclosure time phase, and such monitoring can provide useful information on postclosure performance. 4In this Report, the term “licensing” refers generally to a process of obtaining approval of a disposal facility by any regulatory authorities; it does not necessarily refer only to approval by NRC or an Agreement State. 5This may not be the case at a geologic repository for disposal of spent nuclear fuel and high-level wastes, because such a facility may remain open for decades after waste emplacement is complete.

20 / 2. DEFINITION AND PRINCIPLES OF PERFORMANCE ASSESSMENT The term “prospective modeling evaluations” is used rather than “modeling predictions” to emphasize that performance assessment is directed primarily at building sufficient understanding of the long-term behavior of a disposal system to make a defensible decision about compliance with performance objectives. Such an understanding is achieved by identifying those assumptions about the performance of engineered and natural barriers in a disposal system that are most important to obtaining the projected outcome and then showing that plausible changes in those assumptions would not affect a decision about compliance. The term “reasonable assurance” emphasizes the inexact nature of performance assessment and the crucial role of judgment in conducting performance assessments and in evaluating results. Further discussion of the important concept of reasonable assurance is provided in Section 3.5.3. The term “operating limits” is included in the definition to emphasize that results of performance assessment are used to identify acceptable operating conditions at a disposal facility, especially waste acceptance criteria in the form of limits on concentrations or inventories of radionuclides and requirements on physical and chemical properties of waste forms. Operating limits may change during the period of waste disposal in response to improved understanding of system behavior or changes in facility design, properties of waste, waste containers, waste emplacement, or the closure concept. Any changes in operating limits must be supported by revisions of a performance assessment to incorporate such new information. 2.3 General Principles of Performance Assessment This Section presents a discussion of general principles of performance assessment, based on the definition given in the previous section. It is important to recognize that performance assessment is essential to management of a disposal site and waste disposal operations. An understanding of principles of performance assessment can reduce the potential for misinterpretation of results, and can lead to an efficient process and defensible product. General principles of performance assessment discussed in this Section are identified as follows: • Performance assessment should be an iterative, flexible process of integrating modeling, data collection, and design activities in a manner that identifies those aspects of

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engineered and natural barriers in a disposal system of importance to a decision about compliance with performance objectives. • Performance assessment is a process intended to provide reasonable assurance of compliance with performance objectives. • Since models and parameters used in performance assessment are uncertain and some processes that affect system performance may not be well understood, a variety of results should be presented rather than a single projected outcome. • An integration and interpretation of assumptions and results, in which bases for conceptual models and results of calculations are presented in a manner that reflects the judgments involved and the importance of different aspects of an assessment to a licensing decision, is a critical aspect of performance assessment. In general, the performance assessment process provides a means of building confidence in judgments and models used to determine whether reasonable assurance of compliance with performance objectives can be obtained. 2.3.1

Performance Assessment as an Iterative Process

Characterization of performance assessment as a process, rather than a set of calculations, is important. An early view was that site characterization should be completed and a conceptual facility design developed, after which a performance assessment would be conducted to support a license application (Starmer et al., 1988). However, experience has shown that efficient application of the performance assessment process involves substantial interaction between modeling and data collection and facility design activities. These interactions occur during the period prior to licensing and during facility operations and are dynamically linked. That is, site data and information on facility design are needed to make modeling assumptions and to assign parameter values, evaluation of modeling results can lead to an identification of additional data needs or design changes, and these in turn can lead to use of altered assumptions and parameter values in modeling. Introduction of unanticipated wastes can lead to changes in modeling assumptions and facility design. In general, feedback between different activities identifies aspects of the site, design or models for which altered

22 / 2. DEFINITION AND PRINCIPLES OF PERFORMANCE ASSESSMENT assumptions are needed to improve projected performance in a defensible manner. Thus, performance assessment is a management tool to be used throughout the preclosure phase of facility operations. Performance assessment also is a useful tool for risk management during the postclosure time phase. It can be used to determine needs for institutional control at a site, and it can be used to guide monitoring activities that normally are undertaken after facility closure. Such monitoring activities can provide additional information on the validity of a performance assessment in supporting a licensing decision. Important aspects of data collection, facility design, and modeling are best identified using a flexible, iterative process that allows feedback from each activity to be incorporated in subsequent iterations of other activities (Case and Otis, 1988; DOE, 1985; IAEA, 1997a; 1999; Kozak, 1994a; Kozak et al., 1993; NRC, 2000; Seitz et al., 1992a). Efficiency is improved by starting with simple analyses and using findings of those analyses to identify areas that require more detailed consideration in subsequent analyses. Sensitivity analysis is an important contributor to this feedback. Sensitivity analysis provides an increased understanding of particular aspects of a disposal system that influence overall performance, and it helps to provide justification for those aspects for which additional effort would provide the most benefit to a performance assessment and an appropriate basis for a licensing decision. 2.3.2

Performance Assessment as a Decision Tool

Performance assessment is an essential tool for regulatory decision making. Performance assessment is used to identify conditions at a disposal facility that support a finding of reasonable assurance of compliance with performance objectives. An understanding of how this use of performance assessment differs from its application to the more difficult problem of predicting the actual long-term performance of disposal systems is critical to the conduct of performance assessment and associated data-collection activities and to proper interpretation of results. In the context of performance assessment, it is desirable to develop conceptual models of the long-term performance of disposal systems that bound the range of reasonably foreseeable conditions. Thus, performance assessment can be used, in an inverse sense, to identify conditions that may cause performance objectives to be exceeded. This places the emphasis on defending why such conditions are not likely to occur or on needed changes in facility design

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to minimize the potential for such conditions to occur. Because of inherent uncertainties in any performance assessment, judgment will be necessary in assessing the defensibility of conceptual models. 2.3.3

Uncertainty in Results of Performance Assessment

Models and parameters used in performance assessment are characterized by varying degrees of uncertainty. Thus, performance assessment generally should provide more than a single result in demonstrating compliance with performance objectives, with different results developed on the basis of a variety of plausible assumptions. For example, although a single set of results may provide the primary basis for a comparison with performance objectives, additional results should be provided in the context of a sensitivity and uncertainty analysis to lend credence to the demonstration of compliance. The type of uncertainty of primary concern to this Report is uncertainty of importance to a licensing decision, rather than uncertainty in the actual outcome (projected dose). To emphasize this distinction, this Report uses the term “importance analysis” (Section 7). Importance analysis is used to identify assumptions and parameter values in a performance assessment that have an impact on a decision regarding compliance with performance objectives. This information can then be used to identify areas where further data, design enhancements, or modeling are needed to reduce uncertainty (increase confidence) in a decision. Since results of performance assessment depend on assumptions, data and design of a facility, changes in any of these can result in changes in conclusions resulting from an analysis. 2.3.4

Integration and Interpretation of Results

Given the inexact nature of performance assessment and the need to consider a variety of conditions in an assessment and, thus, to produce a range of projected outcomes, it is logical to ask how those outcomes can be interpreted for comparison with fixed performance objectives. An essential but highly challenging aspect of performance assessment is an integration and presentation of results in a manner that facilitates this interpretation. The term “integration” refers to a need to summarize a variety of results and to identify assumptions and data that are important to a licensing decision, as determined from the analyses performed. The term “interpretation” refers to a need to present results and critical assumptions in a manner that provides the basis for

24 / 2. DEFINITION AND PRINCIPLES OF PERFORMANCE ASSESSMENT identifying conditions under which a disposal facility can be expected to comply with performance objectives. This part of a performance assessment will require substantial judgment, as do all other parts. Peer review can be a useful means of building confidence in assumptions and the defensibility of an analysis. 2.3.5

Summary

Performance assessment is most efficient when conducted in an iterative manner, as shown conceptually in Figure 2.1. These steps in an iterative approach are developed more fully into procedural steps in Section 4. At this stage of the discussion, the important point is that performance assessment produces an evolution of ideas about the behavior of a disposal system by an integration of observations, conceptualizations and interpretations of technical analyses. Analyses become more refined as additional data, information on facility design, detailed models, or better supported assumptions are used. Data collection, design, and modeling activities are interdependent parts of a process that require regular feedback among the multiple disciplines involved. Results of each iteration of a performance assessment should provide feedback to help identify critical data and design needs. Performance assessment should be viewed as a process, the primary purpose of which is to determine whether there is reasonable

Fig 2.1. Conceptual representation of iterative approach to conduct of performance assessment; process normally begins with observation and development of conceptual models.

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assurance that doses to members of the public and impacts on the environment that would result from disposal of waste in a particular facility will be below applicable regulatory requirements. A defensible performance assessment includes the use of results obtained throughout different phases of the process to demonstrate an improved understanding of the importance of assumptions regarding a site and engineered features with respect to conclusions obtained from an analysis of long-term performance. Finally, the most important task in performance assessment is to integrate and interpret results in a manner that facilitates their use in providing justification for a decision about compliance of a disposal facility with applicable regulatory requirements. Acceptance of the need for judgment in conducting performance assessment and in interpreting results is an essential aspect of the licensing process. 2.4 Balance Between Conservatism and Realism in Performance Assessment A central issue that must be confronted in all performance assessments is whether simple and clearly conservative models should be used to describe various aspects of the long-term performance of a disposal system or whether more complex and rigorous modeling should be undertaken in an effort to provide more realistic projections of outcomes at times far into the future. The following discussion describes the point of view on this issue taken in this Report. Incorporation of realism in performance assessment is desirable for many reasons. For example: more realistic assessments could be used to justify that disposal of greater amounts of waste in licensed near-surface facilities would be safe, which could alleviate needs for more costly disposal options; more realistic assessments could more accurately convey to the public and other stakeholders the expected magnitude of doses or risks at far future times, which could alleviate unwarranted concerns about radiological impacts of waste disposal; more realistic assessments are potentially important in comparing alternative disposal systems or facility designs; efforts at realistic assessments may lead to significant advances in modeling and development of improved databases to support modeling activities; and too great a reliance on conservative models in evaluating the performance of operating or planned near-surface disposal facilities may provide precedents that, when applied in evaluating needs for cleanup of radioactively contaminated sites, could lead to costly but unwarranted remediation.

26 / 2. DEFINITION AND PRINCIPLES OF PERFORMANCE ASSESSMENT However, a desire for realism in performance assessments must be tempered by a realization of difficulties in taking this approach. A fundamental difficulty is that results of performance assessments cannot be tested by comparison with observed outcomes when projected impacts of waste disposal often occur many generations into the future.6 In addition: realistic assessments of all aspects of the long-term performance of a disposal facility would require extensive amounts of site-specific data that generally are not available when planning for a new facility begins; collection of required site-specific data would likely be time consuming and unreasonably expensive,7 or it may not be possible to obtain some data that would be required for rigorous modeling; and a demonstrated ability to realistically model some phenomena, such as flow and transport in unsaturated soil, under any conditions that might be encountered in the environment is not yet in hand. If attempts at realistic assessments are to be viewed as credible, they must be accompanied by full disclosure of uncertainties in all aspects of an analysis, including uncertainties in the present state of knowledge of chemical and physical processes that occur in complex natural and engineered systems and uncertainties in model structures used to represent those processes, as well as uncertainties in model parameters. This is a formidable challenge for many aspects of the performance of waste disposal systems. These difficulties illustrate that determination of “realistic” outcomes of waste disposal is not solely an objective exercise. Rather, any approach to modeling requires a significant amount of subjective scientific judgment, and it is practically impossible on the basis of current knowledge to evaluate how those judgments would affect differences between projected and actual outcomes. Although use of clearly conservative models for at least some aspects of the performance of waste disposal systems can circumvent many of the difficulties with attempts at more realistic 6Observations of outcomes of past disposal practices are of limited use when older facilities were not located or designed in accordance with current practices, when observations have been carried out for no more than a few decades, or when observed outcomes resulted from unsatisfactory designs and environmental conditions (see Footnote 3 on page 17). 7The time and costs required to obtain extensive amounts of sitespecific data are more difficult to justify at near-surface disposal facilities for low-level waste than at geologic repositories for spent nuclear fuel and high-level waste or transuranic waste, given that there will be very few geologic repositories and that the latter types of waste usually pose much higher intrinsic hazards.

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assessments described above, this approach also is not without its challenges. For example: conservative models often are implemented using generic data, and there is a need to demonstrate that assumed models and data are conservative at specific sites; assumptions that would lead to conservative projections of outcomes may not be intuitively obvious for some aspects of system performance; and it may be difficult to explain to the public and other stakeholders that an expected outcome is substantially less than a projected outcome obtained using conservative models. In addition, use of conservative models in inappropriate ways can result in misleading comparisons of alternative disposal systems or facility designs. Perhaps the most important concern with overzealous use of conservative models is that near-surface disposal of some wastes would be foreclosed when projected outcomes are compared with performance objectives even though, in reality, disposal of those wastes would not compromise compliance with performance objectives or protection of the public. The point of view taken in this Report is that an appropriate balance between conservatism and more realistic approaches to performance assessment is largely a matter of judgment that should be applied on a site-specific basis, and that there is no single prescription that would be appropriate at all sites. Performance assessments generally should attempt to incorporate some degree of realism to demonstrate an appropriate level of understanding of the long-term performance of disposal systems. The goal at any site should be to provide a cost-effective and defensible assessment that would be commensurate with hazards posed by low-level wastes that are intended for disposal at that site. At well chosen sites with highly desirable characteristics, use of simple and conservative models for some aspects of system performance may not affect a decision on the acceptability of wastes that are intended for disposal. At other sites with less favorable characteristics, efforts at more realistic modeling may be required to demonstrate that wastes that are intended for disposal would be acceptable. At any site, it is important to keep in mind that performance assessment is conducted to inform a decision about the acceptability of waste disposals, and that it is not necessary to obtain realistic projections of outcomes to render such decisions in a defensible manner.

3. Context for Performance Assessment This Section discusses several topics that provide important context for the conduct of performance assessments of low-level waste disposal facilities. These topics include: (1) the definition of low-level waste in the United States; (2) sources and properties of low-level waste; (3) recommendations of the International Commission on Radiological Protection (ICRP) on application of principles of radiation protection to disposal of solid radioactive waste; (4) legal and regulatory requirements for disposal of low-level waste in the United States, with particular emphasis on performance objectives for near-surface disposal systems that address long-term protection of the public and the environment; and (5) other concepts that are important in demonstrating compliance with applicable performance objectives, including the role of active and passive institutional controls, model validation and confidence in model outcomes, and “reasonable assurance” of compliance. 3.1 Definition of Low-Level Radioactive Waste The historical development of the definition of low-level radioactive waste in the United States and the current definition and its implications are discussed in detail elsewhere (Kocher, 1990; NCRP, 2002). Only a summary discussion is given here. 3.1.1

Earliest Descriptions of Low-Level Waste

The term “low-level waste” was first used formally in the radioactive waste management program of the U.S. Atomic Energy Commission (AEC) in the late 1950s (Lennemann, 1972). This term referred originally to liquid wastes produced in chemical reprocessing of spent nuclear fuel that generally contained the lowest concentrations of radionuclides of any liquid wastes produced in fuel reprocessing. Liquid low-level wastes usually were released to holding ponds or lagoons or directly to surface water. Beginning about 1960, the concept that low-level waste generally contained 28

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the lowest concentrations of radionuclides also was applied to solid wastes when AEC initiated interim shallow-land burial services for solid radioactive wastes generated in the commercial sector (Lennemann, 1967). Although low-level radioactive wastes, either liquid or solid, usually were released to or disposed of in the near-surface environment, the earliest descriptions of these wastes were not based on considerations of protection of the public. Rather, they were based primarily on operational requirements for safe handling and storage of wastes at generating sites (Lennemann, 1967; 1972). Liquid and solid low-level wastes were wastes that contained concentrations of radionuclides sufficiently low that only minimal requirements on shielding and other protection systems for workers were required during waste operations. Numerical limits on concentrations of radionuclides in liquid or solid low-level wastes were developed independently at each AEC site that generated waste (Beard and Godfrey, 1967; Marter, 1967), but there was no attempt to develop limits that would be generally applicable to all sites or limits that would apply to both liquid and solid low-level wastes. Early limits were expressed in terms of total activity concentrations of all radionuclides combined. Thus, classification of radioactive wastes as “low level” was based essentially on concentrations of shorterlived radionuclides that did not pose a long-term hazard to the public following disposal or release to the environment. The earliest descriptions of low-level radioactive waste used at AEC sites were not retained when this term was defined in law. In particular: liquid and solid low-level wastes no longer were defined separately; low-level waste no longer was defined on the basis of requirements for protection of workers at waste generating sites, nor was the definition based on considerations of protection of the public at waste disposal sites; and, low-level waste no longer was defined as waste that contains relatively low concentrations of radionuclides only. 3.1.2

Current Definition of Low-Level Waste

The current legal definitions of low-level radioactive waste in the United States are contained in the Nuclear Waste Policy Act (NWPA, 1982), as amended, and the Low-Level Radioactive Waste Policy Amendments Act (LLRWPAA, 1985). In the Nuclear Waste Policy Act, low-level waste is defined as: • radioactive waste that is not high-level waste, spent nuclear fuel, transuranic waste, or uranium or thorium mill tailings; or

30 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT • radioactive waste that NRC, consistent with existing law, classifies as low-level waste. LLRWPAA contains a similar definition, except transuranic waste is not excluded from low-level waste. Thus, low-level waste is defined only by exclusion. The two definitions in law exclude high-level waste, spent nuclear fuel, and uranium or thorium mill tailings, but they differ in regard to whether low-level waste excludes transuranic waste. This inconsistency has little practical significance, because LLRWPAA is concerned only with disposal of low-level waste at sites licensed by NRC or an Agreement State, and there currently is very little transuranic waste that could be sent to such sites (DOE, 1997a). Most transuranic waste has been generated in defense-related activities at AEC and U.S. Department of Energy (DOE) sites, and this waste is excluded from low-level waste on the basis of the definition in the Nuclear Waste Policy Act. The definition of low-level waste depends on the definitions of high-level waste, spent nuclear fuel, transuranic waste, and uranium or thorium mill tailings. Definitions of the different waste classes that arise from operations of the nuclear fuel cycle are summarized in Table 3.1. The definition of high-level waste is the key to the classification system for wastes that arise from operations of the nuclear fuel cycle, because the definitions of low-level waste and transuranic waste exclude high-level waste. The definition of high-level waste is based on its source, not its radiological properties. High-level waste contains high concentrations of fission products, resulting in high levels of decay heat and external radiation, and high concentrations of long-lived, alpha-emitting transuranium radionuclides. High-level waste thus requires extensive safety systems to protect workers during waste handling and storage and highly confining and isolating disposal systems to protect the public (e.g., a geologic repository). Since high-level waste is defined on the basis of its source, other wastes with similar radiological properties that are not produced directly in fuel reprocessing are not classified as high-level waste. Such wastes, as well as incidental wastes that arise in fuel reprocessing, operations at reprocessing plants, or further processing of reprocessing wastes that have been excluded from high-level waste on a case-by-case basis, are classified as low-level waste or transuranic waste, depending on the concentrations of long-lived, alphaemitting radionuclides.

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TABLE 3.1—Summary of current definitions of different classes of radioactive wastes that arise from operations of nuclear fuel cycle in the United States.a Waste Class

Definition

High-level waste

Primary wastes, either liquid or solid, that arise from chemical reprocessing of spent nuclear fuelb

Spent nuclear fuel

Irradiated nuclear fuel that has not been chemically reprocessedc

Transuranic waste

Waste that contains more than 4 kBq g–1 of alpha-emitting transuranium radionuclides with half-lives >20 y, excluding high-level waste

Uranium or thorium mill tailings

Residues from chemical processing of ores for their source material (i.e., uranium or thorium) content

Low-level waste

Waste that is not high-level waste, spent nuclear fuel, transuranic waste, or uranium or thorium mill tailingsd

a Adapted from Kocher (1990) and NCRP (2002); definitions are simplified representations of current definitions in law. bCertain incidental wastes that arise in fuel reprocessing, operations at fuel reprocessing plants, or further processing of reprocessing wastes have been excluded from high-level waste on a case-by-case basis. Excluded incidental wastes generally have lower concentrations of fission products and long-lived, alpha-emitting radionuclides than primary reprocessing wastes that are classified as high-level waste. cSpent nuclear fuel is not waste until so declared. In some laws and regulations, spent nuclear fuel is not distinguished from high-level waste. d Low-level waste also does not include NARM not associated with operations of the nuclear fuel cycle.

The definition of transuranic waste is unique in two respects. First, the definition is quantitative, in that minimum concentrations of particular radionuclides are specified. Second, the definition is based on considerations of protection of the public following waste disposal, because waste with concentrations of long-lived, alpha-emitting transuranium radionuclides <4 kBq g–1 is generally acceptable for near-surface disposal (NRC, 1982b) but waste with higher concentrations of these radionuclides is not. Most transuranic waste contains substantially lower concentrations of fission products and long-lived, alpha-emitting transuranium radionuclides than high-level waste or spent nuclear fuel, but a disposal system similar to that for high-level waste and spent nuclear fuel will be used (i.e., a geologic repository).

32 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT Until recently, DOE sites had the option of managing and disposing of waste other than high-level waste that contained high concentrations of alpha-emitting transuranium radionuclides with half-lives <20 y (e.g., 244Cm or 252Cf) or high concentrations of alpha-emitting nontransuranium radionuclides (e.g., 233U) in the same manner as transuranic waste (DOE, 1988a). However, such wastes are not included in the current legal definition of transuranic waste. They are now classified as low-level waste, and they cannot be sent to the same disposal facility as DOE’s transuranic waste that was generated in defense-related activities. Mill tailings contain naturally occurring radionuclides, principally radium, uranium and thorium. Similar to most low-level waste, mill tailings are intended for near-surface disposal. The exclusion of mill tailings from low-level waste is based primarily on the consideration that the much larger volumes of mill tailings necessitate a separate approach to management and disposal, which emphasizes disposal in place or at sites where tailings were generated. The current definitions of waste classes summarized in Table 3.1 apply only to wastes that arise from operations of the nuclear fuel cycle [i.e., to wastes regulated under the Atomic Energy Act (AEA, 1954), as amended]. Therefore, low-level waste as defined in law also excludes naturally occurring radioactive material other than source, special nuclear, or byproduct material (e.g., mining and chemical processing wastes not included in mill tailings, wastes from treatment of public drinking water supplies, spent radium sources) or any radioactive material produced in an accelerator. The various kinds of nonfuel-cycle waste materials are termed naturally occurring and accelerator-produced radioactive material (NARM). Smaller volumes of more concentrated NARM wastes resemble some low-level wastes in their radiological properties and are managed as low-level waste, especially at DOE sites that are responsible for managing all radioactive wastes produced at those sites under authority of AEA. Larger volumes of more diffuse NARM wastes, especially diffuse radium wastes, resemble mill tailings in their radiological properties. The definition of low-level waste in the Nuclear Waste Policy Act also gives NRC the authority to define radioactive materials as low-level waste, consistent with existing law. That is, NRC has the authority to develop a definition of what low-level waste is, rather than what it is not. NRC has not exercised this authority. Indeed, since NRC can only regulate radioactive materials defined in AEA and the current definition of low-level waste excludes all other classes of radioactive waste regulated under AEA, it is not evident

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how NRC could develop a new definition of low-level waste unless NRC also developed a new definition of high-level waste that is not based on the source of the waste, as authorized by the Nuclear Waste Policy Act (NWPA, 1982). The current definition of low-level waste in law differs significantly from the earliest descriptions of liquid and solid low-level wastes discussed in Section 3.1.1. In particular, it is no longer the case that low-level waste necessarily contains relatively low concentrations of radionuclides compared, for example, with high-level waste. Rather, low-level waste can contain high concentrations of shorter-lived radionuclides, including fission and activation products with half-lives less than ~30 y, and high concentrations of long-lived radionuclides, including longer-lived fission and activation products or isotopes of uranium. Thus, low-level waste can resemble high-level waste in its radiological properties. The definition of low-level waste only by exclusion does not describe its constituents or radiological properties, nor is the definition related in any way to requirements for safe handling and storage during waste operations or safe disposal of waste.8 The definition of low-level waste only by exclusion and the resulting lack of a relationship between the definition and properties of low-level waste or requirements for management and disposal differs from definitions used in most other countries. For example, in the waste classification system currently recommended by the International Atomic Energy Agency (IAEA, 1994a), the definition of low-level waste includes a limit on thermal power density, which serves to limit concentrations of shorter-lived radionuclides, and limits on concentrations of long-lived, alphaemitting radionuclides. Thus, low-level waste, as defined by IAEA, always contains lower concentrations of radionuclides than high-level waste, and the definition of low-level waste is linked to an expectation that most such waste should be acceptable for disposal in near-surface facilities. Further discussion of IAEA’s waste classification system is given in NCRP Report No. 139 (NCRP, 2002).

8

It should be noted that the radioactive waste classification system developed by NRC in 10 CFR Part 61 (NRC, 1982b) and discussed in Section 3.4.1.1 does not provide a definition of low-level waste. Rather, it essentially defines subclasses of low-level waste, including a subclass of such waste that is not generally acceptable for near-surface disposal.

34 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT 3.2 Sources and Properties of Low-Level Waste Low-level waste is produced in many different types of activities (DOE, 1997a). In the commercial sector, the principal sources of low-level waste include: (1) operations of nuclear fuel-cycle facilities, primarily nuclear power reactors but also including chemical conversion, uranium enrichment, and fuel fabrication facilities; (2) decontamination and decommissioning of nuclear fuel-cycle facilities; and (3) activities at hospitals, medical schools, universities, radiochemical manufacturers, research laboratories, and other institutions licensed by NRC or Agreement States to possess and use radioactive materials subject to regulation under AEA (1954). Most low-level waste by volume has been generated at DOE sites. The principal sources of DOE low-level waste include plutonium production and uranium enrichment operations, the naval nuclear propulsion program, and a wide variety of research and development activities. Low-level waste also has been generated by remediation activities at contaminated properties managed by DOE, and large volumes of low-level waste may be generated by remediation of old DOE waste disposal sites under authority of the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA, 1980), as amended (also known as “Superfund”). As a result of its definition only by exclusion (Section 3.1.2) and its many different sources, low-level waste varies widely in its radionuclide compositions and concentrations and in its physical and chemical forms. In this regard, low-level waste differs from many other wastes that arise from operations of the nuclear fuel cycle, particularly high-level waste, spent nuclear fuel, and mill tailings. By volume, most low-level waste is material that has been incidentally contaminated with low levels of radioactivity. However, small volumes of low-level waste, such as ion-exchange resins and activated metals, can contain very high levels of radioactivity, including high concentrations of radionuclides with half-lives of ~30 y or greater. The widely varying physical and chemical properties of low-level waste introduce an important complicating factor into performance assessments of waste disposal facilities. This is particularly the case in regard to developing models to describe mobilization of radionuclides in solid wastes (e.g., by contact with infiltrating water) and their release into the environment. This complication provides an incentive for increased use of monolithic waste forms (e.g., incorporation of waste into grout), because releases of radionuclides from such waste forms may be predicted more reliably than releases from highly heterogeneous untreated wastes.

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3.3 ICRP Recommendations on Disposal of Radioactive Waste This Section considers recommendations of ICRP on application of principles of radiation protection to disposal of radioactive waste. In Publication 77, ICRP (1997b) issued general statements of protection policy for disposal or release to the environment of radioactive waste of any kind, including solid, liquid and gaseous wastes. That report was followed by Publication 81 (ICRP, 1998), which addresses application of protection principles to disposal of solid radioactive wastes. This Section reviews recommendations of ICRP, with particular emphasis on recommended protection criteria and their application, as well as approaches to demonstrating compliance with protection criteria by means of performance assessment. Earlier ICRP recommendations on disposal of solid radioactive waste are given in Publication 46 (ICRP, 1985). IAEA also issued several documents that discuss principles of radioactive waste disposal and their application (IAEA, 1989a; 1994b; 1995a; 1996a; 1997b). Current ICRP recommendations incorporate many of the principles discussed in IAEA documents, and some of these documents are discussed elsewhere in this Report. 3.3.1

General Recommendations on Radiation Protection

The system of radiation protection9 recommended by ICRP for proposed and continuing practices (including waste disposal) is based on three general principles (ICRP, 1991): 1. No practice involving exposure to radiation should be adopted unless it produces sufficient benefit to exposed individuals or society to offset the radiation detriment it causes (justification of a practice). 2. For any source within a justifiable practice, the magnitude of individual doses, the number of people exposed, and the likelihood of incurring exposures when they are not certain to be received should all be kept as low as reasonably achievable (ALARA), economic and social factors being taken into account. This procedure should be constrained by restrictions on doses to individuals (dose constraints) or on risks to individuals in the case of potential exposure (risk 9ICRP generally uses the term “radiological protection” (ICRP, 1991). The term “radiation protection” is used in this Report to be consistent with terminology normally used by NCRP (1993).

36 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT constraints) so as to limit the inequity likely to result from the inherent economic and social judgments (optimization of protection). 3. Exposures of individuals resulting from all justifiable practices combined should be subject to dose limits, or to some control of risk in the case of potential exposures. These limits are aimed at ensuring that no individual is exposed to radiation risks from these practices that are judged to be unacceptable in any normal circumstances (individual dose and risk limits). Similar recommendations on basic principles of radiation protection have been developed by NCRP in Report No. 116 (NCRP, 1993). 3.3.2

General Policy on Application of Protection Principles to Radioactive Waste Disposal

In Publication 77, ICRP (1997b) provides general statements of policy on application of the principles of radiation protection described above to radioactive waste disposal. These policies are summarized as follows. 1. Justification of practices should not be applied to waste management and disposal operations in isolation. Rather, those practices should be included in an assessment of the justification of each practice that generates waste. 2. Optimization of protection is broadly interpreted as doing all that is reasonable to reduce doses. Although optimization of collective dose on the basis of differential cost-benefit analysis has conventionally been one input to optimization of protection, problems of estimating collective dose over long time periods in the future are recognized, and the main emphasis is given to qualitative specification of optimization. A dose constraint is an important component of optimization of protection, and it should be used prospectively to exclude protection options that would cause the dose to a member of the critical group to exceed the constraint. An effective dose to members of the public of no more than ~0.3 mSv y–1 from radioactive waste disposal is the recommended dose constraint. 3. Application of a dose limit for all justifiable practices combined, which currently is an effective dose of 1 mSv y–1 (ICRP, 1991), to waste disposal has inherent difficulties.

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However, a process of constrained optimization (i.e., optimization of protection at each disposal site subject to restriction by a dose constraint) will obviate the need to use a dose limit in control of radioactive waste disposal. 4. Accidents and disruptive events, if they occur, may cause exposures that are larger than normal. Such exposures should be treated as potential, and their magnitude and probability should be taken into account in reaching waste management decisions. However, the role of potential exposure in risk assessment for disposal of long-lived radionuclides is not yet clear. In the second statement given above, the critical group should be representative of individuals in an exposed population who are reasonably expected to receive the highest doses from a source, or group of sources, under consideration. The dose to the average individual in the critical group then could be taken as representative of the highest dose delivered by the source or sources. ICRP’s recommendations on potential exposures mentioned in the fourth statement are contained in Publication 64 (ICRP, 1993a) and elaborated on in Publication 76 (ICRP, 1997a). In Publication 46, ICRP (1985) had recommended that a suitable risk constraint for potential exposure was the value 10–5 y–1, which is the same order of magnitude as the fatal cancer risk implied by a dose limit for the public of 1 mSv y–1 when a risk coefficient for fatal cancers of ~10–2 Sv–1 is assumed (ICRP, 1977). The recommended risk constraint corresponds to a lifetime fatal cancer risk of ~10–3. 3.3.3

Application of Protection Principles to Disposal of Solid Radioactive Wastes

In Publication 81, ICRP (1998) has applied the general principles of radiation protection discussed in Section 3.3.1 and the general policies on application of protection principles to radioactive waste disposal discussed in Section 3.3.2 to disposal of long-lived solid radioactive waste. Selected recommendations of interest to this Report are summarized as follows: Protection of Future Generations • The principal means of achieving protection of the public is through the process of constrained optimization, taking into account the recommended upper bound for the dose constraint of 0.3 mSv y–1 or its risk equivalent of ~10–5 y–1, assuming a detriment coefficient of 0.05 Sv–1 (ICRP, 1991),

38 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT in cases of potential exposure (i.e., when the probability of exposure according to an assumed scenario may be less than one). • Individuals and populations in the future should be afforded at least the same level of protection from actions taken today as is the current generation. This objective implies the use of current dose and risk constraints as indicators of the long-term performance of waste disposal systems, which are compared with estimates of dose or risk over long time periods to give an indication of whether a disposal facility is acceptable given current understanding of its performance. • It cannot be assumed that future generations will have knowledge of disposals undertaken at the present time. Therefore, protection of future generations should be achieved primarily by passive measures at the development stage of a disposal facility, and should not rely unduly on active measures taken in the future. However, institutional controls maintained after facility closure may enhance confidence in the safety of a facility, particularly by reducing the likelihood of human intrusion. There is no reason why these controls may not continue for extended periods of time and, at near-surface disposal facilities in particular, they may make a significant contribution to overall safety. Critical Group • Exposures should be assessed on the basis of the mean annual effective dose in the critical group, which should be small enough to be relatively homogeneous with respect to age, diet, and those aspects of behavior that affect annual doses received. Due to the long time scales of concern, the critical group must be hypothetical. Habits and characteristics of the critical group should be chosen on the basis of reasonably conservative and plausible assumptions, taking into account current lifestyles and available site- or regionspecific information. • In many cases, different scenarios for future events and processes that affect the performance of a disposal system and resulting exposures of the public, each associated with different critical groups, may have different probabilities of occurrence and, therefore, the highest dose may not be linked to the highest risk. • Critical groups cannot be defined independently of an assumed biosphere. Major changes in the biosphere may occur over long time periods, but consideration of such

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changes should be limited to those due to natural forces. A critical group and biosphere should be defined using either site- or region-specific information or a stylized approach based on more general habits and conditions. Stylized approaches will become more important for longer time scales. • Given that radioactive contamination in the biosphere resulting from waste disposal is likely to remain relatively constant over periods considerably longer than the human life span, it is reasonable to calculate an annual dose or risk averaged over the lifetime of individuals, and it is not necessary to calculate doses or risks to different age groups. An average annual dose or risk can be adequately represented by the annual dose or risk to an adult. Potential Exposures • Exposures might result from processes that affect the expected behavior of a disposal system and often have an assumed probability of occurrence in a given time of less than one. Therefore, the objective of protecting individuals from potential exposures associated with credible processes is best achieved by considering both the probability of occurrence and the magnitude of exposure in a manner similar to recommendations for potential exposure situations at the present time (ICRP, 1993a). A previous recommendation (ICRP, 1985) that the normal evolution of a disposal system and probabilistic situations (accidents and disruptive events) should be treated separately may not be needed. Optimization of Protection • Constrained optimization is the main approach to evaluating the radiological acceptability of a waste disposal option. Although there are a number of methodological options for implementing constrained optimization in the context of potential exposures, formal optimization techniques for application to potential exposures remain to be developed. • Optimization of protection for waste disposal is a judgmental process that takes social and economic factors into account and should be conducted in a structured, essentially qualitative way. The goal is to ensure that reasonable measures have been taken to reduce future doses to the extent that required resources are in line with these reductions. Optimization of protection should be applied in an iterative manner during the development stage of a disposal system and should cover site selection and facility design phases.

40 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT • Two broad categories of exposure scenarios have to be considered: those that involve natural processes and those that involve human intrusion. Optimization should explore and apply reasonable measures to reduce the probability and/or magnitude of exposures due to natural processes by considering, for example, seismic properties of a site, the retention capability of a facility, and waste package design, and due to human intrusion by considering, for example, the presence of natural resources, selection of a depth of the disposal facility below the ground surface, and institutional control measures. • Assessed doses and risks to individuals are inputs to the optimization process. However, radiation protection criteria are applied differently to natural processes and human intrusion. For natural processes, estimated doses or risks are compared with a dose or risk constraint to establish the acceptability of potential exposures of individuals. In contrast, when considering implications of human intrusion, it is not appropriate to apply a dose or risk constraint for radioactive waste disposal, because there is little or no scientific basis for predicting the nature or probability of future human activities and because, by definition, an intrusion event bypasses some or all of the barriers that have been incorporated in the disposal facility as part of optimization of protection. • Although collective dose is of limited use in optimization of protection for waste disposal facilities, consideration of the number of people potentially exposed and the distribution of individual doses in time can be of some help. Radiological Criteria Applied to Scenarios Representing Natural Processes • Approaches to show whether a dose constraint (e.g., ~0.3 mSv y–1) or a risk constraint (e.g., ~10–5 y–1) for natural processes is satisfied can involve either: (1) aggregation of risks by combining doses and probabilities or (2) presentation of the dose and probability of occurrence for each exposure situation separately. The same degree of protection can be achieved by using either, in the first approach, a risk constraint or, in the second approach, a dose constraint supplemented by consideration of the probability that doses would be incurred. • In an aggregated approach, the total risk from all credible natural processes that may result in doses to individuals in

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the future is compared with a risk constraint. Although this approach is conceptually satisfying, it requires a comprehensive evaluation of all relevant exposure situations and their associated probabilities within the time period under consideration. • In a disaggregated dose/probability approach, likely, or representative, release scenarios are identified and calculated doses from those scenarios are compared with a dose constraint. The radiological significance of other, less likely, scenarios can be evaluated by separate consideration of resultant doses and their probability of occurrence. This approach does not require precise quantification of the probability of occurrence of less likely scenarios but, rather, an appreciation of radiological consequences balanced against an estimated probability. Such other considerations as the duration and extent of calculated doses or risks also may be taken into account in evaluating the significance of less likely scenarios.

Radiological Criteria Applied to Human Intrusion • There are two concerns about the consequences of human intrusion into a waste disposal facility: (1) direct exposure of nearby populations to waste that is brought to the surface by an intrusion event, and (2) an increase in releases to the environment and exposures of the public associated with natural processes (e.g., increased releases to groundwater as a result of drilling through a facility). • Protection from exposures associated with human intrusion is best accomplished by efforts to reduce the possibility of such events (e.g., by disposal well below the ground surface, incorporation of robust design features that make intrusion more difficult, active and passive institutional controls). Since a future society may be unaware of exposures resulting from intrusion, any protective actions required should be considered during development of a disposal facility. • Because the occurrence of human intrusion cannot be totally ruled out, the consequences of one or more typically plausible stylized intrusion scenarios should be considered by the decision maker to evaluate the resiliency of a disposal facility to potential intrusion. However, since there is no scientific basis for predicting the nature or probability of future human actions, it is not appropriate to include probabilities of such events in a quantitative performance

42 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT assessment that is to be compared with a dose or risk constraint. Nevertheless, a measure of the significance of human intrusion for radiological protection is necessary. • In circumstances where human intrusion could lead to doses to those living around a disposal site sufficiently high that intervention based on current criteria would almost always be justified, reasonable efforts should be made to reduce the probability of human intrusion or to limit its consequences. ICRP previously advised that an annual effective dose of ~10 mSv may be used as a generic reference level below which intervention is not likely to be justified, and that an annual effective dose of ~100 mSv may be used as a generic reference level above which intervention should be considered almost always justifiable.

Technical and Managerial Principles • Technical and managerial principles to enhance confidence that radiation safety will be maintained throughout the postclosure period should be applied during the disposal system development process and in a manner consistent with the intrinsic hazard of the waste and the level of residual uncertainty identified in the performance assessment. • The key principle is defense-in-depth that provides successive passive safety measures and is applied mainly by using multiple barriers that provide a combination of different lines of defense against potential challenges to the safety of a disposal system. • Other technical and managerial principles to enhance confidence in the safety of a disposal system include: (1) use of sound engineering principles and practices that are proven by testing and experience to the extent feasible, considering the need for research and innovation to improve safety, (2) a comprehensive system of quality assurance, (3) iterative conduct of performance assessments until closure to identify potential vulnerabilities and sensitivities in the disposal system considering the inherent limitations in an assessment methodology, potential gaps in data, and alternative interpretations of existing data to ensure that numerical results adequately represent or bound system performance, and (4) establishment of a feedback mechanism so that results of assessments can be taken into account to guide the development process.

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Consistency with Radiological Criteria • Assessments of compliance with radiological protection criteria should be based on a comprehensive safety case supported by multiple lines of reasoning including a good understanding of system behavior and quantitative performances, supported with and complemented by qualitative arguments including information from observation of the natural system at the disposal site and natural analogs. • The sophistication of a performance assessment should be consistent with the hazard posed by the waste. When a robust assessment shows that results are well within recommended constraints, there may be no need for further analysis. • Calculated doses or risks should be considered as performance measures or safety indicators. To provide additional insight, it may be useful to make qualitative comparisons particularly for the distant future (e.g., comparisons of the remaining hazard potential with risks imposed by other natural or human-induced sources). • To evaluate the performance of waste disposal systems over long time scales, one approach is to consider quantitative estimates of dose or risk for time periods on the order of 1,000 to 10,000 y. This approach recognizes the possibility that risks associated with major geologic changes (e.g., glaciation, tectonic movements) may obscure risks associated with a waste disposal system over longer time frames. Another approach is to consider quantitative calculations farther into the future by making increasing use of stylized assumptions and taking into account the time periods when judging the calculated results. Qualitative arguments would provide additional information to this judgmental process. • Demonstration of compliance with radiation protection criteria is best achieved using a stepwise or iterative approach at various stages of facility development and critical review as the development process makes progress. • Uncertainties, some of which are unquantifiable, may be characterized as: (1) data uncertainty that reflects incomplete knowledge of the performance of a disposal system as influenced by the design and immediate environment after closure, (2) future states uncertainty that reflects an imperfect ability to predict future human actions and states of the environment, and (3) model uncertainty that reflects uncertainty about a conceptual description of a disposal system, a mathematical description of the concept, and implementation of a mathematical description in a computer code.

44 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT • Efforts should be undertaken during development of a disposal system to understand the significance of uncertainties and to reduce or bound uncertainties through site characterization and experimentation. Expert judgment should be used to evaluate the significance of residual uncertainties. • Uncertainty analysis should be an integral part of the performance assessment process to calculate dose or risk and, whenever possible, reported results should include ranges on possible values rather than a single point value. An analysis of uncertainty should be adequate for the purpose of an assessment. • Demonstration of compliance with radiation protection criteria in the future is not simply a straightforward comparison of estimated doses or risks with constraints. Proof of compliance cannot be absolute because of the inherent uncertainties, especially in understanding the evolution of the geologic setting, biosphere, and engineered barriers over long time periods. A decision on the acceptability of a disposal system should be based on reasonable assurance, rather than an absolute demonstration of compliance. • Evaluation of compliance requires judgment. Dose or risk constraints should increasingly be considered as reference values for time periods farther into the future. Numerical compliance alone should not compel acceptance of a proposed disposal system. Adequate evidence of the quality of supporting data and analyses, as well as an assessment that the design and construction of a facility comports to technical and managerial principles should also be required. Transgressions of constraints do not necessarily oblige rejection of a safety case, due to unquantified conservatism that is likely to be incorporated in a performance assessment; as the time frame increases, some allowance should be made for an assessed dose or risk exceeding the constraint. Any transgression must be justified and system safety must be supported by other evidence, or reasons for a transgression must be evaluated to determine if additional measures would result in better protection. • Judgment is required in optimization of protection and application of technical and managerial principles, but this should not be viewed as an open-ended process. Provided that the appropriate constraint for natural processes has been satisfied, that reasonable measures have been taken to reduce the probability of human intrusion or limit its consequences, and that sound engineering, technical and managerial principles have been followed, the radiation protection principles can be considered satisfied.

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Discussion of ICRP Recommendations

Many discussions in this Report on the nature and general principles of performance assessment and use of performance assessment in determining acceptable disposals of low-level waste in near-surface facilities are consistent with ICRP recommendations on application of radiation protection principles to disposal of solid radioactive wastes. Examples include: (1) the importance of judgment in the conduct of performance assessment and in applying results of performance assessments in regulatory decision making, (2) the need for iteration of performance assessments and establishment of feedback mechanisms so that results can be used to guide development of a disposal facility, (3) the need to consider uncertainty analysis as an integral part of performance assessment, (4) the importance of tailoring the sophistication and robustness of a performance assessment and an analysis of uncertainty to the hazards posed by the waste and the purpose of the assessment, and (5) the need to consider effects of natural processes on the performance of disposal systems separately from effects of human intrusion. It also is the case that important aspects of regulatory requirements for near-surface disposal of low-level waste in the United States are consistent with ICRP recommendations. For example, as indicated in the next section, regulations emphasize use of dose constraints for individual members of the public of ~0.3 mSv y–1 to judge the acceptability of near-surface disposal systems with respect to projected releases of radionuclides beyond the site boundary as a result of natural processes, and those constraints are applied for a period after facility closure of 1,000 or 10,000 y. Regulations also make a clear distinction between protection of inadvertent intruders and protection of members of the public beyond the site boundary, and they incorporate dose criteria to limit effects of human intrusion that are substantially higher than the dose constraint for individual members of the public beyond the site boundary. In some respects, however, regulations for near-surface disposal of low-level waste in the United States and the conduct of performance assessments are not consistent with ICRP recommendations. Regulations do not include a risk constraint that would apply to potential exposure situations that result from natural processes for which the probability of occurrence may be less than one. Thus, in demonstrating compliance with a dose criterion for off-site members of the public, scenarios for release of radionuclides and exposure of the public that are included in an assessment (e.g., mobilization by infiltrating water and transport in groundwater)

46 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT are assumed to occur with a probability of one. Other scenarios that are presumed to be less likely or to have lower consequences usually are assessed with less rigor, and probabilities of occurrence usually are considered only qualitatively. Another difference between regulations for near-surface disposal of low-level waste in the United States and ICRP recommendations concerns the treatment of human intrusion. ICRP has recommended that the consequences of one or more stylized intrusion scenarios as they affect releases from a disposal facility and exposure of the public should be evaluated and considered by decision makers. Although a requirement to evaluate potential impacts on the public on the basis of an assumption that drilling through a single waste package would occur is included in regulations for disposal of spent nuclear fuel and high-level waste in the proposed geologic repository at the Yucca Mountain Site in Nevada (EPA, 2001a; NRC, 2001), a requirement to consider impacts of intrusion on off-site releases is not included in NRC and DOE regulations for near-surface disposal of low-level waste. Justifications for not including such a requirement are discussed in Section 6.7. Stylized intrusion scenarios at near-surface facilities discussed in Section 6 are applied in evaluating potential impacts on intruders or other individuals who might reside on a disposal site after an assumed loss of institutional control, and those evaluations are used to establish limits on concentrations of radionuclides that would be acceptable for near-surface disposal.10 Dose criteria used to establish concentration limits at near-surface facilities in the United States are substantially lower than the range of 10 to 100 mSv y–1 that ICRP recommends for use in determining a need to reduce the probability of human intrusion or to limit its consequences. 3.4 Requirements for Near-Surface Disposal of Low-Level Waste This Section discusses the types of disposal facilities that have been authorized in the United States for disposal of low-level waste and legal and regulatory requirements for disposal of low-level waste in near-surface facilities, including performance objectives 10

It is generally recognized that there are no reasonable means of protecting inadvertent intruders in the unlikely event that intrusion into highly hazardous spent nuclear fuel or high-level waste in a deep geologic repository should occur (e.g., by drilling). Rather, the best that can be done is to take reasonable measures to reduce the likelihood of intrusion (NAS/NRC, 1995).

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and other criteria that address long-term protection of the public and the environment. Other approaches to regulating lowlevel waste disposal that have been developed elsewhere also are discussed. 3.4.1

Authorized Disposal Systems

This Section discusses the types of disposal systems for low-level waste that are specified in laws and regulations and the historical development of near-surface disposal technologies. 3.4.1.1 Legal and Regulatory Specifications. Disposal of low-level waste is subject to general provisions of AEA (1954), as amended, in regard to the need to protect public health and safety. Disposal of non-DOE low-level waste also is subject to provisions of the Low-Level Radioactive Waste Policy Act (LLRWPA, 1980), as amended (LLRWPAA, 1985), which established a system of State Compacts to promote siting of new facilities. None of these laws specify a particular disposal technology for low-level waste, in contrast to high-level waste, spent nuclear fuel, and transuranic waste for which particular disposal technologies (and facilities) are specified in law. The lack of specificity regarding acceptable disposal technologies reflects the existence of historical disposal practices and the wide variety of low-level wastes that can be accommodated most efficiently by a variety of technologies. Current law presumes that disposal near the land surface will be used for most low-level waste. However, the Low-Level Radioactive Waste Policy Amendments Act (LLRWPA, 1985) directed NRC to identify alternatives to shallow-land burial for non-DOE low-level waste and to establish technical guidance and requirements for licensing of alternative disposal facilities. In addition, systems involving deeper disposal have been used for the highest-activity low-level wastes generated at DOE sites, including so-called greater confinement disposal facilities at the Nevada Test Site (DOE, 1997b). Disposal of non-DOE low-level waste in near-surface facilities, either above or below grade, is subject to licensing criteria established by NRC in 10 CFR Part 61 (NRC, 1982b; 1993) (Section 3.4.2.1) or comparable licensing criteria established by Agreement States. An important feature of NRC’s licensing criteria in 10 CFR Part 61 is a classification system for near-surface disposal, which defines waste that is generally acceptable for near-surface disposal (Class-A, -B or -C waste) and waste that is not (greater-

48 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT than-Class-C waste). An amendment to the licensing criteria (NRC, 1989) specifies that disposal of greater-than-Class-C waste in a geologic repository that will be used for disposal of spent nuclear fuel and high-level waste is required unless disposal elsewhere, including the possibility of a near-surface facility, is approved by NRC on a case-by-case basis. NRC’s licensing criteria in 10 CFR Part 61 serve to emphasize that particular disposal systems for low-level waste are not specified in law or regulations. Rather, given the wide variety of low-level wastes that require disposal, different types of disposal technologies have been used to provide protection of the public. The particular choice often depends on the radiological properties of waste and properties of a specific disposal site. It also should be emphasized that NRC’s waste classification system in 10 CFR Part 61 does not define low-level waste, because greater-than-Class-C waste is still a form of low-level waste. Rather, 10 CFR Part 61 essentially defines subclasses of low-level waste (NCRP, 2002).

3.4.1.2 Historical Development of Disposal Technologies. The earliest disposals of most solid low-level waste, except waste that was dumped in the oceans, took place in near-surface facilities that were constructed without engineered barriers. Lower-activity solid waste usually was packaged in plastic bags, cardboard or wooden boxes, or metal drums and placed in shallow trenches, and higher-activity solid waste usually was placed in shielded packages to protect workers and then placed in deeper trenches or auger holes. Liquid low-level waste often was released directly to the land surface or surface water, or injected underground. Two significant differences between current disposal systems for low-level waste and past practices are evident. First, disposal of liquid low-level waste on land is no longer allowed. Second, many disposal systems now place considerable reliance on engineered barriers to provide protection of the public, including inadvertent intruders. For example, many disposal systems use waste compaction, chemical treatment of waste, waste encapsulation (e.g., in grout, bitumen or polyethylene), physical barriers to water infiltration and human intrusion (e.g., concrete vaults or waste packages), leachate collection systems, and disposal well below the ground surface. In the future, disposal of untreated waste in shallow trenches without engineered barriers probably will be used only for low-activity wastes or at arid sites with unusually favorable hydrologic conditions (e.g., the Envirocare Facility in Utah).

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Requirements for Protection of the Public

Disposal of low-level waste in the near-surface environment has been practiced since the earliest times that such waste was generated. Until about 1960, there were no regulatory requirements on disposal of low-level waste to define acceptable exposures of the public. Some protection of the public was provided indirectly by limits on concentrations of radionuclides in low-level waste that were developed at each generating site (Beard and Godfrey, 1967; Marter, 1967) and by natural dilution and dispersion following releases to the environment. Beginning about 1960, disposal of low-level waste was subject to radiation protection standards for the public in 10 CFR Part 20 (AEC, 1960) and similar standards that applied at AEC sites (AEC, 1958). Those standards, which applied to releases beyond site boundaries from all controlled sources combined, including all site operations as well as waste disposal, specified limits on annual dose equivalent to individual members of the public of 5 mSv to the whole body, 30 mSv to bone or the thyroid, or 15 mSv to any other organ. Compliance with radiation protection standards for the public was verified by environmental monitoring, and prospective performance assessments of waste disposal facilities were not conducted. Beginning in the late 1970s, efforts were undertaken to develop standards, in the form of performance objectives and other criteria, that would apply specifically to near-surface disposal of low-level waste. Such standards would define adequate protection of the public for this practice and would apply for long times in the future. Selected performance objectives and other criteria that have been established by NRC and DOE are summarized in Table 3.2 and discussed in the following two sections. While the criteria in Table 3.2 directly affect the types of analyses that must be included in performance assessments, it is important to emphasize that, first, NRC and DOE regulations include other qualitative requirements that are important to facility design and operations and, second, institutional controls are an important means of protecting inadvertent intruders. A third section considers views of the U.S. Environmental Protection Agency (EPA) on acceptable standards for disposal of low-level waste. 3.4.2.1 Licensing Criteria Established by NRC. NRC’s licensing criteria for near-surface disposal of non-DOE low-level waste are given in 10 CFR Part 61 (NRC, 1982b); these criteria also apply to disposal of DOE low-level waste if such waste is sent to a facility licensed by NRC. This regulation includes the following performance objectives:

Regulations 10 CFR Part 61 (NRC, 1982b)

Order 435.1 (DOE, 1999a; 1999b)c

Applicability

Numerical Performance Objectives and Other Criteria

Off-site members of the public

Limits on annual dose equivalent from all release and exposure pathways of 0.25 mSv to whole body, 0.75 mSv to thyroid, and 0.25 mSv to any other organ

Inadvertent intruders onto disposal sites

Protection of inadvertent intruders must be ensured at any time after active institutional controls over disposal sites are removedb

Off-site members of the public

Limit on annual effective dose equivalent from all release and exposure pathways of 0.25 mSv, excluding dose from radon and its progeny in air Limit on annual effective dose equivalent from releases to air of 0.1 mSv, excluding dose from radon and its progeny Limit on release rate of radon to air of 0.7 Bq m–2 s–1 or, alternatively, concentration of radon in air at boundary of facility of 20 Bq m–3

Inadvertent intruders onto disposal sites

Performance measures of effective dose equivalent of 1 mSv y–1 for scenarios involving continuous exposure and 5 mSv for scenarios involving single acute exposured

Water resources

Analysis of potential impacts on water resources to establish limits on quantities of radionuclides for disposal based on site-specific criteriae

a Performance objectives and other criteria are discussed in Section 3.4.2. In addition, NRC and DOE regulations both include provisions that releases should be maintained ALARA. bRequirement is implemented by means of waste classification system, which includes limits on concentrations of radionuclides that are generally acceptable for near-surface disposal in Class-A, -B and -C wastes (Table 3.3) and technical requirements for disposal of waste in each class. cOrder 435.1 applies to waste disposals after July 9, 1999, and retroactively to waste disposals between September 26, 1988, and July 9, 1999, that were conducted in accordance with requirements in DOE Order 5820.2A (DOE, 1988a), except Order 5820.2A may continue to apply in some cases. d Performance measures for inadvertent intruders are used to establish limits on concentrations of radionuclides that are acceptable for near-surface disposal on site-specific basis. eDisposals must comply with other requirements for environmental protection (DOE, 1988b; 1990) and any applicable federal, state and local laws and regulations.

50 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT

TABLE 3.2—Summary of selected performance objectives and other criteria in NRC and DOE regulations for near-surface disposal of low-level waste.a

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• For off-site members of the public, limits on annual dose equivalent from all release and exposure pathways of 0.25 mSv to the whole body, 0.75 mSv to the thyroid, and 0.25 mSv to any other organ. • Reasonable effort should be made to maintain releases of radioactivity to the general environment ALARA. • Design, operation and closure of disposal facilities must ensure protection of individuals who might inadvertently intrude into disposal sites or contact the waste at any time after active institutional controls over the sites are removed. In the performance objective for protection of inadvertent intruders, loss of active institutional controls over disposal sites is assumed to occur at 100 y after facility closure. The performance objective for off-site members of the public does not specify a time period over which the dose criteria should be applied; this issue is discussed in Section 3.4.3.1. The dose criteria in NRC’s performance objective for off-site members of the public are the same as those in EPA standards for routine releases of radionuclides from uranium fuel-cycle facilities in 40 CFR Part 190 (EPA, 1977). These criteria are a small fraction of dose limits in radiation protection standards for the public at the time the performance objective was established (AEC, 1960), as described in the previous section. This performance objective was based in part on generic performance assessments of reference disposal facilities at several types of sites, which indicated that the specified dose criteria should be reasonably achievable at any suitable disposal site (NRC, 1981a; 1982a). As a matter of policy, NRC now considers that an annual total effective dose equivalent of 0.25 mSv is an appropriate performance objective, although the dose criteria for the whole body or individual organs specified in 10 CFR Part 61 can be used. This policy is consistent with NRC’s use of total effective dose equivalent in its radiation protection standards for the public in 10 CFR Part 20 (NRC, 1991a). In addition, when a probabilistic approach to performance assessment is used to obtain a probability distribution of projected doses, NRC staff have recommended that the highest mean dose in any year should be <0.25 mSv total effective dose equivalent, and that the upper 95th percentile of projected annual doses over time should be <1 mSv (NRC, 2000). The latter criterion is the dose limit for chronic exposure in NRC’s radiation protection standards for the public (NRC, 1991a). NRC staff recommendations have not been approved by NRC and, thus, should not be viewed as official regulatory policy.

52 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT The performance objective for protection of inadvertent intruders is implemented by means of NRC’s waste classification system, which includes: (1) generally applicable limits on concentrations of radionuclides in Class-A, -B and -C waste; (2) requirements on physical forms, characteristics, and stability of waste in each class; and (3) requirements on disposal of waste in each class (NRC, 1982b). Limits on concentrations of radionuclides in each waste class were developed on the basis of several considerations, including: (1) dose criteria for an inadvertent intruder expressed as an annual dose equivalent of 5 mSv to the whole body and bone or 15 mSv to any other organ, which were consistent with radiation protection standards for the public at that time (AEC, 1960); (2) generic analyses of doses associated with selected exposure scenarios for inadvertent intruders (Section 6.3), which are assumed to occur after loss of active institutional controls over a disposal site at 100 y after disposal; and (3) information on radionuclide concentrations in non-DOE low-level wastes. Limits on concentrations of radionuclides in Class-A, -B and -C wastes established by NRC are given in Table 3.3. An important consequence of NRC’s waste classification system is that it obviates the need to include an analysis of potential impacts of low-level waste disposal on future inadvertent intruders in performance assessments for NRC-licensed near-surface facilities. Rather, protection of inadvertent intruders is provided by compliance with the waste classification system in 10 CFR Part 61, including requirements on waste forms and disposal of Class-A, -B and -C wastes as well as the generally applicable limits on concentrations of radionuclides in each waste class. 3.4.2.2 Requirements Established by DOE. Current requirements for near-surface disposal of low-level waste at DOE sites are given in DOE Order 435.1 (DOE, 1999a; 1999b). These requirements include the following performance objectives: • For off-site members of the public, a limit on annual effective dose equivalent from all release and exposure pathways of 0.25 mSv, excluding the dose from radon and its progeny in air. • For off-site members of the public, a limit on annual effective dose equivalent from releases to the atmosphere of 0.1 mSv, excluding the dose from radon and its progeny. • A limit on release rate of radon at the surface of a disposal facility of 0.7 Bq m–2 s–1 or, alternatively, a limit on concentration of radon in air at the boundary of the disposal facility of 20 Bq m–3.

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TABLE 3.3—Concentration limits of radionuclides in NRC's generally applicable waste classification system for near-surface disposal of radioactive waste.a Concentration (TBq m–3)b Radionuclide Class A

Class B

Class C

—c

—c

—c

—c

—c

—c

Shorter lived Total of all isotopes with half-lives <5 y H-3 Co-60

26 1.5 26

Ni-63

0.13

2.6

Ni-63 in activated metal

1.3

Sr-90

0.0015

5.6

260

Cs-137

0.037

1.6

170

26

26 260

Longer livedd C-14

0.03

0.3

C-14 in activated metal

0.3

3

Ni-59 in activated metal

0.81

8.1

Nb-94 in activated metal

0.00074

0.0074

Tc-99

0.011

0.11

I-129

0.0003

0.003

Alpha-emitting transuranium isotopes with half-lives >5 y

0.4e

4e

Pu-241

13e

130e

Cm-242

74e

740e

aBased

on Table 1 and Table 2 of 10 CFR Part 61 (NRC, 1982b). For mixtures of radionuclides, the sum of ratios of concentrations to the respective concentration limits must not exceed unity. Waste classification system also includes requirements on waste forms and disposal of Class-A, -B and -C wastes. b Except as noted, concentrations are obtained from values specified in 10 CFR Part 61 in units of Ci m–3 (1 Ci = 0.037 TBq). c There are no limits established for these radionuclides in Class-B or -C wastes. dThere are no Class-B limits for longer-lived isotopes. e Values are in units of kBq g–1 and are obtained from values specified in 10 CFR Part 61 in units of nCi g–1 (1 nCi = 0.037 kBq).

54 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT These performance objectives are similar to those established previously in DOE Order 5820.2A (DOE, 1988a). Specification of dose criteria in terms of effective dose equivalents is consistent with DOE’s radiation protection standards for the public (DOE, 1990). However, if DOE low-level waste is sent to a facility licensed by NRC or an Agreement State, performance objectives and other criteria established by those authorities apply. DOE Order 435.1 specifies that performance objectives for disposal of low-level waste apply for 1,000 y after disposal (DOE, 1999b). This choice was based on a number of considerations. First, DOE considers that projections of dose beyond 1,000 y are so uncertain that they are just as likely to result in a poor decision about the acceptability of waste disposal at a particular site as a good decision. Second, waste must be disposed of somewhere, and as time increases, DOE considers that uncertainty increases to such an extent that differences between disposal options (i.e., differences in impacts from disposal at different sites) become indistinguishable. Third, on the basis of a study by the National Academy of Public Administration (Finger et al., 1997), DOE considers that near-term impacts should be given greater weight in decision making than hypothetical impacts in the distant future. Calculations of dose beyond 1,000 y should be considered only qualitatively at DOE sites, and only in comparing disposal alternatives that are otherwise indistinguishable in their projected impacts. The issue of time of compliance with performance objectives is discussed further in Section 3.4.3.1. DOE Order 435.1 includes other requirements and criteria that directly affect the conduct of performance assessments but are not stated as performance objectives. First, performance assessments must include a demonstration that releases from a disposal facility will be maintained ALARA (DOE, 1999b; 1999c).11 Second, site-specific assessments of potential doses to inadvertent intruders are required for the purpose of establishing limits on concentrations of radionuclides for disposal at each site (DOE, 1999b). Performance measures for inadvertent intruders are the same as previous performance objectives for protection of inadvertent 11An ALARA requirement was included in performance objectives in the previous Order 5820.2A (DOE, 1988a). Removal of this requirement from performance objectives in Order 435.1 was based on the view that ALARA is a process that must be conducted to reduce potential doses to the public to levels that are as low as reasonably achievable, rather than a performance objective of the same kind as numerical criteria to limit releases and exposures of off-site members of the public.

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intruders in Order 5820.2A (DOE, 1988a).12 Third, an analysis of potential impacts on water resources is required for the purpose of establishing limits on quantities of radionuclides that are acceptable for disposal at each site on the basis of site-specific criteria (DOE, 1999c). This requirement replaces a previous performance objective in Order 5820.2A (DOE, 1988a), which specified that groundwater resources should be protected in accordance with applicable federal, state and local requirements. Order 435.1 also specifies that disposals must comply with other DOE requirements on environmental protection (DOE, 1988b; 1990). DOE requirements that are directed at protection of water resources, especially the previous performance objective in Order 5820.2A (DOE, 1988a), usually have been interpreted in terms of compliance with existing or proposed EPA standards for radioactivity in public drinking water supplies in 40 CFR Part 141 (e.g., McDowell-Boyer et al., 2000; ORNL, 1997a). Drinking water standards for radionuclides are discussed in the following section. A provision in Order 435.1 that was not included in the previous Order 5820.2A (DOE, 1988a) and is not indicated in Table 3.2 is a requirement to perform a composite analysis at each disposal site (DOE, 1999b; 1999c). DOE’s low-level waste disposal facilities often are located in the same area as old, inactive waste disposal sites or other sources, and releases from multiple sources may overlap. The purpose of a composite analysis is to assess sources of public exposure that may overlap with releases from a low-level waste disposal facility of concern. Results of a composite analysis are used to determine if there is reasonable confidence that combined releases from disposal facilities and other overlapping sources will be limited to an extent sufficient to provide long-term protection of the public and, if not, to include appropriate plans for mitigation of potential future releases in DOE’s environmental management program. An example of a composite analysis is given in ORNL’s Composite Analysis for Solid Waste Storage Area 6 (ORNL, 1997b). Performance objectives and other criteria in DOE Order 435.1 differ in some respects from those in NRC’s 10 CFR Part 61, as described in Section 3.4.2.1. First, for off-site releases, DOE specifies separate performance objectives for airborne releases of all radionuclides and for releases of radon. Those performance 12

Removal of a requirement for protection of inadvertent intruders from performance objectives in Order 435.1 was based on the view that performance objectives should be applied only to undisturbed performance of disposal systems, and that intrusion into a near-surface facility is not an acceptable condition but should be averted.

56 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT objectives are consistent with EPA standards for airborne emissions of radionuclides from DOE facilities in 40 CFR Part 61 (EPA, 1989a). Second, DOE has not developed a generally applicable waste classification system to provide protection of inadvertent intruders but, rather, requires site-specific assessments of inadvertent intrusion for the purpose of establishing limits on concentrations of radionuclides for disposal at each site. Differences in NRC and DOE approaches to protection of inadvertent intruders are discussed further in Section 6.2.2. Third, NRC regulations do not include separate requirements to address protection of water resources. Rather, adequate protection of water resources is assumed to be provided by dose criteria in the performance objective that applies to all pathways of exposure of the public. Finally, NRC regulations do not include requirements to address combined impacts on the public of a low-level waste disposal facility and other sources that could overlap. This would not be an issue as long as low-level waste disposal facilities are located away from other potential sources of public exposure at future times. 3.4.2.3 Implementation of the ALARA Requirement. Current NRC and DOE regulations for disposal of low-level waste discussed in the previous two sections both call for maintaining releases to the environment and exposures of the public ALARA. For operating nuclear facilities (e.g., nuclear power plants), compliance with that objective often is evaluated based, in part, on analyses of costs of reducing releases of radionuclides to the environment in relation to benefits in health risks averted in the general population (i.e., on an optimization of collective dose). However, as indicated by discussions of current ICRP recommendations on disposal of radioactive waste (ICRP, 1997b; 1998) in Section 3.3, limitations in the use of collective dose in situations that involve exposure at times far in the future have been widely recognized; see also IAEA (1996b) and NCRP (1995). DOE’s approach to evaluating an ALARA objective in disposal of low-level waste is to investigate the cost-effectiveness of alternative disposal concepts (e.g., use of concrete vaults instead of earthen trenches), improvements in facility design for the chosen disposal concept (e.g., an increase in cover thickness or other engineered barriers, use of more robust waste packages), and changes in facility operations (e.g., additional requirements on waste treatment) in reducing projected releases and doses (DOE, 1999c). Such evaluations can focus on projected doses to individuals or to a static population, which is fixed at current levels and locations, and qualitative considerations of whether various more costly alternatives

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would result in substantial reductions in projected individual or population doses. However, if projected impacts in a base-case performance assessment are sufficiently low (e.g., projected annual doses to individuals are well below applicable performance objectives), the cost of evaluating a different disposal concept could outweigh potential benefits in reduced impacts on the public, and further analysis would not be required (DOE, 1996a). NRC regulations and staff guidance on performance assessment (NRC, 2001) do not specify particular approaches to implementing an ALARA objective for disposal of low-level waste. However, consistent with an approach to implementing that objective in regulations for cleanup of radioactively contaminated sites in 10 CFR Part 20, Subpart E (NRC 1997), NRC favors an approach of evaluating effects of alternatives on reductions in doses to maximally exposed individuals and does not consider reductions in collective dose; see also NCRP (2004). 3.4.2.4 EPA Views on Requirements for Disposal of Low-Level Waste. The Atomic Energy Act (AEA, 1954), as amended, gives EPA the authority to establish environmental standards for disposal of low-level waste. However, EPA has not developed such standards. EPA standards for disposal of spent nuclear fuel, high-level waste, and transuranic waste in 40 CFR Part 191 (EPA, 1985; 1993a) and 40 CFR Part 197 (EPA, 2001a), provisions on protection of groundwater in EPA standards for control and cleanup of uranium mill tailings in 40 CFR Part 192 (EPA, 1995a), and EPA guidance on cleanup of radioactively contaminated sites under authority of AEA and CERCLA (Luftig and Weinstock, 1997) indicate that EPA has historically considered two performance objectives to be appropriate in situations similar to near-surface disposal of low-level waste: (1) a limit on annual effective dose equivalent of 0.15 mSv to an off-site member of the public from all release and exposure pathways, and (2) protection of underground sources of drinking water in accordance with standards for radioactivity in public drinking water supplies in 40 CFR Part 141 (EPA, 1976; 2000b). EPA’s drinking water standards for radionuclides, which apply to natural background as well as man-made sources, are summarized in Table 3.4; the history of the development of these standards is discussed elsewhere (EPA, 1991a; 1997; 2000a; 2000b; NCRP, 2004). EPA’s preferred performance objective for protection of off-site members of the public is somewhat less than the annual effective dose equivalent of 0.25 mSv used by NRC and DOE and is intended to correspond to a lifetime cancer risk of ~10–4, which is consistent

58 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT TABLE 3.4—EPA standards for radioactivity in public drinking water supplies [40 CFR Part 141 (EPA, 1976; 2000b)]. Standard

Interim standards (EPA, 1976)

Criteria

Concentration limitsa • 0.2 Bq L–1 for 226Ra plus 228Ra • 0.6 Bq L–1 for gross alpha-particle activity, including 226Ra but excluding radon and uranium Dose limit • annual dose equivalent of 0.04 mSv to whole body or any organ from man-made beta/gamma-emitting radionuclidesb

Final standards (EPA, 2000b)c

Maintain interim standards (EPA, 1976) for radium, gross alpha-particle activity, and man-made beta/gamma-emitting radionuclides Concentration limit of 30 µg L–1 for uraniumd

a

Concentration limits are obtained from values specified in 40 CFR Part 141 in units of pCi L–1 (1 pCi = 27 Bq); specified concentration limits for radium and gross alpha-particle activity are 5 pCi L–1 and 15 pCi L–1, respectively. Standards for radium and gross alpha-particle activity apply to natural background and man-made sources. b Radionuclide-specific concentration limits, except limits for 3H and 90Sr, are calculated on the basis of a daily intake of drinking water of 2 L and dose coefficients for ingestion of radionuclides (i.e., 50 y committed dose equivalents per unit activity intake) given in NCRP Report No. 22 (NCRP, 1959), as amended in August 1963. Radionuclide-specific concentration limits (i.e., maximum contaminant levels) so calculated, rather than specified limit on dose to the whole body or any organ, are operational standards for man-made, beta/gamma-emitting radionuclides. Operational standards for 3H and 90Sr are 740 and 0.30 Bq L–1, respectively. c Proposed standard for radon in drinking water was issued in 1991 (EPA, 1991a), but was later withdrawn (EPA, 1997). Another proposed standard for radon issued in 1999 (EPA, 1999a) has not been promulgated as a final rule. Proposed standard for radon in drinking water includes maximum contaminant levels of 11 or 150 Bq L–1. Higher limit would apply if a state has adopted or implemented an approved multimedia mitigation program to address radon in indoor air. dRelationship between mass and activity of uranium (and, therefore, radiation dose) depends on relative abundance of naturally occurring isotopes 234U, 235 U, and 238U in drinking water. Typical activity-to-mass ratios for uranium in drinking water are in the range of 0.025 to 0.056 Bq µg–1 and average ratio is expected to be ~0.033 Bq µg–1 (EPA, 2000b). Thus, 30 µg L–1 of uranium in drinking water is expected to correspond approximately to 1 Bq L–1.

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with EPA’s overall risk-management policies. EPA’s view that groundwater resources should be protected in accordance with drinking water standards is acknowledged in DOE standards and is included in most performance assessments at DOE sites, but is not incorporated in NRC regulations. Rather, it is NRC’s judgment that the performance objective for off-site members of the public from all release and exposure pathways provides adequate protection of water resources. 3.4.2.5 Implications of Performance Objectives. Performance objectives for near-surface disposal of low-level waste established by NRC or DOE and discussed above have important implications for the conduct of performance assessments. First, performance objectives apply to the overall performance of disposal systems, and separate performance objectives for particular natural or engineered barriers are not specified. This approach differs from the first licensing requirements for geologic repositories for disposal of high-level wastes developed by NRC in 10 CFR Part 60 (NRC, 1981b), which included separate subsystem performance objectives for waste packages, the engineered barrier system, and local groundwater flow.13 The absence of subsystem performance objectives allows considerable flexibility in complying with performance objectives within constraints imposed by siting and design requirements. Siting and design of disposal facilities, development of waste acceptance criteria including requirements on waste treatment and packaging, and development of methods of performance assessment then can focus on those attributes of a particular site and disposal technology that contribute most importantly to overall safety of the system. Second, performance objectives for exposure of off-site members of the public are expressed in terms of annual dose to individuals. Therefore, maximum concentrations of radionuclides at assumed receptor locations must be calculated. Furthermore, time histories of concentrations are required to calculate the maximum annual dose from all radionuclides combined. The types of analyses needed to calculate maximum concentrations at any location differ substantially from analyses that could be used to demonstrate compliance with containment requirements for disposal of spent nuclear fuel, 13NRC regulations in 10 CFR Part 60 will not be applied to the proposed geologic repository for disposal of spent nuclear fuel and high-level waste at the Yucca Mountain Site in Nevada. NRC regulations that apply to the Yucca Mountain Site are given in 10 CFR Part 63 (NRC, 2001); those regulations do not specify performance objectives for particular natural or engineered barriers.

60 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT high-level waste, and transuranic waste in EPA’s 40 CFR Part 191 (EPA, 1985; 1993a), which are expressed as limits on cumulative releases of radionuclides over 10,000 y.14 In calculating cumulative releases, detailed time histories of releases and resulting concentrations of radionuclides in the environment are not required. Another difference between standards for low-level waste disposal and those for disposal of spent nuclear fuel, high-level waste, and transuranic waste in 40 CFR Part 191 (EPA, 1985; 1993a) and 40 CFR Part 197 (EPA, 2001a) that can affect approaches to performance assessment is the following. At geologic repositories, the boundary that defines locations at which compliance with performance objectives must be demonstrated is a substantial distance from emplaced waste. An exposed individual is assumed to be located at a distance up to ~18 km from the proposed facility for disposal of spent nuclear fuel and high-level waste at the Yucca Mountain Site in Nevada (EPA, 2001a; NRC, 2001), and the distance for compliance demonstrations for disposal of DOE’s defense transuranic waste at the Waste Isolation Pilot Plant (WIPP) Facility in New Mexico is up to 5 km (EPA, 1985; 1993a). Such substantial distances mean, for example, that regional-scale groundwater flow models generally are appropriate. In addition, modeling of releases from a disposal facility (the source term) can be a relatively unimportant aspect of performance assessment at sites with highly favorable geologic, hydrologic and geochemical conditions. At such sites, an ability of the natural geologic, hydrologic and geochemical system to inhibit transport of radionuclides away from a disposal location may provide adequate protection of the public even if there were no engineered barriers to releases from the facility.15 At 14EPA standards in 40 CFR Part 191 (EPA, 1985; 1993a) have been applied to disposal of DOE’s defense transuranic waste only. Those standards also include a separate performance objective for protection of off-site members of the public that is expressed in terms of an annual dose, as in NRC and DOE performance objectives for low-level waste disposal. Those standards will not be applied to the proposed geologic repository for disposal of spent nuclear fuel and high-level waste at the Yucca Mountain Site in Nevada. EPA standards that apply to the Yucca Mountain Site are given in 40 CFR Part 197 (EPA, 2001a); they do not include containment requirements expressed in terms of limits on cumulative releases of radionuclides. 15Nonetheless, use of multiple natural and engineered barriers in disposal facilities for spent nuclear fuel and high-level waste generally is required (NRC, 1981b; 2001), and the long-term performance of waste packages, waste forms, and other engineered barriers can be important to achieving acceptable overall performance at some repository sites, including the proposed Yucca Mountain Site in Nevada (Rickertsen, 2002).

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near-surface disposal facilities for low-level waste, the boundary at which compliance with performance objectives for off-site releases is evaluated often is assumed to be only 100 m from disposed waste. Thus, detailed hydrologic modeling on a local scale generally is appropriate, and modeling of the source term often is quite important in demonstrating acceptable performance. Finally, by expressing performance objectives in terms of dose, performance assessments do not need to consider probabilities of occurrence of assumed exposure scenarios. Rather, exposure scenarios can be assumed to occur with a probability of unity, provided that assumed scenarios are credible for the particular disposal site and facility design. In contrast, EPA’s containment requirements for disposal of spent nuclear fuel, high-level waste, and transuranic waste in 40 CFR Part 191 (EPA 1985; 1993a), which apply to disposal of DOE’s defense transuranic waste at the WIPP Facility in New Mexico, require explicit consideration of probabilities of occurrence of exposure scenarios, and EPA standards for disposal of spent nuclear fuel and high-level waste at the Yucca Mountain Site in 40 CFR Part 197 (EPA, 2001a) specify that assumed exposure scenarios should be weighted by their probability of occurrence in demonstrating compliance with an annual dose criterion that applies to normal (undisturbed) performance. 3.4.2.6 Requirements of States and State Compacts. The Low-Level Radioactive Waste Policy Act (LLRWPA, 1980), as amended (LLRWPAA, 1985), specifies that disposal of most low-level waste not generated at DOE sites is the responsibility of states or State Compacts. These authorities may establish their own standards for disposal of low-level waste, and those standards may differ from those established by NRC (1982b; 1993), provided the two sets of requirements are compatible and requirements of the state or State Compact are at least as restrictive as those of NRC. Standards for low-level waste disposal established by states or State Compacts were not reviewed for this Report. 3.4.3

Unresolved Issues in Performance Objectives for Low-Level Waste Disposal

As indicated in the previous section, performance objectives for disposal of low-level waste have been established for some time. Nonetheless, a number of issues regarding performance objectives and their implementation remain unresolved. Many of these issues have a significant impact on suitable approaches to performance assessment.

62 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT 3.4.3.1 Time Period for Compliance. A uniform policy regarding the time period over which performance objectives for disposal of low-level waste should be applied has not been established. Current NRC and DOE regulations address this important issue differently. NRC regulations in 10 CFR Part 61 (NRC, 1982b; 1993) do not address the time period for compliance with performance objectives, thus implying that there is no time cutoff. NRC staff have recommended that performance objectives should be applied for 10,000 y, but that calculations should be extended in time if projected doses for particular radionuclides are increasing at 10,000 y (NRC, 2000). Results of such calculations could be used to establish site-specific limits on inventories of long-lived radionuclides, especially if projected doses exceed the annual dose limit for the public in 10 CFR Part 20 (NRC, 1991a) of 1 mSv (NRC, 2000). As noted in Section 3.4.2.1, however, NRC staff recommendations have not been approved by NRC and, thus, do not constitute official regulatory policy. DOE Order 435.1 specifies that performance objectives apply for 1,000 y (DOE, 1999b); this time period also is applied to intruder dose analyses used to establish site-specific limits on concentrations of radionuclides (Section 3.4.2.2). Calculations should be carried out to the time of maximum dose, if it is beyond 1,000 y, but such doses would not be used to establish site-specific limits on disposals of radionuclides (DOE, 1999c). Calculations over longer time frames could be used, for example, to choose between disposal alternatives that have comparable projected impacts over 1,000 y. DOE’s position on the issue of time of compliance is based on a consideration that results of performance assessments are but one factor in determining acceptable waste disposal practices. Other important factors include the need for active and passive institutional controls and associated legal and societal responsibilities, such as the need to prioritize allocation of finite resources to environmental problems, and the concept of a “rolling present” (Finger et al., 1997).16 An important implication of DOE’s position is that performance assessments and composite analyses must be continually maintained. Unless a low-level waste disposal site can be released for unrestricted use by the public under conditions that meet requirements of DOE Order 5400.5 (DOE, 1990), the site must be maintained under DOE control. During a period of DOE control of a site, performance assessments and composite analyses will be continually updated, and needed corrective measures can be identified. When a disposal facility contains high concentrations of long-lived radionuclides, the required period of DOE control could be indefinite.17

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Resolution of the issue of time of compliance does not have a firm technical basis, but is primarily a matter of public policy.18 This issue is potentially important in determining acceptable near-surface disposals of long-lived and relatively immobile radionuclides and radionuclides, such as isotopes of uranium, that decay to radiologically significant decay products that build up in waste only at times far in the future (Kocher, 1995). 3.4.3.2 Inclusion of Doses Due to Radon. NRC and DOE regulations for disposal of low-level waste are inconsistent in regard to whether doses due to radon are included in performance objectives. This issue is potentially important in determining acceptable disposals of waste that contains radium, thorium and uranium, although the importance for uranium also depends on the time period over which compliance with performance objectives must be demonstrated (Kocher, 1995; McDowell-Boyer et al., 2000; ORNL, 1997a). NRC’s 10 CFR Part 61 (NRC, 1982b; 1993) does not address radon. However, NRC staff have recommended that the dose due to radon should be included in complying with the performance objective for off-site releases (NRC, 2000). On the other hand, NRC expects that non-DOE low-level waste will not contain significant quantities of radium, thorium or uranium (NRC, 1981a; 1982a), in which case inclusion of the dose due to radon should not have a significant impact on acceptable waste disposals. DOE Order 435.1 (DOE, 1999b) includes a separate performance objective for releases of radon, with the dose due to radon and its short-lived progeny excluded from other performance 16The concept of a “rolling present” is embodied in the “chain of obligation” principle, whereby each generation’s primary obligation is to provide for the needs of the living and next succeeding generations (Finger et al., 1997). This principle allows for incremental decision making, meaning that the current generation should provide the next generation with the skills, resources and opportunities to deal with any problems the current generation bequeaths, and the next generation should do likewise for the next succeeding generation. In this way, future generations are considered and compensated for any harms passed on by the previous one. The “rolling present” thus involves an iterative decision process, in which succeeding generations have the responsibility to reevaluate policies and decisions of the past using their own values and priorities, and can incorporate new knowledge and facts to make appropriate policy changes. The concept of a “rolling present” essentially means that DOE’s time of compliance is interpreted as 1,000 y from a present that is advancing in time, rather than fixed in time.

64 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT objectives (Table 3.2). Performance measures used in establishing site-specific limits on concentrations of radionuclides for disposal on the basis of an intruder dose analysis (Section 3.4.2.2) also exclude the dose due to radon. Consideration of radon separately from other radionuclides is consistent with federal radiation protection standards for the public (DOE, 1990; NRC, 1991a), EPA standards for airborne emissions of radionuclides at DOE facilities in 40 CFR Part 61 (EPA, 1989a), and EPA standards for control and cleanup of uranium and thorium mill tailings in 40 CFR Part 192 (EPA, 1983). 3.4.3.3 Performance Objective for Protection of Groundwater. Current NRC and DOE regulations differ in regard to requirements for protection of groundwater. As discussed in Section 3.4.2, NRC’s 10 CFR Part 61 (NRC, 1982b; 1993) does not address protection of groundwater, whereas DOE Order 435.1 requires a separate analysis of impacts of low-level waste disposal on water resources and compliance with applicable federal, state and local laws and regulations, as well as other DOE requirements for environmental protection (DOE, 1988b; 1990). Protection of groundwater is a cornerstone of EPA policies on protection of the environment (EPA, 1991b). In addition, in its regulatory standards for radioactive waste disposal and in cleanup of radioactively contaminated sites, EPA has historically incorporated groundwater protection objectives that are consistent with 17This

especially could be the case if release of DOE’s low-level waste disposal sites is subject to requirements of CERCLA (1980) and EPA’s implementing regulations in 40 CFR Part 300 (EPA, 1990), given that EPA’s goal for lifetime cancer risk of 10–4 under CERCLA is well below lifetime risks that correspond to performance measures for inadvertent intruders used by DOE to establish limits on concentrations of radionuclides at each disposal site (Table 3.2). Indefinite control of near-surface disposal facilities for non-DOE low-level waste also could be required, given that the dose criterion of 0.25 mSv y–1 for unrestricted release of contaminated sites established by NRC in 10 CFR Part 20, Subpart E (NRC, 1997), is well below dose criteria that were used to establish limits on concentrations of radionuclides in NRC’s waste classification system in 10 CFR Part 61 (Section 3.4.2.1). NRC’s 10 CFR Part 61 requires that proposed near-surface disposal sites be owned by a state or the federal government before a license can be issued (NRC, 1982b), and Section 151 of the Nuclear Waste Policy Act (NWPA, 1982), as amended, provides for federal ownership of commercially operated low-level waste disposal sites following license termination if necessary or desirable in order to protect public health and safety and the environment.

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standards for radioactivity in public drinking water supplies (Section 3.4.2.4 and Table 3.4). Protection of groundwater near waste disposal sites also is an important concern of states in carrying out their responsibilities for ensuring the safety of public drinking water supplies under authority of the Safe Drinking Water Act (SDWA, 1974), and states may impose requirements that are consistent with drinking water standards. However, there is not a uniform set of federal requirements on protection of groundwater that apply at all low-level waste disposal sites. Inclusion of a separate performance objective for protection of groundwater is potentially important in determining allowable releases of radionuclides from any disposal facility; this would also be the case if the performance objective were applied to protection of surface water as well (ORNL, 1997a). Application of drinking water standards summarized in Table 3.4 to protection of groundwater resources at low-level waste disposal sites has been controversial, in large part because effective doses and cancer risks corresponding to the standards, especially standards for beta/gamma-emitting radionuclides, are much lower than dose limits or dose constraints in radiation protection standards for the public, including the primary public dose limit of 1 mSv y–1 and a dose constraint for individual sources of 0.25 mSv y–1 recommended by NCRP (1993); see NCRP (2004) for further discussion. This controversy perhaps can best be understood 18EPA standards for disposal of spent nuclear fuel, high-level waste, and transuranic waste (EPA, 1985; 1993a; 2001a) include a time of compliance with performance objectives of 10,000 y; that time of compliance also is included in NRC standards in 10 CFR Part 63 (NRC, 2001) that apply to the Yucca Mountain Site in Nevada. However, in a decision dated July 9, 2004, the U.S. Court of Appeals for the District of Columbia Circuit found that the 10,000 y compliance period selected by EPA for its standards in 40 CFR Part 197 that apply to the Yucca Mountain Site (EPA, 2001a) violated Section 801 of the Energy Policy Act of 1992 (ENPA, 1992), which directed EPA to establish standards for that site that were based upon and consistent with findings and recommendations of the National Academy of Sciences (NAS/NRC, 1995). NAS found that there was “no scientific basis for limiting the time period of the individual risk standard to 10,000 y or any other value.” At the time of this writing, EPA had not revised 40 CFR Part 197 to address the Circuit Court’s ruling and to be consistent with NAS’s findings and recommendations on the matter of time of compliance for the Yucca Mountain Site. The Court’s ruling does not affect EPA standards in 40 CFR Part 191 (EPA, 1985; 1993a) that apply to disposal of DOE’s defense transuranic waste at the WIPP Facility in New Mexico, and it does not apply to disposal of low-level waste.

66 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT and resolved by recognizing that EPA has applied drinking water standards for radionuclides to groundwater resources at waste disposal sites as part of a larger pollution prevention policy to protect current and future sources of drinking water, rather than as a direct means of providing radiation protection of the public. 3.4.3.4 Interpretation of Performance Objectives for Compliance Purposes. Performance objectives for low-level waste disposal in NRC and DOE regulations include criteria that are expressed in terms of dose. Such a performance objective implies that calculated doses below the criteria indicate regulatory compliance, but calculated doses above the criteria do not. However, given the substantial uncertainty in any performance assessment, fixed performance objectives may pose a significant challenge to regulatory authorities in judging compliance. In the presence of uncertainty, results of assessments that attempt to portray the state of knowledge about the performance of disposal systems by taking uncertainties in models and parameters into account cannot be expressed as a single expected outcome. Rather, results must be expressed as estimated ranges or probability distributions of projected outcomes, or as statements of confidence (credibility) that the actual outcome would not exceed some value. Unless disposal of only trivial amounts of radionuclides is allowed, assessments that take uncertainties into account usually will produce some outcomes that exceed performance objectives, although the estimated likelihood may be low. Such projected outcomes do not necessarily indicate noncompliance with performance objectives, but regulatory authorities must make a judgment about whether the facility is in compliance. The problem of comparing highly uncertain model calculations with fixed performance objectives does not necessarily indicate that a different type of standard for disposal of low-level waste should be developed, such as a probabilistic standard as currently specified in containment requirements for disposal of spent nuclear fuel, high-level waste, and transuranic waste (EPA, 1993a), which apply to disposal of DOE’s defense transuranic waste at the WIPP Facility in New Mexico. As noted in Section 3.4.2.1, a recommendation by NRC staff, which has not been approved by NRC as official regulatory policy, is that when a probabilistic approach to performance assessment is used, the highest mean annual dose in any year should be less than a performance objective of 0.25 mSv y–1 and the upper 95th percentile of a probability distribution of projected doses should be less than the annual dose limit for the public of 1 mSv y–1 (NRC, 2000). Regardless of whether or not a

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probabilistic approach to performance assessment is used, the problem of dealing with uncertainty in regulatory decision making can be addressed using the concept of reasonable assurance (Section 3.5.3). This concept, which is used by DOE (1999c), acknowledges that decisions about regulatory compliance involve some degree of judgment and, thus, should not be based only on objective comparisons of model calculations with performance objectives. In essence, decisions about regulatory compliance are no different than performance assessments themselves, because both represent judgments about disposal systems in the presence of uncertainty. 3.4.4

Other Approaches to Regulating Waste Disposal

This Section considers a variety of alternative approaches to regulating waste disposal. Alternative approaches to regulating disposal of radioactive waste that have been considered or implemented are discussed in Section 3.4.4.1, and Section 3.4.4.2 discusses the current approach to regulating disposal of hazardous chemical wastes in the United States. 3.4.4.1 Approaches to Regulating Radioactive Waste Disposal. Dose to individual members of the public has been the measure of performance used to judge the acceptability of low-level waste disposals in the United States (Section 3.4.2). In addition, performance objectives for low-level waste disposal expressed in terms of dose apply to the overall performance of disposal systems, and separate performance objectives for particular natural or engineered subsystems are not included. This Section discusses other types of performance objectives or measures of disposal system performance that have been considered in regulating radioactive waste disposal. ICRP (1997b; 1998) has recommended that risk to individuals in critical groups is an appropriate measure of disposal system performance when exposure scenarios are expected occur with a probability in any year substantially less than one [i.e., exposures should be considered as potential rather than virtually certain to occur (Section 3.3)]. If risk is used as a performance measure, probabilities of occurrence of exposures according to postulated scenarios would need to be considered, as well as projected consequences of such scenarios. The National Academy of Sciences/National Research Council (NAS/NRC, 1995) recommended that a performance objective expressed in terms of risk to individuals is appropriate for disposal of high-level wastes at the Yucca Mountain Site in Nevada. This recommendation was based in part on a desire to develop a standard that would not need to be changed to provide a particular

68 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT level of health protection if knowledge of cancer and other health risks at low levels of exposure changes in the future. A variation on the use of risk in performance objectives for waste disposal systems has been developed by regulatory authorities in the United Kingdom (HMIP, 1995). Risk is adopted as the preferred measure of performance, rather than dose, based in part on an argument that only by considering probabilities of occurrence of processes and events that affect the performance of a disposal system can a reasonably complete understanding of impacts of disposal on human health be obtained. However, instead of specifying an upper bound on allowable risk, an annual risk of 10–6 would be specified as a target (goal) for waste disposal systems. If regulatory authorities would be satisfied that good engineering and science have been applied in developing a disposal facility and that an estimated risk would be below the specified target, no further reductions in risk, based on the ALARA principle, would be sought and the facility would be judged acceptable. A facility also could be judged acceptable if an estimated risk is above the target, but only if regulatory authorities were satisfied that an appropriate level of safety is assured and that any further improvements in safety could be achieved only at a disproportionate cost. Regulations do not define “an appropriate level of safety” for waste disposal systems (i.e., the maximum risk above the target that would be acceptable). However, consistent with ICRP recommendations on radiation protection of the public (ICRP, 1991), the maximum acceptable risk presumably would not exceed the risk target by more than about two orders of magnitude. The approach described above of using a specified risk as a target for waste disposal, rather than an upper bound on allowable risk, is based on a previous finding that reliance cannot be placed exclusively on quantitative estimates of risk to determine whether a disposal facility is safe (Department of the Environment, 1995). While risk calculations can inform a judgment on the safety of a facility, other technical factors, including some of a more qualitative nature, will also need to be considered. Therefore, it is considered inappropriate to rely on a specified risk as an acceptance criterion for a disposal facility after control over the facility is withdrawn, but it is considered appropriate to apply a risk target in the design process for a disposal facility. The decision process defined by regulations for waste disposal in the United Kingdom is similar to the decision process for remediation of radioactively contaminated sites in the United States under authority of CERCLA (1980) and EPA’s implementing regulations in 40 CFR Part 300 (EPA, 1990); see NCRP (2004) for further discussion.

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Other supplementary measures of disposal system performance could be used in making the case that waste disposal would be safe, in spite of large uncertainties in evaluating system performance at far future times. Rather than relying on a single measure of performance, such as dose or risk, IAEA (1994b) noted that the case for long-term safety can be made most effectively by use of several performance measures, called safety indicators; these include risk, dose, environmental concentrations, fluxes into the biosphere, fluxes through various barriers in a disposal system, and time itself. Such safety indicators become particularly valuable when they can be supported by observations from natural analogs (e.g., environmental concentrations of naturally occurring radionuclides and fluxes into the biosphere). IAEA emphasizes, however, that risk and dose are the most fundamental safety indicators and should always be used. These ideas are incorporated in current ICRP recommendations discussed in Section 3.3. IAEA (1994b) also discussed the usefulness of various safety indicators in different time periods after disposal. Those discussions are concerned with disposal of high-level wastes in geologic repositories, but they are relevant to near-surface disposal of low-level waste to the extent that low-level waste contains significant amounts of long-lived radionuclides. In the time period up to ~104 y, IAEA recommends that the case for safety should be based on quantitative performance assessments using dose or risk calculations, supported by calculations of other safety indicators. For the time period of ~104 to 106 y, the case for safety can still be based on quantitative performance assessments. However, calculations should be simplified and stylized, and such calculations should only be viewed as illustrative and calculated doses as indicative. Finally, little credibility can be attached to integrated safety assessments for time periods beyond ~106 y. Given the performance objectives for low-level waste disposal in the United States discussed in Section 3.4.2, IAEA’s recommendations on use of safety indicators other than dose or risk have not been used in performance assessments of particular disposal facilities. Use of other safety indicators would become less important when regulations specify a time period for compliance with performance objectives of 1,000 or 10,000 y. 3.4.4.2 Approach to Regulating Disposal of Hazardous Chemical Waste. Under authority of the Resource Conservation and Recovery Act (RCRA, 1976), waste containing hazardous chemicals has been regulated by EPA (and by states with regulatory programs approved by EPA) on the basis of a set of legal and regulatory

70 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT requirements that differ from requirements for disposal of lowlevel radioactive waste. Those differences have potentially important implications for the role of performance assessment in determining acceptable disposals of low-level radioactive waste under AEA, even though RCRA specifies that waste containing source, special nuclear, and byproduct materials, as defined in AEA (1954), is excluded from regulation under RCRA. The potential importance of RCRA and its implementing regulations to performance assessments of low-level waste disposal facilities arises from the existence of waste that contains both radioactive material (i.e., source, special nuclear, or byproduct materials) and hazardous chemicals. Such materials are referred to as “mixed waste” and are subject to dual regulation under AEA and RCRA. The history and current status of mixed waste regulation and difficulties in managing and disposing of mixed waste, especially mixed low-level waste, are discussed elsewhere (NAS/NRC, 1999a; NCRP, 2002). Issues of dual regulation also arise in management and disposal of waste containing radionuclides regulated under AEA and hazardous chemicals regulated under the Toxic Substances Control Act (TSCA, 1976) including, for example, dioxins, polychlorinated biphenyls (PCBs), and asbestos, but those issues are not considered further here. The purpose of this discussion is to highlight areas in which the approach to regulation of hazardous waste under RCRA differs from the approach to regulation of low-level radioactive waste under AEA. RCRA and its implementing regulations lay out a comprehensive and complex management system for the generation, transport, treatment, storage and disposal of hazardous chemical wastes. The management system includes: (1) detailed reporting requirements to provide continuous accountability in handling of hazardous waste from generation to disposal (“cradle-to-grave”); (2) detailed and prescriptive technical standards for treatment, storage and disposal of hazardous waste; and (3) a permitting system requiring adherence to the technical standards and development of a closure plan as prerequisites for granting an operating permit to any hazardous waste management facility. As in management of radioactive wastes under AEA, an important objective of RCRA is protection of human health and the environment in all aspects of managing hazardous chemical wastes. However, the approaches to protecting human health and the environment under the two laws are quite different. Management of radioactive waste under AEA essentially is performance-based, meaning that facilities that manage or dispose of radioactive waste must comply with standards that specify

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acceptable overall performance, such as the performance objectives for disposal of low-level waste discussed in Section 3.4.2, but each facility has considerable flexibility in determining how compliance with those standards is to be achieved, taking into account such factors as the nature of the wastes and characteristics of a particular disposal site and facility. In contrast, management of hazardous chemical waste under RCRA essentially is technology-based, meaning that detailed and prescriptive technical standards are applied to every facility that generates, stores, treats or disposes of waste. For example, land disposal restrictions and universal treatment standards specified by EPA in 40 CFR Part 268 apply to every disposal facility without regard for the waste isolation capabilities of a disposal facility at a particular site. Similarly, operating standards for hazardous waste disposal facilities specified in 40 CFR Part 264 include detailed and prescriptive requirements for construction and operation of facilities, such as requirements for liner and leachate collection systems and facility closure, that apply at any disposal site, without regard for site characteristics. Thus, the approach to management of hazardous chemical waste under RCRA provides a high degree of uniformity in treating wastes prior to disposal and in developing, operating, and closing disposal facilities, and any waste that meets the treatment standards is acceptable for disposal in any permitted facility. For purposes of this discussion, perhaps the most important difference between management of radioactive and hazardous chemical wastes is that RCRA and its implementing regulations do not call for assessments of the long-term performance of specific waste disposal facilities in meeting the objective of protecting human health and the environment. Rather, technical standards for waste treatment and design and operation of waste disposal facilities were based only on generic assessments of health risks to the public that were conducted to support development of regulations, and protection of human health and the environment in disposal of hazardous chemical waste relies mainly on real-time monitoring at disposal sites. In regard to releases to groundwater, for example, each facility is required to comply with groundwater protection standards specified in 40 CFR Part 264. If measured concentrations of contaminants in groundwater exceed applicable standards, a site operator must undertake corrective actions to mitigate releases. RCRA also calls for risk assessments of reasonably foreseeable potential releases and accidents in applications for final permits at waste disposal facilities. However, such risk assessments have been applied only to the operating period of a disposal facility and not to the postclosure time phase.

72 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT Thus, under RCRA and its implementing regulations, sitespecific performance assessments of the kind discussed in this Report are not used to support decisions on permitting and closure of specific disposal facilities. Rather, the approach to protecting human health and the environment in disposal of hazardous chemical waste relies on generally applicable requirements on waste treatment, engineered barriers in the design of disposal facilities, monitoring of releases of contaminants, and maintenance of institutional control over disposal sites for as long as the waste remains hazardous.19 Differences in approaches to waste management under AEA and RCRA summarized above have led to difficulties in obtaining operating permits for treatment, storage and disposal facilities for mixed waste under RCRA (NCRP, 2002). In recognition of difficulties in managing mixed waste, as well as waste that contains hazardous chemicals and NARM, EPA has undertaken efforts to reduce the burden of dual regulation. Regulations in 40 CFR Part 266 (EPA, 2001b) specify conditions under which storage and treatment of mixed low-level waste at a generating site is exempt from RCRA requirements when a generator is licensed by NRC or an Agreement State and conditions under which such mixed low-level waste is exempt from RCRA requirements on waste manifests, transportation and disposal, except waste treatment standards under RCRA remain in effect. Those regulations also apply to chemically hazardous waste that contains NARM waste, provided the radioactive material is regulated by a state. Relaxation of requirements for storage and treatment do not apply to mixed waste managed by DOE, but provisions of 40 CFR Part 266 related to transportation and disposal do apply to DOE if DOE follows NRC or Agreement State requirements (e.g., on waste manifests). EPA also has indicated its interest in developing regulations that would allow disposal of low-activity mixed waste (i.e., waste that contains hazardous chemicals regulated under RCRA and low-level radioactive waste or NARM waste that contains low concentrations 19A

similar approach to regulation, except for requirements on waste treatment, is applied to disposal of uranium or thorium mill tailings in EPA’s 40 CFR Part 192 (EPA, 1982; 1983). Those regulations specify design objectives, which were based on generic assessments of health risks to the public and apply at all sites, and requirements on monitoring and corrective actions, especially in protecting groundwater, but site-specific performance assessments are not required. A technology-based approach to regulation of waste disposal thus is not unique to disposal of hazardous chemical waste.

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of radionuclides) in disposal facilities for hazardous chemical wastes permitted under RCRA (EPA, 2003). The intended benefits of such regulations would be to (1) ease the burden of dual regulation of mixed low-level waste by allowing disposal of low-activity wastes in RCRA facilities and (2) promote a more consistent framework for disposal of low-activity radioactive waste than is provided by the current approach of regulating non-DOE waste containing NARM separately from waste containing radioactive material regulated under AEA. Such regulations also could apply to DOE wastes. One outcome of efforts to ease difficulties in disposing of mixed low-level waste by allowing disposal of such waste in RCRA facilities may be to promote greater harmony in approaches to management and disposal of radioactive and hazardous chemical wastes. For example, there could be increased attention to the need to assess potential long-term impacts on the public of disposal of any hazardous wastes (radioactive or chemical) in near-surface facilities on a site-specific basis. Thus, performance assessments of the kind discussed in this Report may eventually be carried out at disposal facilities for hazardous chemical wastes.

3.4.5

Requirements for Protection of the Environment

Standards for low-level waste disposal discussed in Section 3.4.2 focus on protection of human health and protection of water resources that could be used by humans, but they do not include requirements directed at protection of plants and animals. This approach is consistent with a view that requirements for radiation protection of humans are adequate to ensure that populations of other plant and animal species would not be threatened, even though individual members of species may be harmed (IAEA, 1996b; ICRP, 1991). Requirements for radiation protection of the public and the environment established by DOE (1990) include an interim limit on absorbed dose to aquatic animals of 10 mGy d–1. In revising those requirements, DOE has indicated an intention to establish limits on absorbed dose to terrestrial plants and animals of 10 mGy d–1 and 1 mGy d–1, respectively, as well (DOE, 1996b; 2002). Performance assessments of low-level waste disposal facilities at DOE sites generally have not considered doses to aquatic or terrestrial biota, mainly because performance objectives and other performance criteria are much less than dose limits for biota and the latter are applied to populations of species at a site, rather than individual members of species.

74 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT 3.5 Other Concepts in Performance Assessment This Section discusses other concepts that are important to conduct of performance assessments of low-level waste disposal facilities, including institutional controls, model validation and confidence in model outcomes, and reasonable assurance of compliance with applicable performance objectives. 3.5.1

Institutional Controls

Institutional controls over low-level waste disposal facilities play an important role in ensuring protection of the public. Two general types of institutional controls, referred to as “active” and “passive,” may be applied. Although a distinction between active and passive institutional controls is not always clear-cut and it can be argued that all passive controls require some deliberate actions by institutions, it can be useful to distinguish between controls that involve deliberate human intervention or other actions at waste disposal sites and other controls that essentially involve maintaining societal knowledge of the existence of a disposal facility at a site. For example, monitoring, testing and research can be conducted at a disposal site during a period of active institutional control for the purpose of enhancing confidence that results of a performance assessment adequately represent the long-term performance of a facility. When sufficient information to support confidence in results of a performance assessment has been obtained, there can be a transition to more passive types of institutional control. 3.5.1.1 Active Institutional Controls. Active institutional controls over waste disposal sites include government ownership combined with use of fences and guards to prevent unauthorized entry. Regulations for disposal of low-level waste (DOE 1999b; NRC, 1982b) and the first regulations for disposal of spent nuclear fuel, high-level waste, and transuranic waste (EPA, 1985; 1993a; NRC, 1981b) specify that active institutional control will be maintained for at least 100 y after facility closure.20 Thus, there is an implicit assumption that active institutional controls cannot be relied upon beyond 100 y. Especially at near-surface disposal 20EPA and NRC regulations that apply to disposal of spent nuclear fuel and high-level waste at the Yucca Mountain Site in Nevada (EPA, 2001a; NRC, 2001) do not specify a minimum time over which active institutional control must be maintained.

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facilities for low-level waste, determinations of acceptable disposals on the basis of an assumption that active institutional controls will not be maintained beyond 100 y provide a kind of defense-indepth when active controls may be required for substantially longer times (Section 3.4.3.1). Maintaining active institutional control for a substantial period of time after facility closure serves two purposes. First, inadvertent human intrusion onto the disposal site is assumed not to occur during the period when concentrations of shorter-lived radionuclides and, thus, potential doses to an intruder are the highest. This assumption is embodied in approaches used by NRC and DOE to establish concentration limits of radionuclides in near-surface facilities (Section 3.4.2). Second, an environmental monitoring program can be established at a disposal site to provide an additional measure of confidence that the long-term performance of the facility will be acceptable, and maintenance of the facility or other remedial actions can be undertaken in the event that actual performance is poorer than expected. 3.5.1.2 Passive Institutional Controls. As noted above, a transition from active institutional controls that involve the presence of governing authorities at a disposal site and deliberate exclusion of the public to more passive types of control may occur when there is sufficient confidence in the long-term performance of a disposal facility that active controls are no longer considered necessary to protect the public. Passive institutional controls include government ownership without use of fences and guards, permanent markers at a disposal site, public records of disposal activities and requirements for periodic public disclosure of those activities, and legal restrictions on future land uses. Passive institutional controls are intended to enhance societal memory of waste disposal at a site while still allowing public access to the site and its immediate surroundings. Engineered barriers in disposal systems, which are intended to inhibit mobilization and transport of radionuclides and prevent inadvertent human intrusion, usually are not considered to be a form of passive institutional control. For some period of time after loss of active institutional control over a disposal site, passive institutional control presumably would have an effect in precluding inadvertent intrusion, as well as exposures of nearby residents who might choose not to live near a site if they were aware of the facility. An assumption that passive institutional control would preclude or reduce exposures usually has not been taken into account in performance assessments and in using results to establish waste acceptance criteria at specific sites. Nonetheless, NRC and DOE approaches to protection of

76 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT inadvertent intruders discussed in Section 3.4.2, which allow higher doses to inadvertent intruders than to off-site individuals, are based in part on an assumption that societal memory of disposal activities will be maintained for the foreseeable future. Inadvertent human intrusion thus is regarded as an accidental occurrence, rather than an event that is virtually certain to occur, in which case an allowance of higher doses is compensated by the lower probability that postulated exposure scenarios for inadvertent intruders would actually occur. 3.5.2

Model Validation and Confidence in Model Outcomes

Evaluations of the long-term performance of radioactive waste disposal systems necessarily involve use of mathematical or physical models, including natural analogs, because it is not practical to ensure protection of the public over long time periods by direct observation of a disposal facility and monitoring of all releases to the environment. While models used in performance assessments should, in some sense, be valid and extensive model validation exercises related to radioactive waste disposal have been undertaken (Larsson et al., 1997), the concept of model validation probably has only limited application to waste disposal. As used in this Report, the term “model validation” refers to comparisons of model predictions with relevant data that were not used in developing a model. This meaning of the term differs from that often used, for example, in computer sciences, where “validation” often refers to a confirmation that a code solves a problem as intended. This difference in meaning causes occasional confusion between code users and code developers, and the difference can be important if quality assurance guidelines require use of “validated” codes. An example of a model validation exercise for waste disposal is the following. Controlled laboratory experiments could be used to estimate model parameters that are important in predicting release rates of radionuclides from a waste form. Model validation then would involve placing the same waste form in an environmental setting that mimics a disposal facility and observing releases over time for comparison with model predictions. Field tests using lysimeters are one means of obtaining data to validate model predictions that are based on laboratory studies. Another example is use of field studies to confirm predictions of models of flow of water in the unsaturated zone (Young et al., 1999a; 1999b). Some types of model validation can be undertaken while maintaining a performance assessment during the period of active institutional control after closure of a disposal facility.

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There have been numerous international programs on comparisons of models with data (IAEA, 2004a; Larsson et al., 1997; SSI, 1996a; 1996b; 1996c) and intercomparisons of alternative modeling approaches (IAEA, 1995b; NEA, 1989); see also SENES (2005) and Thiessen et al. (1997). Those programs have contributed greatly to an understanding of performance assessment and the degree to which model validation and intercomparisons may be relied upon to develop confidence in results of performance assessment. Such studies also have contributed to the long evolution in thought on the issue of model validation in the radioactive waste management community (Kozak, 1994b). The following discussion reflects the current state of opinion on this controversial issue. The general consensus is that validation of models of waste disposal systems is inherently impossible (Davis et al., 1991; Kocher et al., 1985; Kozak and Olague, 1995; NAS/NRC, 1990; Oreskes et al., 1994; Seitz et al., 1990). This view is based on several considerations, including: (1) the paucity of data, which leads to ambiguity about the most appropriate approaches to modeling; (2) an inability to directly observe a disposal system of interest; (3) lack of knowledge of the future behavior of a disposal system; and (4) the high degree of site-specificity of phenomena of interest, which makes it difficult to transfer information obtained at one site to another (Kozak and Olague, 1995). As a consensus for this view developed beginning in the late 1980s, there was an attempt to redefine a “valid” model to mean a model that is adequate for purposes of regulatory decision making (Davis et al., 1991). However, this attempt led to frequent semantic and technical disputes between those who would redefine the term and those who were using it in its usual sense. As a result, there has been an increasing tendency not to discuss model validation, since it is recognized to be impossible. Rather than discussing model validation, there has been an increasing focus on the concept of developing confidence or credibility in the use of models for purposes of regulatory decision making (IAEA, 1999; Kozak, 1994a; Seitz et al., 1990; Vovk and Seitz, 1995). This concept is quite different from the more difficult task of using models to predict actual outcomes of waste disposal. In some cases, an important factor in developing confidence in models for decision making is a comparison of model predictions with independent data. However, few performance analysts attempt to claim that such comparisons are directed toward “validation” of models, or that positive results of such a comparison lend more than qualitative confidence about the appropriateness of models for use in performance assessment.

78 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT Achieving credibility in results of performance assessments for the purpose of regulatory decision making requires proper completion of three processes: quality assurance, model calibration, and evaluation of conservative bias. 3.5.2.1 Quality Assurance. For a performance assessment to be credible, information contained in an assessment and upon which the assessment is based must, to the extent possible, be documented and traceable. This information comes in several forms, including: (1) data, either site-specific or generic; (2) interpretations of data, which are embodied in assumptions; and (3) computer codes, in which data and assumptions are implemented. The purpose of quality assurance is to ensure that proper documentation in each of these areas is provided. Site-specific data must be obtained in a way that is well documented to provide confidence that an assessment is founded on legitimate measurements. Proper documentation also allows, for example, that extreme values of importance to an assessment can be reexamined to evaluate possible causes. In the case of commonly used generic data, such as bioaccumulation factors in models of terrestrial food-chain pathways, documentation is needed to show where the values came from and why they are appropriate for use at a particular disposal site. In assuring data quality, the emphasis should be on providing sufficient documentation to allow a skeptical outsider to trace the legitimacy of values used in an assessment. Perhaps the most difficult, but also the most essential, aspect of documenting a performance assessment is documentation of assumptions. Assumptions are an essential part of an assessment and are important in data interpretation, in developing conceptual models for a disposal system, and in implementing conceptual models of the system using mathematical or physical models. There is no standard guidance on the level of detail needed in documenting assumptions, but it is no longer considered acceptable simply to identify a computer code used and assume that regulators and other stakeholders will appreciate why that code is suited to a system of concern. On the other hand, assumptions generally need not be documented in minute detail, unless details are important to an argument for safety of a disposal system. The appropriate level of detail is that which will convince a skeptical audience that all significant alternatives for describing a disposal system and the consequences of waste disposal have been considered, and that a reasonable suite of assumptions has been developed. Documentation of computer codes is easier than documentation of assumptions. Standard approaches to production and

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documentation of codes have been developed in the software industry and by numerical analysts. Key techniques in these approaches include verification, benchmarking and assessment of numerical error. Verification refers to a determination that a particular computer implementation of mathematical equations to represent a model is without significant error. Codes can be verified, for example, by comparing code outputs with known analytical solutions of specified problems. Benchmarking refers to intercomparisons of outputs of different codes that purport to solve similar problems. To the extent that outputs of different codes agree and some of the codes have been verified, these methods can increase confidence in use of a code. The magnitude of errors associated with numerical methods used to solve differential equations in many computer codes, including finite-element, finite-difference, or NewtonRaphson methods, can be evaluated by using commonly available standard techniques. Documentation that numerical error has been addressed can also increase confidence that a numerical analysis has been performed properly. Although quality assurance of a performance assessment is important, a quality assurance plan should not be too detailed and prescriptive. Otherwise, a quality assurance plan may require so much attention that little work on an assessment is accomplished. It is important, therefore, to develop a quality assurance plan that contains only common-sense technical requirements, and to implement it in a common-sense fashion. The goal of a plan should be to assist in developing confidence in an assessment. 3.5.2.2 Model Calibration. Model calibration refers to a process of fitting a model to represent site-specific conditions using sitespecific data. Most often, model calibration is a process by which input data for a model are deduced from output data obtained from other sources including, for example, field or laboratory measurements. Those input data are used either to project the same type of output information under different forcing conditions or to project different types of output. An example of a common type of calibration used in groundwater analyses for performance assessments is calibration of an unknown spatial variability of hydraulic conductivities using hydraulic heads measured in the field. The calibrated conductivity field is then used to project the spatial distribution of hydraulic head under differing conditions (e.g., increased infiltration), or to project groundwater velocities or transport of radionuclides. Model calibration can be used in most aspects of performance assessment. However, the ability to calibrate a model may lead to

80 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT an unfounded level of confidence in applications of the model, particularly in the case of models of groundwater flow. There are three issues associated with model calibration that may limit the degree of confidence it lends to an analysis: nonuniqueness, extrapolation, and lack of observation of pertinent quantities. The issue of nonuniqueness is that there often is more than one possible set of conditions (and there may be an infinite number) that match existing data sets equally well. This issue is well known in the conductivity calibration example noted above, where it is termed the “inverse problem.” If there is an infinite number of possible hydraulic conductivity fields that match available data equally well, then there is little confidence that any one of them will be close enough to reality to permit extrapolation to different conditions. However, this is an important problem only if different credible alternatives give significantly different extrapolations. The issue of extrapolation is that if conditions of an analysis are changed, a model might then be applied to a regime in which a calibration has little meaning. For example, under conditions of increased infiltration, the water table might rise and saturate different geologic strata than those that were saturated when calibrating a model. This kind of situation can only be addressed by introducing additional untestable assumptions about the behavior of an altered system. The issue of extrapolation is complicated by the issue of nonuniqueness described above, because an analyst does not know whether the original model was a reasonable match to reality, or whether the match was fortuitous. The third issue is that models usually are calibrated on the basis of quantities that can be observed, but observable quantities are not necessarily the quantities of interest in performance assessment. In the example described above, an analyst is interested in groundwater velocity and radionuclide transport, not hydraulic head. In general, an analyst cannot know how well a model that was calibrated on the basis of measured hydraulic heads has matched an existing Darcy velocity field. For seepage velocity or radionuclide transport, the uncertainty is even greater. From this discussion, it is clear that calibration can provide only limited confidence in the ability of a model to provide realistic projections of disposal system performance. However, for many aspects of performance assessment, calibration remains the only available tool for developing a model that represents, however imperfectly, specific conditions at a disposal site. 3.5.2.3 Evaluation of Conservative Bias. The importance of evaluating conservative bias in a performance assessment follows from

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the recognition that the purpose of an assessment is to demonstrate compliance with regulatory requirements. If there is confidence that a disposal facility will comply with those requirements, then knowledge of the actual outcome (e.g., maximum dose to an individual member of the public) is not important to a regulatory decision. The important issue of achieving a reasonable balance between conservative bias and realism in performance assessment is discussed in Section 2.4. In demonstrating compliance with regulatory requirements, it may be desirable and, in some cases, necessary to choose conditions, within constraints provided by existing knowledge and data, that are believed to overestimate actual outcomes of waste disposal. However, extreme and highly unlikely assumptions generally should not be used. Rather, use of conservative assumptions should be tempered by a rational assessment of constraints provided by existing knowledge, leading to incorporation of some degree of “realism” in an analysis, to the extent that a more realistic assessment is practical and defensible. An appropriate balance between conservatism and realism in an assessment is difficult to achieve, since it sometimes is difficult to know if certain assumptions are indeed conservative and “realism” can never be determined. Nonetheless, successful resolution of this issue is at the core of the problem of developing performance assessments that are acceptable for purposes of regulatory decision making. In Section 4, a performance assessment process that provides an approach to achieving an appropriate balance between conservatism and realism is described. However, this process can never obviate the need for professional judgment by analysts and reviewers in determining the degree to which “realism” can be incorporated in an analysis. When regulatory authorities and any other stakeholders that have a defined role in the decision-making process judge that an appropriate balance between conservatism and defensible realism has been developed and that a disposal system complies with applicable regulatory requirements, it then can be said that “reasonable assurance” of compliance has been demonstrated. This important concept is described in the following section. 3.5.3

Concept of Reasonable Assurance

The essential purpose of performance assessment, and the purpose given greatest emphasis in this Report, is to provide a basis for determining whether a waste disposal facility complies with applicable performance objectives. Such determinations will be

82 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT based, in large part, on the use of models that are intended to provide a technically credible representation of a disposal system. However, given that there will always be uncertainty in model calculations used in demonstrating compliance with performance objectives, absolute assurance of compliance is not attainable by any means. This statement is true even when the intent is to use models with a conservative bias, unless one trivializes the problem by allowing disposal of only insignificant quantities of radionuclides. Uncertainties in performance assessments of waste disposal systems may be divided into two types: (1) those uncertainties for which quantification is possible and makes sense, and (2) those uncertainties for which quantification is not possible or does not make sense. Sometimes an uncertainty may have both quantifiable and unquantifiable components. It is important to understand that uncertainties in outcomes of waste disposal generally are not amenable to a rigorous and purely objective quantification (i.e., evaluations of uncertainty will always be subjective to a significant degree). Since absolute assurance of compliance with criteria that specify limits on long-term impacts from radioactive waste disposal cannot be achieved, except in trivial situations, and because scientific judgment is an essential component of regulatory decision making for waste disposal, an appropriate goal for regulatory acceptance of a disposal facility is to achieve “reasonable assurance” of long-term safety. This concept has long been used in waste disposal (NRC, 1981b) and is often called “reasonable expectation” in regulations for waste disposal in the United States (DOE, 1999b; EPA, 1985; 1993a; 2001a; NRC, 2001).21 Dealing with uncertainty is not unique to disposal of radioactive wastes. Determining the regulatory acceptability of a waste disposal facility is simply a particular example of technical decision making in the presence of uncertainty. However, waste disposal provides a unique challenge, due to the long time frames over which safety needs to be provided. This Section discusses how “reasonable assurance” should be interpreted and some of the means by which it can be achieved. 21

Proponents of the term “reasonable expectation” may argue that it provides a more realistic approach to regulatory decision making that does not rely on overly conservative assumptions when relevant data or other information are lacking. An assumption in this Report is that there is little, if any, practical difference in interpretation of the two terms as long as all assumptions are documented.

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There is some overlap between this discussion and those in Section 3.5.2 and later discussions in Section 4. However, there are additional factors of importance to achieving reasonable assurance of safety that are not directly associated with performance assessment, and these factors are discussed here. 3.5.3.1 Description and Interpretation of Reasonable Assurance. A critical issue in regulatory decision making for waste disposal systems is the level of assurance needed in demonstrating compliance with applicable performance objectives (i.e., in demonstrating long-term safety). As indicated above, the nature of uncertainties in performance assessment and a conclusion that absolute assurance of compliance generally cannot be achieved lead to a realization that “reasonable assurance” is the appropriate concept for decision making. It is inherent in the concept of reasonable assurance that this term cannot be given a precise definition. In this sense, “reasonable assurance” is similar to such concepts as “beyond a reasonable doubt” or “preponderance of evidence” used in courts of law.22 It also is inherent in this concept that the level of proof that performance objectives will be met cannot be defined in advance of a regulatory decision. Although this situation is potentially frustrating to a license applicant, due to the resulting uncertainty in conditions for regulatory acceptance of a disposal facility, it recognizes that subjective judgments are an essential aspect of the process of deciding on the acceptability of disposal facilities. Furthermore, the decision process is not simply a matter of invoking technical arguments, because decisions will be made in a social and legal setting in which additional factors, which cannot be specified in advance, will be important. Those factors include, for example, the possibility that performance objectives, and other relevant safety criteria, as well as public attitudes may change over time. On the other hand, it clearly is desirable for an applicant, regulatory authorities, and other stakeholders in a decision to engage in a continuing dialog from which the level of assurance to be required in making the case for safety will emerge. Some requirements related to achieving reasonable assurance of safety, 22A potentially important difference between the concept of “reasonable assurance” and similar concepts used in courts of law is that there is a substantial body of precedent for interpretation of the latter. Precedents for the meaning of “reasonable assurance” in regard to the safety of waste disposal systems can only be established by an accumulation of sitespecific regulatory decisions.

84 / 3. CONTEXT FOR PERFORMANCE ASSESSMENT although they may be hedged with caveats, should be specified in advance by regulatory authorities, since they will amount to general policy for acceptance or rejection of proposals for disposal facilities (e.g., the acceptability of particular calculational methodologies or data sets). Otherwise, the regulatory process may suffer from a lack of credibility with politicians and the public. Applicants and regulators may need to work together to resolve thorny issues, particularly those of a technical nature. The need for the concept of reasonable assurance is based on a realization that scientific judgment is an essential element of all aspects of performance assessment and, furthermore, that such judgments generally are subjective to a significant extent. Although complex calculations using sophisticated mathematical models, including quantification of uncertainties, often will be performed in an assessment, the validity of model outputs and their estimated uncertainties are limited by the validity of judgments used in developing conceptual models that provide the foundation for mathematical models. This is not to say that such calculations should not be performed. Rather, results need to be considered and interpreted in their proper context as an aid to regulatory judgments, and should not be used as a substitute for such judgments. Achieving reasonable assurance of safety of waste disposal systems results primarily from an interaction between analysts and regulatory authorities. The role of analysts is to make the case for safety on the basis of available information. The role of regulatory authorities then is to judge the adequacy of the safety case. Reasonable assurance thus may be viewed as equivalent to building confidence in a regulatory decision. 3.5.3.2 An Approach to Achieving Reasonable Assurance of Compliance. A multifaceted approach to achieving reasonable assurance of compliance with performance objectives for waste disposal systems in the presence of large and, to a significant extent, unquantifiable uncertainty in evaluating system performance has been discussed by IAEA (1997b) and is summarized in Table 3.5. Some aspects of this approach undoubtedly are more important than others in making the case for safety of waste disposal systems (e.g., use of multiple lines of reasoning consistent with available information, use of judgment, analyses of sensitivity and uncertainty, an overall analysis of system safety, quality assurance and peer review), and their relative importance probably depends on particular characteristics of each disposal system and the types of wastes intended for disposal therein. Nonetheless, use of the different aspects in a complementary manner should provide a means of achieving reasonable assurance of compliance.

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The process of achieving reasonable assurance of compliance with performance objectives and other similar regulatory requirements places considerable demands on the conduct of performance assessments. Probably the most important are the need to justify all assumptions, to demonstrate an understanding of the importance of all assumptions to results of an assessment, and to interpret results with respect to their relevance to a regulatory decision. Approaches to performance assessment that are potentially important in achieving reasonable assurance of compliance are emphasized in discussions in Section 5 on modeling of different components of a disposal system.

Factor

Description

Multiple lines of reasoning

Consideration of all assumptions, conceptual models, and databases that are credible and reasonably consistent with available information on performance of disposal facility; may include use of observations on natural systems to complement analytical techniques, simplified methods as well as more complete treatments, and screening models

Use of judgment

Recognition of importance of scientific judgment as essential element in demonstrations of compliance, and development of methods of performance assessment that consider judgments systematically; does not obviate the need to justify credibility of assumptions

Multiple barriers

Use of multiple and redundant natural or engineered barriers in disposal systems to compensate for uncertainties in predicting performance of individual barriers and to ensure safety should a barrier fail to function as represented in an assessment

Generic approaches

Use of generic modeling studies to support assumption that low-level waste disposal should be safe and to identify general characteristics of disposal systems that help ensure safety

Well-structured methods of assessment

Use of assessment methods that are structured to help ensure that all credible events and processes affecting disposal system performance are considered and that basis for scientific judgments is transparent, but without imposing an overly rigid structure or excessive formalization

Modular approach

Use of modular approach in constructing performance assessment models, based on phenomena being modeled, to facilitate descriptions of individual components of disposal system and their interfaces, examinations of intermediate results, sensitivity and uncertainty analyses, and replacement of subunits of an overall model

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TABLE 3.5—Factors of potential importance in achieving reasonable assurance of compliance with performance objectives for low-level waste disposal.a

Incorporation of overdesign (safety factors) in disposal facilities and use of assessment models that are tolerant to uncertainties in calculational methods and assumptions (e.g., initial and boundary conditions); level of robustness should be directly linked to importance of assessed processes or features to regulatory decision

Sensitivity and uncertainty analyses

Determination of model parameters or assumptions with greatest effect on model outputs and investigation of variability in model outputs due to uncertainties in model parameters or assumptions; important for demonstrating understanding of behavior of assessment models and increasing confidence in use of models for decision making

Different safety indicators and role of qualitative assessments

Use of safety indicators other than dose or risk in making case for safety (e.g., concentrations, transit times, fluxes, barrier performance, time); recognition that quantitative results should be complemented by qualitative discussions to provide proper context for interpretation of results and regulatory decision making in presence of uncertainty

Overall analysis of system safety

Analysis of total system performance (i.e., calculations of dose or risk) to determine attributes of system of importance to overall safety and required performance of different system components to achieve safety

Iterative approach to assessment

Revisions of assessments as new information on performance of disposal system becomes available, with intent of continuously improving confidence in assessment for regulatory decision making

Simplicity

Development of equivalent simplified representations of complex models and results that convey essence of an assessment in making case for safety; facilitates communication with regulators and other audiences with stake in regulatory decision

Quality assurance and peer review

Systematic and auditable documentation of assessment and development of disposal facility; independent technical appraisal of assessment to develop confidence in methods, models, judgments, and regulatory decisions

Acceptability to different audiences

Tailoring of assessment to different audiences with stake in decisions about acceptability of waste disposals

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a Adapted from approach to achieving reasonable assurance of safety for disposal of long-lived wastes in geologic repositories discussed by IAEA (1997b).

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Robustness of design and assessment

4. Framework for Performance Assessment This Section discusses a general approach to conducting performance assessments. A flexible framework for efficient and defensible conduct of the performance assessment process is emphasized. Section 4.1 discusses considerations related to selection of models and data sets, and Section 4.2 presents a suggested framework for conduct of performance assessments. The desire for a “cookbook” approach to conducting performance assessments has often been expressed. In such an approach, a particular computer code or set of codes to be used in all assessments would be specified by regulators, and a license applicant would simply run the codes using appropriate input data. This approach is used, for example, in evaluating compliance with EPA standards for airborne emissions of radionuclides (EPA, 1989a). Although parts of the performance assessment process may be standardized (e.g., selection of dose coefficients for inhalation and ingestion of radionuclides, requirements for documentation of data and assumptions), a completely standard approach is unworkable because it lacks the flexibility needed to consider important characteristics of specific sites and facility designs. For example, the link between the performance of engineered barriers and characteristics of a site usually would not be taken into account in a generic assessment. In order to make defensible decisions regarding regulatory compliance, the performance assessment process must result in a sufficient understanding of those aspects of a disposal system that are important to a licensing decision. A critical means of achieving such an understanding involves use of alternative conceptual models that are consistent with available information on a specific site. A process for choosing among plausible assumptions is more conducive to developing confidence in results than is use of a single modeling structure. A standardized approach to performance assessment also may introduce biases toward a particular type of site or facility that may not be cost-effective in protecting the public. Licensing of a disposal facility will involve intense scrutiny of the performance assessment process. This scrutiny often is not 88

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related to particular computer codes used, but is focused on conceptual models, assumptions, and data sets (RAEC, 1994). For example, the review process for the Ward Valley Site in California focused on the potential for migration of tritium gas and the potential for contamination of water supplies, not on computer codes used in the assessment. This example also illustrates that sitespecific issues will be a common concern in the licensing process. Thus, the performance assessment process should emphasize the need to exercise sound judgment and provide a thorough defense of assumptions made as part of the process, rather than development of a standardized approach to performance assessment. It is important to keep in mind the purpose of performance assessment throughout the process of data collection, conceptual model development, and modeling itself (Vovk and Seitz, 1995). Performance assessment is used to identify conditions under which reasonable assurance of compliance with regulatory requirements related to protection of the public can be demonstrated. Reasonable assurance of compliance requires a sufficient understanding of those aspects of a disposal system that are important to a licensing decision, but it does not require a perfect representation of all processes that affect long-term performance. Thus, the performance assessment process should focus on obtaining a reasonable and defensible representation of a disposal system for the purpose of demonstrating compliance with regulatory requirements. 4.1 Data Collection, Conceptual Models, and Mathematical Models Data collection, development of conceptual models for different aspects of the performance of disposal systems, and development of mathematical models to represent conceptual models and data sets and their implementation in computer codes, often by means of numerical methods, are integral parts of the performance assessment process. The relationship between different steps in an assessment is shown in Figure 4.1, which also indicates the level of subjectivity (judgment) associated with each step. The importance of subjective judgment is high in interpreting data and developing conceptual models, but is low in implementing conceptual models and data sets in mathematical models and codes. Figure 4.1 also indicates, as discussed above, that scrutiny of performance assessment is expected to focus on data and conceptual models, rather than computer codes used in an assessment. Codes and mathematical models must be properly tested, but they are not expected to be the main focus of reviews.

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Fig. 4.1. Logical ordering of different steps in formulating and implementing calculational models in the performance assessment process; importance of subjective judgment in each step and primary focus of reviewers are indicated.

4.1.1

Data Collection

Defensibility of a performance assessment begins with the quantity and quality of data used in the assessment. Site-specific data generally are necessary in developing conceptual models of a disposal facility (Case and Otis, 1988; NRC, 1991b). Both the quantity and quality of data are important. The existence of large amounts of data does not necessarily mean that data are defensible (Seitz et al., 1992b). Similarly, sparse data sets, regardless of their quality, generally are not sufficient to characterize a complex and highly variable disposal system. A variety of generic data is available in the literature for many parameters used in performance assessment calculations. Many computer codes provide generic data sets for use in an analysis

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(e.g., dose coefficients for intakes of radionuclides, parameters describing accumulation of radionuclides in terrestrial food chains, and exposure of humans). In some cases, those data are deliberately chosen to be conservative and, thus, may be useful in initial screening calculations to identify radionuclides or exposure pathways that do not warrant further consideration in an assessment. Despite the availability and appeal of generic data sets, their applicability to a disposal facility of concern must be justified. Unusual behaviors of radionuclides not captured in generic data can occur at some sites (e.g., bioaccumulation of radiocesium in environments deficient in potassium, retardation of iodine in transport in soils rich in organic matter). An intention to use conservative assumptions also must be justified, because such assumptions can be misleading in cases where counterintuitive behavior occurs. For example, low values of solid/solution distribution coefficients of radionuclides may be chosen in an attempt to provide conservative estimates of concentrations in groundwater and travel times to off-site locations. However, such assumptions may not be conservative for 238U and its progeny, due to long-term ingrowth of radiologically significant decay products. Site-specific data probably will be necessary for many aspects of a disposal facility that have a significant impact on projected performance. Information on a disposal site and its surroundings that is likely to be essential for a defensible assessment include, for example, characteristics of aquifers, geochemistry of soil and rocks, meteorology, and food preferences and other living habits of nearby populations. Data on any engineered systems used in a facility also will be essential. Data collection should be driven by needs of performance assessment. This view is a cornerstone of an iterative approach to performance assessment introduced in Section 2.1 and discussed in more detail in Section 4.2. Since data collection is costly, it is inefficient to obtain many types of site- and radionuclide-specific data at the start of the performance assessment process. In an iterative approach, results of initial modeling studies that are based on limited data are used to identify important data needs. Subsequent iterations then become more detailed in those areas of greatest concern, and they may require that additional data be collected. 4.1.2

Development of Conceptual Models

A conceptual model essentially is a way of thinking, or point of view, about the behavior of a system of concern; it is a description (in words) of assumptions about system behavior. Conceptual

92 / 4. FRAMEWORK FOR PERFORMANCE ASSESSMENT models are essential to performance assessment because they provide the basis for all mathematical models used in an assessment. In some cases, conceptual models are assumed implicitly and are rarely questioned. Examples include common assumptions that groundwater flow can be modeled using Darcy’s Law for porous media and that transport of radionuclides in groundwater can be modeled using equilibrium solid/solution distribution coefficients. In other cases, however, considerable effort is required to develop conceptual models for a particular disposal system and to defend them in licensing reviews. For example, even though use of Darcy’s Law may not be questioned, the conceptual model of groundwater flow at a site generally will receive careful scrutiny in reviews. Except in cases where particular conceptual models generally are accepted by regulators, selection of conceptual models for use in performance assessment must be defended on the basis of sitespecific data (NAS/NRC, 1990). Defense of conceptual models forces the need to identify those aspects of the behavior of a site and engineered features that are important to overall performance (i.e., an analyst must demonstrate an understanding of how data and assumptions that form the basis for conceptual models influence a decision about compliance with performance objectives). In the end, this understanding will form the basis for an integration and interpretation of results in a demonstration of compliance (Dodge et al., 1991). This objective should be considered during development of conceptual models, while recognizing that obtaining agreement on conceptual models can be the most challenging part of the performance assessment process. Simplifying assumptions will be necessary in developing conceptual models of all aspects of the performance of a disposal system. Those assumptions will be an important focus of reviews. An initial conceptual model may be based on limited data and designed to be very simple and conservative, in order to identify areas that require more detailed consideration. Further development of conceptual models should reflect an increased focus on radionuclides and processes of greatest concern. At all stages of the process, simplifying assumptions should be clearly identified and justified based on the current level of understanding of the system. The complexity of a conceptual model generally should be commensurate with the amount and quality of available data for a site (NAS/NRC, 1990). For example, it may not be appropriate to use a fully three-dimensional groundwater flow model when only limited site-specific data are available. Conceptual models also should anticipate the availability and defensibility of mathematical models for processes of concern. Efforts at data collection and development

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of conceptual models can be wasted if defensible mathematical models are not available to represent the level of detail provided. For example, flow of water in unsaturated, fractured rock is not well understood from the perspective of physical and mathematical modeling, so collection of large amounts of data on fracture size and distribution may not contribute to solving the problem. Use of a model also may be limited by the availability of data. For example, data on the behavior of concrete over time periods of concern to performance assessment (hundreds of years or more) are not available, so approximate representations of the long-term performance of concrete barriers must be used. Modeling of engineered systems generally is limited by a lack of long-term data in specific environments of concern. A natural geologic system is not as susceptible to this concern, because many geologic formations are quite stable over long time periods. Some needs for additional data may be addressed by evaluations of existing literature or by new experiments, including studies of natural analogs. Given the intended regulatory purpose of performance assessment, the complexity of modeling included in an assessment may be less than in models appropriate for other applications. In many cases, detailed models may need to be abstracted to a simpler level that is more readily defensible in a regulatory setting. Consequently, there may be a hierarchy of modeling approaches applied to a disposal system, with more complex models providing a conceptual basis for justification of simpler abstracted models. This approach is depicted in Figure 4.2 (DOE, 2001), which shows the performance assessment models (denoted by TSPA23 in the figure) as resting on the support of a variety of other models and sources of information. 4.1.3

Selection and Implementation of Mathematical Models

Following development of conceptual models, mathematical models to represent the concepts must be selected and implemented. There are three types of mathematical models, referred to here as generic, compliance and research models, that can be used in performance assessment. Generic models are relatively simple and are used primarily in establishing regulations and in studies intended to give an appreciation for the scope of the problem. Compliance models are site-specific and are developed for the purpose 23TSPA denotes the Total System Performance Assessment for the Yucca Mountain Facility for disposal of spent nuclear fuel and high-level waste in Nevada.

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Fig. 4.2. Hierarchy of modeling approaches in performance assessment of waste disposal systems (DOE, 2001). Performance assessment models, denoted by TSPA (Total System Performance Assessment), are shown to be supported by, and abstracted from, more detailed and rigorous modeling and site information.

of demonstrating compliance with applicable regulatory requirements. Research models are highly complex representations of specific processes that rely on detailed data. Each type of mathematical model is distinctly different, and is typically developed and used in different circumstances and for different reasons. Each type has specific data requirements that may further define or limit its applications. Models used in establishing regulations often are intended to represent processes that may bound many technologies, conditions, and environmental settings, and generic data used in those models also may be intentionally conservative. Use of conservative models in establishing regulations is appropriate, if kept within reason, to help ensure that an adequate margin of safety will be provided by different types of disposal facilities in different environmental settings. Many important aspects of compliance models must rely on site-specific data and models, to assure that results are tailored to fit a particular disposal facility of concern. New data may need to be collected, even though data collection is potentially timeconsuming and costly. Because compliance models and data

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are site-specific, they should not be used to set regulations that apply to a wide variety of sites, facility designs, and environmental conditions. Research models, and their data requirements, can be quite complex, and they often divide a conceptual model into numerous components or processes. Such models often require data that are not generally available and may not be practically obtainable at most sites. Given their limited scope, research models are seldom used to set regulations or evaluate compliance with regulatory requirements. However, they can provide useful support for a less detailed compliance evaluation (i.e., an increased understanding of a particular aspect of the performance of a disposal facility that may be obtained by use of research models can provide the basis for development of simpler yet defensible models for use in performance assessment). In such cases, the need for a research model may be justified by the scrutiny applied to a particular aspect of the performance assessment. The concern with migration of tritium gas at the Ward Valley Site in California is an example of this need. Mathematical models used in performance assessment may range in complexity from simple multiplicative-chain models, in which a quantity of interest is obtained as the product of several input parameters, to detailed process-level models that are expressed, for example, as second-order partial differential equations for which analytical (closed-form) solutions do not exist and numerical methods, such as finite-element or finite-difference, are required to obtain model outputs. Mathematical models that require numerical methods to obtain solutions are commonly used in compliance and research models. Mathematical models of any complexity normally are implemented in performance assessment by means of computer codes. 4.2 Process for Conducting Performance Assessments The performance assessment process can be a useful aid to decision making for many aspects of the development of a low-level waste disposal facility. The framework discussed in this Section is intended to provide a logical process that aims to maximize the benefits of conducting a performance assessment while minimizing efforts on activities that will have little or no benefit in supporting a licensing decision. The general approach is to begin with conservative assumptions, as appropriate, and to collect more data and refine models to reduce the degree of conservatism for those aspects of a system of importance in demonstrating adequate performance. The fewer assumptions that must be defended, the more likely it is that an analysis will stand up to scrutiny.

96 / 4. FRAMEWORK FOR PERFORMANCE ASSESSMENT The framework for performance assessment described in this Section accommodates the modular nature of many performance assessment models. Approaches to modeling for the different modules often used in performance assessments are discussed in Section 5. 4.2.1

Historical Perspective

Currently accepted approaches to performance assessment were developed starting from traditional engineering analyses of physical systems, including analyses of uncertainty using rigorous statistical methods. However, experience has shown that judgment regarding an appropriate balance between conservatism and realism in an analysis, given available data, is the crux of performance assessment, rather than an intent to provide a rigorous representation of the performance of a disposal system and its uncertainty (Section 2.4). The presence of uncertainty must be made known to decision makers, but limitations in models and data will limit an ability to use rigorous statistical techniques of uncertainty analysis in place of sound judgment and peer review. Engineering analyses based on fundamental principles of physics and chemistry have been used to supplement physical testing of structures and systems for thousands of years. For example, mathematical analyses have been used as a tool to design structures, such as bridges and buildings, at differing levels of sophistication at least since the time of the ancient Egyptians. This experience has been refined in recent years to such an extent that analyses often are conducted in lieu of experiments during engineering design. The value of such analyses has been demonstrated by the shorter times from concept to production and lower costs of conducting experimental testing in the presence of increasingly complex designs, with no loss in safety. Similar analyses have been used to assess the safety of nuclear reactors. Reactor designs are analyzed largely on the basis of theoretical models and empirical understanding of reactor behavior that has been gained from years of operating experience and experimental testing. Normal operational safety relies on proven designs, well established empirical and theoretical models, quality assurance, and testing. The most challenging analyses related to reactor operations are those concerned with potential accident (abnormal) conditions. However, even for analyses of abnormal conditions, many parts of a reactor system can be studied experimentally. Individual reactor systems include standard components (e.g., pumps, valves, pipes and wiring) that are well supported by

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performance data, including mechanical tolerances and reliability statistics, even under experimentally simulated accident conditions. Safety analyses can, therefore, draw on substantial data, in the form of statistical information, that can be used to quantify effects of failures of different components on the overall performance of a system. Rigorous statistical techniques have been developed for this purpose, such as fault tree analysis and probabilistic risk assessment. Since the early 1980s, considerable effort has been devoted to development of methodologies and models to assess the long-term performance of waste disposal systems. At first glance, it seems that statistical techniques developed for reactors or other engineered structures could be applied to waste disposal systems. However, while reactors or other engineered structures typically are well-controlled and -characterized systems with regular maintenance to keep components within a specified range of conditions, a disposal system represents a substantially different situation. Not only does the system have to perform for substantially longer periods of time without maintenance, but there also is little ability to observe the system. Furthermore, a disposal facility is placed in a heterogeneous, dynamic environmental setting. These complications force the need for an approach to performance assessment of disposal systems that differs from reactor analyses. The following paragraphs provide a perspective on the development of current approaches to performance assessment. It has long been recognized that performance assessment is limited by the availability of data (Case and Otis, 1988; DOE, 1982; Seitz et al., 1992b). Thus, data collection has always been an important part of the licensing process. An early perception in the low-level waste arena was that the licensing process would be conducted serially, starting with site characterization followed by a single assessment and defense of the results (Starmer et al., 1988). However, such an approach did not prove successful in practice, and the relationship between data collection and modeling has evolved into an iterative process discussed in this Report. In the mid-1980s, DOE became increasingly concerned with the cost and schedule of site characterization activities for the high-level waste repository program (DOE, 1985). To address this concern, efforts began on a process to link data-collection activities to licensing requirements. Exercises were conducted in which all project managers responsible for data collection had to identify the end user of data or a specific licensing requirement that mandated the need for data (i.e., a data-collection activity had to be justified by the end use of data). While a majority of data-collection

98 / 4. FRAMEWORK FOR PERFORMANCE ASSESSMENT activities were needed to support performance assessment, a large number of activities were identified for which there was no end user. This result suggested that the performance assessment process can be used to improve the cost-effectiveness of data-collection activities. In effect, results of preliminary performance assessments were used to justify data needs. This parallel and iterative approach is reflected in site characterization plans that were produced (DOE, 1986a; 1986b; 1986c; 1986d). A parallel activity in the high-level waste program was a process called “performance allocation.” This process attempted to quantify the importance of different activities to a performance assessment. This activity highlighted the value of sensitivity analyses of models used in performance assessment and provided valuable insight into limitations of numerical confidence intervals in describing uncertainties in models of natural systems. Initially, the intent was to categorize data needs on the basis of a statistical confidence required for a defensible result. The process involved an identification of data needs for each analysis activity, followed by discussions to address the relative importance of different parameters. As DOE’s exercise in performance allocation progressed, it became clear that general statements could be made about the relative importance of parameters, but it was difficult to justify results without supporting analyses. Furthermore, even in cases where substantial analyses had been conducted, it often became apparent that large amounts of data would be required to assign a statistically significant numerical confidence level to the data. It was eventually recognized that efforts to assign numerical confidence levels were not productive, and the focus shifted to developing relative confidence levels (low, medium and high) without any indication of numerical equivalents. Given that data are the cornerstone of performance assessment, the difficulty in assigning numerical confidence to data leads to problems in defending confidence limits in results that may be obtained from an analysis. This is not the case for most engineered systems, because data are already available or may be obtained relatively easily through testing. Performance assessments of low-level waste disposal facilities are faced with many of the same concerns as occurred in the high-level waste program. An important difference is that a substantial number of performance assessments of low-level waste disposal facilities have been conducted and submitted for peer review and licensing. Some experiences with these assessments are summarized in the following paragraphs.

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Reviews conducted by DOE’s Performance Assessment Peer Review Panel (DOE, 1988a) highlighted the importance of conceptual models and supporting data (RAEC, 1994). The majority of the panel’s comments addressed a lack of justification for an assumption used in a conceptual model or a particular parameter. An analysis of the importance of different assumptions and parameters has proved to be effective in providing information to address potential concerns and to justify assumptions and parameters in response to a comment. Peer reviews also have highlighted the benefit of using defensibly conservative models, rather than detailed processspecific models. The time and effort required to defend a simpler, more transparent model and commensurate data are much less than the time required to defend a complex model with its detailed data requirements. Experience with performance assessments of low-level waste disposal facilities at DOE sites has reinforced the need for an iterative approach in two ways. By requiring formal peer reviews early in the performance assessment process, the process became iterative simply because multiple reviews were required before approval of a facility could be granted. More importantly, early reviews were beneficial because they occurred at a time when it was necessary to use a less sophisticated modeling approach to obtain preliminary results. Results of a preliminary analysis were then reviewed, and feedback was generated by consensus on those areas that required additional effort prior to further review. Given the preliminary nature of initial results, those reviews focused on assumptions used in an analysis. In this manner, it became clear that an iterative approach could identify activities that would contribute to the defensibility of an analysis and activities that would have minimal impact. The experience of the Peer Review Panel also emphasized the need for judgment and consensus in evaluating the degree of conservatism in models, rather than simply relying on a structured analysis to provide an answer. Experience in conducting performance assessments of non-DOE low-level waste disposal facilities has resulted in many of the same insights as those obtained from assessments at DOE sites. NRC conducted performance assessments for early disposal sites and found that an iterative approach beginning with relatively simple models and progressing to more sophisticated models was a practical approach to the problem (Bergeron et al., 1991a; 1991b). In general, reviews of performance assessments at non-DOE sites have focused on justifications of data or conceptual models of various aspects of the performance of a disposal facility. In some cases, however, models themselves were deemed inadequate by reviewers.

100 / 4. FRAMEWORK FOR PERFORMANCE ASSESSMENT When this occurred, a model was based on generic data and was unsupported by data from the site of concern. This experience again highlights the need for site-specific data to support models. Another important lesson from past experience is that performance assessment requires a multi-disciplinary team of analysts (Case and Otis, 1988; Seitz et al., 1992a). The many different aspects that are involved in translating a given amount of waste in a particular disposal facility to a projected dose over long time frames makes it difficult for a few individuals to adequately address all aspects. Models in each area of a performance assessment usually must be simplified while retaining sufficient accuracy for the purpose of an assessment. The choice of data and simplifications in models in each area must be justified by someone who is knowledgeable in that field and has a clear understanding of the goals and peculiarities of performance assessment. Since simplifications in one area may impact other aspects of an assessment in a nonconservative manner, justification for simplifications requires close interaction with everyone involved in an assessment and a clear understanding of the unusual nature of performance assessment. Insights gained from past experience have illustrated that performance assessment is more than just a set of calculations that are carried out at the end of the site characterization process. Instead, performance assessment is an integrated process of developing a sufficient basis to support a licensing decision. This basis is developed through an interaction among analyses, design activities, and data collection. These activities need to be integrated in an appropriate manner that optimizes confidence in a licensing decision with respect to cost. Insights also have highlighted difficulties in defending standard engineering analyses using rigorous statistical techniques in the presence of limited data. Numerous rigorous methods have been considered and applied, but when assumptions, input data, and results must be defended, judgment, peer review, and exploration of multiple lines of reasoning have proven to be the primary factors in achieving a defensible assessment. 4.2.2

General Process for Conduct of Performance Assessment

The performance assessment process is complex and involves interactions among a variety of disciplines. Thus, a significant challenge in the process is to provide effective oversight of all assessment activities and to integrate results into a defensible package. Effective oversight involves producing a defensible product while

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allocating resources to those areas that provide the most benefit. These considerations form the basis of the performance assessment framework recommended in this Report. A general framework for the conduct of performance assessment is shown in Figure 4.3 (IAEA, 2004b). This figure focuses the more conceptual representation of an iterative approach to performance assessment in Figure 2.1 and the logical ordering of different steps in formulating and implementing calculational models used in performance assessment in Figure 4.1, which emphasizes the importance of subjective judgment in each step on the practical problem of using performance assessment in the licensing process. The process shown here is equivalent in all respects to those developed independently by other investigators (e.g., Case and Otis, 1988; DOE, 1985; IAEA, 1999; Kozak, 1994a; NRC, 2000; Seitz et al., 1992a). The general framework involves a sequence of steps in an iterative process. An iterative process is intended to provide an integration of data collection, modeling, calculations, and interpretation of results into a coherent framework that focuses on a licensing decision. Those activities are organized as a sequence that involves a description of the context for an assessment, a description of the disposal system, development and justification of scenarios for the future evolution of the system and its impacts on the public and the environment to be evaluated in a performance assessment, development and justification of conceptual and mathematical or physical models to be used in an assessment, and conduct of calculations, followed by a sequence involving interpretation of results, rendering of a licensing decision, and consideration of modifications if the licensing decision is unfavorable. The process includes a loop to represent multiple iterations, which can be applied to any step in the process. 4.2.2.1 Description of Context for Performance Assessment. The initial step of describing the context for a performance assessment is important because it defines the scope and content of an analysis. Regulatory requirements and decision goals, time frames of concern, aspects of an assessment that may be standardized (e.g., dose coefficients), the bias of the analysis (e.g., an intention to represent the disposal system conservatively), and requirements on quality assurance and documentation are among the important issues addressed in this step of the process. Ideally, the regulatory context for conduct of performance assessments, including standard assumptions that can be used in an assessment and requirements on quality assurance and

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Fig. 4.3. Performance assessment process involving sequential steps in an iterative process leading to a final licensing decision (IAEA, 2004b).

documentation, will be specified in advance of activities at a site and will not change prior to a final licensing decision. However, this may not always be the case. For example, low-level waste disposal facilities at many DOE sites have been developed and evaluated in accordance with requirements in Order 5820.2A (DOE, 1988a), but interpretations of performance objectives and requirements for conduct of performance assessments evolved over several years after the Order was issued (DOE, 1996c; Wood et al., 1994). Furthermore, Order 435.1 (DOE, 1999a; 1999b), which superseded Order 5820.2A, applies retroactively to these sites, and performance objectives and requirements on performance assessment are sufficiently

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different that revisions of existing performance assessments may be required. Similarly, interpretation by NRC staff of performance objectives in 10 CFR Part 61 (NRC, 1982b; 1993) and staff recommendations on approaches to performance assessment have evolved over time (NRC, 2000). 4.2.2.2 Description of Disposal System. The second step in the process consists of providing a complete description of a disposal facility of concern. Information provided in this step should include characteristics of the disposal site (e.g., geography and demography, meteorology and climatology, geology, seismology and volcanology, hydrology, geochemistry, and natural resources), the principal design features of the facility (i.e., disposal units and waste packages), and characteristics of waste including radiological and nonradiological properties (e.g., organic materials, free liquids). A description of the disposal facility should include available data and their interpretations that are used in subsequent evaluations of the facility, especially in developing and implementing conceptual models. In the first iteration of the performance assessment process, a description of a disposal facility normally would be based on information that is available before site characterization directed specifically at improving a performance assessment is undertaken and before detailed engineering designs have been developed. Nonetheless, some site-specific data and preliminary information on facility design should be in hand. The data may be sparse, but they should be sufficient to permit development of a simplified modeling approach, including use of conservative screening models. In subsequent iterations, a description of the disposal system, especially site characteristics and design features of the facility, would reflect new information developed for the purpose of improving the analysis. 4.2.2.3 Development and Justification of Scenarios. The third step in the process involves the development and justification of scenarios to be evaluated in a performance assessment. A scenario essentially is a set of assumptions about processes and events that result in releases of radionuclides from a disposal facility, transport in the environment to receptor locations beyond the facility boundary, and exposures of humans. Development of scenarios essentially involves developing one or more conceptual models of the evolution of a disposal system and future human exposures. Scenarios would be based on information used in preparing the description of a disposal facility, including available data and their interpretations.

104 / 4. FRAMEWORK FOR PERFORMANCE ASSESSMENT In the early 1980s, IAEA published a list of phenomena that are potentially relevant to release scenarios for waste repositories (IAEA, 1982; 1983). This list was presented as a “suggested checklist of phenomena” and has been cited as the starting point for scenario development activities in a number of repository safety studies. Also during that time, Sandia National Laboratories was developing its scenario selection methodology to support regulatory activities by NRC (Cranwell et al., 1990). Those initial studies were followed by a long history of development of methods of generating and analyzing scenarios, which is summarized by IAEA (2004b). The IAEA (2004b) report also provides, for the first time, a comprehensive set of features, events and processes that are intended for application to near-surface waste disposal facilities. Applications of the features, events and processes list to specific test-case problems, which were intended to test the comprehensiveness of that list, are presented by IAEA (2004c). With the publication of two IAEA (2004b; 2004c) reports, the step of generating scenarios in Figure 4.3 progressed from an essentially ad hoc approach, which had been used for many years in performance assessments of near-surface waste disposal facilities, to a more formal, justifiable, and traceable approach. 4.2.2.4 Formulation and Implementation of Models. The next step in the process involves the formulation and implementation of analytical models to describe the long-term performance of a disposal facility and its projected impacts on humans and the environment. As indicated in Figure 4.1, formulation of analytical models begins with development of conceptual models of the behavior of a disposal system. In essence, this step takes conceptual models to a more detailed and practical level. For example, a conceptual model of the hydrologic system at a disposal site and data that provide the basis for the conceptual model are used to develop a mathematical model to describe flow of water. Conceptual models can serve two functions. The first is to support development of simple, conservative screening models for the purpose of eliminating unimportant radionuclides and release and exposure pathways from further consideration in an assessment (Seitz and Kocher, 1993). Screening also can be used to eliminate unimportant features, events and processes. Screening models normally are developed only in the first iteration of the performance assessment process, unless the models are not accepted by reviewers and revisions are required. The essence of a screening model is that it must be demonstrably conservative. For example, a screening model for potential

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releases of radionuclides in water might assume that the expected inventory of a radionuclide would be dissolved in pore water in a facility. If the dose that would result from direct ingestion of pore water, with no further dilution, were below applicable performance objectives, that radionuclide could be eliminated from further concern in an assessment. As an example of screening applied to processes, conservative assumptions were used to argue that biotic intrusion into disposed waste is an unimportant release mechanism at a disposal site in Oak Ridge, Tennessee (ORNL, 1997a). Screening can be conducted by taking into account properties of a site, such as groundwater travel time and conservative retardation factors, and simple hand or spreadsheet calculations can be used to minimize the effort. The second, and more important, function of conceptual models is that they provide the basis for development of analytical models to be used in an assessment. Thus, conceptual models are at the heart of performance assessment. Given the importance of conceptual models, it is important that they be transparent and fully justified in each iteration of a performance assessment. The importance of conceptual models and their justification is evidenced by the intense scrutiny that they normally receive in reviews by technical peers and regulators. Justification of analytical models used in an assessment also is required. The level of detail in a model should be justified with respect to the purpose of an assessment. For example, simple equilibrium models of the behavior of radionuclides in terrestrial and aquatic food chains normally are sufficient but should be justified nonetheless. The level of detail in models of complex processes, such as groundwater flow, should be justified in relation to the quantity and quality of available data to support a model. Implementation of models by means of computer codes should be shown to be free of significant calculational error.

4.2.2.5 Conduct of Calculations (Consequence Analysis). In the next step in the process, calculations of projected consequences of waste disposal at a site are carried out. Outputs of those calculations must be commensurate with the assessment context defined in the first step. Thus, the result of this step is a set of numerical values (e.g., projected annual doses to off-site individuals) that can be compared with performance objectives specified in regulations. Provided proper attention has been paid to the previous steps, the process of carrying out calculations should be straightforward.

106 / 4. FRAMEWORK FOR PERFORMANCE ASSESSMENT 4.2.2.6 Interpretation of Results. Owing to the nature of performance assessment, especially the importance of judgment in most steps in the process and, thus, an inability to provide a rigorous quantification of uncertainty in projected outcomes, a simple and objective comparison of results of performance assessment calculations with applicable performance objectives generally is insufficient to provide reasonable assurance of regulatory compliance. There remains the additional, crucial step of providing an interpretation of results. Interpretation of results essentially involves a discussion of key assumptions (i.e., conceptual models, data sets, and analytical models) that led to projected outcomes. Results of sensitivity and uncertainty analyses and demonstrations of conservative bias in modeling are particularly important aspects of a proper interpretation of results with respect to regulatory compliance. If a performance assessment indicates compliance with regulatory criteria, and if models and data used in an assessment are deemed sufficient to support a licensing decision, then the conclusion of this step should be that the case for reasonable assurance of compliance with performance objectives has been made. Interpretation of results also should discuss implications of a performance assessment for site characterization, the design of a disposal facility and any waste packages, development of a closure plan, acceptable physical and chemical characteristics of waste, limits on disposal of different radionuclides, monitoring of the facility, and any other aspects of the disposal system of importance to a licensing decision. That is, results of an assessment will be based on assumptions about a disposal system that then must be reflected in the design and operation of the facility and characteristics of the site. Thus, interpretation of results is the step in the performance assessment process where data collection, facility design, and modeling activities in the other steps are integrated into a coherent whole that demonstrates a sufficient understanding of a disposal facility for the purpose of supporting a licensing decision. If a performance assessment does not indicate that reasonable assurance of compliance with regulatory requirements is obtained, then another iteration of the assessment process normally would be warranted. In such cases, it is important that an interpretation of results include an identification of reasons why the case for compliance has not been made. Important factors affecting a licensing decision may include inadequate data, insufficient support of models used in an assessment, and omission of some scenarios or models without sufficient justification. Even if all aspects of an

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analysis are judged to be sufficient, estimated inventories or concentrations of specific radionuclides in wastes intended for disposal may be too high. 4.2.2.7 Modifications of Assessment. If a regulatory authority judges that a performance assessment does not provide reasonable assurance of compliance of a disposal facility with regulatory requirements, then modifications of the assessment are needed if the facility is to be licensed. There are essentially three options for modifying an assessment: (1) development of alternative conceptual models, which may require additional data; (2) modification of engineered features (e.g., waste containers, backfill, engineered structures, cover) to reduce projected exposures of humans; or (3) collection of additional data to reduce the degree of conservatism in assumptions used in the assessment. A fourth option, which may not require changes in the performance assessment itself, is to reduce allowable inventories of problem radionuclides. The approach that should be taken in modifying a performance assessment depends on the reasons why the case for compliance was not made, as identified in an interpretation of results. If data used in an analysis are judged inadequate, additional data would need to be collected. Insufficient support of models might require development of new models or a better justification for models used in an analysis. Concerns about omission of potentially important scenarios or models could be addressed by providing proper justification, but evaluation of additional scenarios and use of additional models in an analysis may be required. An important consideration in evaluating options for modifying an assessment is whether they would be cost-effective. If none of the needed modifications identified in an interpretation of results is cost-effective, limits would need to be placed on disposal of problem radionuclides identified in the assessment, or the site would be rejected. Cost-effectiveness is a particularly important consideration if collection of additional data at the site is an identified need. It is possible that obtaining required data would be so expensive that a better solution would be to reject the disposal concept and select a different site or facility design. It should be recognized that, in some circumstances, a site or facility design may not be acceptable and, therefore, further analysis would be useless. For any accepted disposal concept at a suitable site, reasonable assurance of compliance with regulatory requirements presumably could be obtained by placing restrictions on inventories or concentrations of radionuclides for disposal, provided the performance assessment itself were judged adequate for purposes of a licensing

108 / 4. FRAMEWORK FOR PERFORMANCE ASSESSMENT decision. This solution would obviate the need for an additional iteration of the performance assessment process. However, this would not be a practical solution if disposal limits were so restrictive that substantial amounts of waste intended for disposal in a facility would be disqualified. The foregoing discussions in this Section mainly address modifications of a performance assessment at the stage of licensing a disposal facility. During facility operations and the period of institutional control after facility closure, monitoring data or new scientific developments may provide information that casts doubt on an existing performance assessment or that could be used to refine an analysis (e.g., to eliminate conservative or erroneous assumptions). Such occurrences could be addressed as part of an ongoing process of maintaining a performance assessment that continues throughout the postclosure period of institutional control. 4.2.2.8 Iterations of Performance Assessment. One or more iterations of an initial performance assessment normally will be needed in the process of developing a disposal facility and obtaining approval by regulators. An initial assessment may be in the form of a screening model, or it may be a simplistic representation of a disposal system, based on limited site-specific data, that is developed for the purpose of identifying needs for additional data or conditions on facility design or operations. Such assessments should not be expected to be sufficient for purposes of licensing. Even for well-developed disposal systems, it is unlikely that the first effort at a performance assessment will pass muster with regulators. The diagram in Figure 4.3 indicates that each iteration of a performance assessment may involve any or all of the steps of an assessment. Requirements that a disposal facility must meet or requirements for conduct of performance assessment might change during the licensing process. Significant changes in waste that is intended for disposal, such as changes in waste forms or important radionuclides, generally should trigger a revision of an assessment. The description of a disposal facility would need to be revised if there are significant changes in facility design or in site-specific data that were used to develop conceptual models of the disposal system. Scenarios (conceptual models) would need to be revised on the basis of changes in a disposal concept or new site-specific data, and new scenarios may require evaluation. Changes in scenarios might require new analytical models. Finally, each iteration would require a new set of calculations and proper interpretation of results.

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The need for modification of each step will depend on inadequacies of the previous assessment. Each iteration may be increasingly complex and require increasing amounts of data. If, at the end of an iteration, regulators agree that an assessment provides reasonable assurance of compliance with performance objectives, the process is complete. 4.2.2.9 Summary. This Section has laid out a general process for conduct of performance assessment. The crux of the process is that it provides for integration and interpretation of data collection, facility design, and modeling activities in a way that supports a licensing decision by focusing on those aspects of a disposal system that are important to long-term performance and by requiring justification of all assumptions and documentation of all aspects of an assessment. The general steps in the performance assessment process described in this Section should be viewed as guidelines. The process is intended to be flexible, so that it can accommodate characteristics of particular sites, facility designs, and wastes. The framework discussed here is intended to highlight the logical flow of an analysis and to emphasize aspects of the process that are expected to be important at any site; it is not intended to lay out specific requirements that would apply to all assessments. A general consideration throughout the performance assessment process is the need to include regular reviews to maintain communication between staff working on different parts of an analysis.

5. Performance Assessment Models The purpose of this Section is to provide guidance on the types of models that should be acceptable for use in performance assessments of near-surface disposal facilities for low-level radioactive waste. Interesting features of models that have been brought to light by their use in performance assessment also are discussed. Discussions in this Section are concerned with selection of models to represent physical and chemical processes, rather than selection of computer codes to implement those models. Specific codes that may be used in performance assessment are intentionally not mentioned. Experience has shown that the defensibility of an assessment rests on proper justification of models and their application at specific disposal sites. If models are shown to provide a reasonable representation of a disposal system for the purpose of performance assessment, then reviewers will be more likely to accept technical judgments embodied in an assessment. Rarely do reviews of an assessment focus on specific features of computer codes. Consequently, if models and their application are properly justified, selection and justification of computer codes to implement the models should be straightforward. Discussions in this Section focus mainly on mathematical models. This focus is not meant to imply that physical models are not important in performance assessment. Indeed, studies of natural or man-made physical systems have been important in developing and parameterizing mathematical models of several aspects of performance assessment including, for example, the performance of cover systems, degradation of concrete, release of radionuclides from waste forms by infiltrating water, flow of water in natural geologic environments, and transport of radionuclides in the environment and through various food chains. Such studies also can provide data for use in efforts to validate predictions of mathematical models. 5.1 General Approach to Modeling of Disposal Systems The behavior of a radioactive waste disposal system generally is complex, even when the system only needs to be understood with 110

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respect to issues of regulatory compliance. Some of the physical processes of concern to long-term performance are indicated in Figure 5.1 for an intact disposal facility and in Figure 5.2 for a facility that has undergone significant degradation (Kozak and Olague, 1995). In reality, the hydrologic system has few natural boundaries, but this system generally is simplified considerably in performance

Fig. 5.1. Conceptual diagram of flow processes in and around an intact near-surface disposal facility (Kozak and Olague, 1995).

Fig. 5.2. Conceptual diagram of flow processes in and around a degraded near-surface disposal facility (Kozak and Olague, 1995).

112 / 5. PERFORMANCE ASSESSMENT MODELS assessment. A disposal facility is intimately linked with the local hydrologic system, and both of these are linked with the surrounding biological system, which also is complex in its own right, but these interactions generally are not considered in performance assessment. Furthermore, there is a continuum of intermediate physical conditions between a completely intact and completely degraded disposal facility and the evolution from one condition to the other occurs over extended time frames, but this evolutionary process usually is represented as one or more step changes at discrete times. Over long time periods, the biosphere and climate also can be expected to evolve, but those aspects of system performance generally are ignored or treated in a stylized manner. Even these figures and this description are significant simplifications of the behavior of a disposal system. Treating a complex disposal system as an integrated whole is beyond current scientific, mathematical and conceptual capabilities. As a result, simplified representations of disposal systems must be sought, while still retaining enough of the physical characteristics of a system to allow defensible licensing decisions to be made. Three types of simplifications generally are incorporated in modeling. First, environmental and ecological systems of concern are decoupled from a disposal system by describing transport and exposure pathways between disposed waste in a facility and a human receptor in such a way that information flows only in one direction. For example, the groundwater pathway is represented as a release into and transport via groundwater, with subsequent release into the biosphere, but possible influences of the biosphere on a groundwater flow system are treated in a simplified manner to eliminate consideration of fully coupled, nonlinear behavior. Second, physical phenomena of interest are divided into modules, which are treated as though they interact only in linear, unidirectional ways. That is, results from one module (e.g., projected releases of radionuclides from a disposal facility) provide input to the next (e.g., transport in the vadose zone), and these linkages are assumed to be independent of quantities of radionuclides in waste. Division of physical phenomena into modules is done in the interest of mathematical and conceptual convenience, rather than rigor. Third, as noted previously, evolution of a disposal system over time is evaluated in a highly stylized manner, which usually involves step changes at particular times rather than continuous changes over time. This treatment of time dependencies in system behavior is an unavoidable consequence of conceptual and computational limitations.

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Decoupling and Simplifying an Analysis

Performance assessment requires an integration of results of a number of different types of models to provide an overall description of a disposal system. Performance objectives for disposal systems summarized in Section 3.4.2 require calculations of annual doses to off-site members of the public over time that take into account all radionuclides and all release, transport and exposure pathways. However, the number of radionuclides in low-level waste usually is large, as is the number of conceivable pathways from disposed waste to human receptors, and there may be interactions between parts of a pathway. As a result, it is necessary to simplify an analysis in a justifiable manner. In general, the inherent complexity of an analysis is reduced by screening of possible pathways, interactions, processes and radionuclides, in order to select for analysis only those aspects of the problem that are expected to have a significant effect on results (e.g., projected doses). In addition, analyses generally are simplified by neglecting any nonlinear interactions between pathways or between portions of a pathway. A perspective on the possible scope of performance assessment is provided by a comprehensive summary of potential pathways between disposed waste and human receptors, including sequences of up to four intermediate environmental and biological media (Shipers, 1989). Possible pathways reasonably can be combined in more than 8,000 different ways (Shipers and Harlan, 1989). Even this number constitutes an initial screening, because a limitation to four media is arbitrary. That limitation was justified on the basis of arguments about dilution in transport from one medium to the next, but additional media may need to be considered in some circumstances. The large number of pathways then was screened to select a reasonable set on the basis of generic arguments about the behavior of a well-constructed facility (Shipers and Harlan, 1989), and this list was subsequently revised to account for engineered alternatives (Rao et al., 1992). However, those lists of potentially important pathways are generic and were intended for use in developing a generic performance assessment methodology. For specific sites and facility designs, a complete set of pathways should be considered initially and then screened on the basis of site-specific considerations. However, there is no single approach to conducting such a screening assessment that would apply at all sites. NRC (1991b) has indicated that exposure pathways that comprise <5 % of the total potential dose can be excluded from an analysis, and that simple screening calculations can be used to exclude pathways. However, as noted above, such an approach requires, in

114 / 5. PERFORMANCE ASSESSMENT MODELS principle, a screening calculation that considers more than 8,000 pathways and a complete set of potentially important radionuclides. A more common approach is to use judgment in a screening process. Arguments are made that pathways can be excluded from an analysis because they are unlikely to occur or they are expected to be minor contributors to dose if they do occur. Similarly, many radionuclides can be excluded on the basis of their half-lives or knowledge that their inventory in waste would be insignificant compared with inventories of other similar radionuclides included in an analysis. Use of a judgment-based approach to screening is supported by a substantial number of performance assessments which have shown that only a few pathways and radionuclides usually are major contributors to projected doses. However, site-specific analyses also provide ample evidence of unusual conditions that can be important. Consequently, screening of pathways and radionuclides on the basis of judgment, although it is a potentially satisfactory approach to reducing the number of pathways and radionuclides that need to be considered in a site-specific analysis, requires proper justification. 5.1.2

Analysis by Modules

Once the important pathways for transport of radionuclides from locations of disposed waste to a location of an assumed human receptor beyond the boundary of a disposal facility have been selected by means of a screening analysis, those pathways need to be analyzed to identify important radionuclides in waste. Performance assessments of low-level waste disposal facilities have been found to have a number of common components, and patterns have developed in the way that representations of physical components of a disposal system interact in an overall model of the system. A representation of these patterns is shown in Figure 5.3. The diagram in Figure 5.3 indicates that a performance assessment generally can be divided into modules that are conceptually or physically separate, and this division provides a convenient framework for discussion of an assessment. At most disposal sites, the most important medium for release and transport of radionuclides is water. As a result, the first step in an assessment usually is an analysis of flow rates of water through a disposal system. A flow analysis is used in radionuclide source-term (release) and transport analyses. Results of a transport analysis generally are in the form of concentrations or fluxes at locations where radionuclides are assumed to enter the biosphere. A biosphere analysis

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Fig. 5.3. Conceptual representation of principal components of performance assessment for waste disposal systems.

translates concentrations of radionuclides in the environment to estimates of external dose rates to a human receptor or rates of intake of radionuclides by various ingestion and inhalation pathways, which are combined with assumptions about living habits and external and internal dosimetry data to obtain estimates of annual dose. The basic modules in Figure 5.3 often are subdivided into components that represent models of specific physical systems, an example of which is shown in Figure 5.4. In this representation, in which releases of radionuclides from a disposal facility into groundwater are assumed to be the most important, a flow analysis is subdivided into analyses of infiltration, flow in the vadose zone, and flow in an aquifer. For the purpose of estimating release and transport of radionuclides, an analysis of concrete degradation describes the evolution of physical parameters used in flow modeling, thus providing input to infiltration and vadose-zone flow analyses. Once a flow analysis, including the influence of concrete degradation, is completed, the results form the basis for modeling of the source term and radionuclide transport. A source-term analysis is distinguished from the rest of a transport analysis in that it provides an estimate of releases of radionuclides into solution, which then is used as input to a transport analysis. An exposure pathway and dose analysis uses results of a transport analysis and translates these into an appropriate output (e.g., annual dose). The representation in Figure 5.4 would be different if releases of radionuclides to surface water or the atmosphere are of concern at a specific site. When a performance assessment is divided into modules, mathematical interfaces between modules must be defined. In some cases, logical or physical connections between modules are well established. For example, use of results of a groundwater transport analysis in the form of concentrations of radionuclides in water as input to an exposure pathway analysis represents a straightforward

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Fig. 5.4. Conceptual representation of components often used in performance assessments of waste disposal systems.

logical link between the two modules. Similarly, at the water table, which is the boundary between unsaturated and saturated flow regimes, a condition of zero pressure head is physically well established. On the other hand, an interface condition for the transport equation between unsaturated and saturated zones is not well defined. On physical grounds, the two zones should be analyzed as a coupled unit, in which case only a continuity condition at the interface need be applied. However, this approach frequently is difficult or even impossible, owing to the difficulty in modeling variably saturated flow fields, and subdividing a groundwater system into two modules is a pragmatic necessity. Other interfaces in Figure 5.4 for which natural interface conditions are difficult to define include: (1) interfaces between the source term and surrounding soil for both flow and transport, (2) the interface between climatology and hydrology (i.e., the soil surface at which infiltration occurs), and (3) the interface between radionuclide concentrations in groundwater and at a well head. Various interfaces of concern are discussed in appropriate parts of Section 5.

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Analysis of Time Dependence

Time dependencies of physical processes generally are represented in a stylized manner in performance assessment. A time dependence normally is accounted for in modeling properties of engineered materials that degrade with time and in modeling releases of radionuclides and transport in the environment. Time-dependent behaviors of those processes that normally are assumed in performance assessment are discussed in appropriate parts of Section 5. However, other aspects of an analysis, including the behavior of hydrologic systems and the biosphere, normally are represented by steady-state models, even though time-dependent behaviors undoubtedly will occur. 5.1.4

Organization of Section

The remainder of Section 5 is devoted to discussions of specific aspects of modeling for components (modules) of a performance assessment shown in Figure 5.4. Other models that may be needed at specific sites to analyze other transport pathways also are discussed. Components of performance assessment discussed in Section 5 include cover performance and infiltration (Section 5.2), the performance of concrete barriers (Section 5.3), the source term (Section 5.4), unsaturated zone flow and transport (Section 5.5), aquifer flow (Section 5.6), radionuclide transport in groundwater and surface water (Section 5.7), atmospheric transport (Section 5.8), biotic transport (Section 5.9), and exposure pathways and radiological impacts (Section 5.10). While reading various parts of Section 5, it is important to bear in mind that performance assessment must represent an entire disposal system, and that a model of each component of the system needs to be coupled with all other models, as appropriate, to achieve a satisfactory model of the entire system. Interface conditions between components should be scrutinized by analysts, as well as details of models of each component, because interfaces between models can be as important as the models themselves. An analysis needs to be checked to ensure that mass of contaminants is conserved (e.g., that the same mass of material is not assumed to be transported through multiple pathways). 5.2 Cover Performance and Infiltration 5.2.1

Introduction

The presence of water in a disposal facility is expected to be a primary cause of releases of radionuclides at most sites. Water

118 / 5. PERFORMANCE ASSESSMENT MODELS accelerates degradation of engineered structures and waste containers, it provides a medium into which radionuclides can dissolve, and its movement is a primary means by which radionuclides can be transported from a facility into the surrounding environment. Therefore, the purpose of most engineered barriers included in facility designs is to minimize contact between water and emplaced waste. As a result, modeling of infiltration into a disposal facility is a natural starting point for performance assessment. The purpose of an analysis of infiltration is to determine the flux of water (i.e., incident precipitation) that moves into the subsurface environment.24 Processes that must be represented in an analysis of infiltration are depicted in Figure 5.5. Precipitation incident on the ground surface is partitioned among four processes: evaporation, transpiration, surface runoff, and infiltration. Infiltration is

Fig. 5.5. Hydrologic processes affecting infiltration of water into waste disposal units (Smyth et al., 1990). 24

In some analyses, downward movement of water through a cover system in a disposal facility is referred to as “percolation,” and the term “infiltration” refers only to downward movement through native soil around the facility. In this Report, the term “infiltration” is used to describe both processes.

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the primary concern in performance assessment, because the other processes shown in Figure 5.5 serve to direct water away from waste. Meyer et al. (1996) identify other processes that could be considered, including snow accumulation and melt, thermal effects and vapor phase flow, subsurface lateral drainage, heterogeneity, hysteresis, and anisotropy of soil properties, and they subdivide transpiration into plant growth and plant uptake of water. However, infiltration is the process of primary concern in performance assessment, because it affects release rates of radionuclides (i.e., the source term) and flow in the surrounding geologic setting. Facility designs often attempt to maximize surface runoff to reduce the potential for infiltration into waste, and most closure plans include vegetation that is intended to maximize transpiration. Furthermore, most near-surface disposal facilities are developed with the intention of installing a low-permeability cover to further reduce the amount of water that reaches waste. Formal frameworks developed by Ho et al. (2002; 2004) or Meyer et al. (1996) may provide useful approaches to evaluating the net effect of engineered components of a cover system on infiltration into waste and the surrounding geologic environment. The framework developed by Meyer et al. (1996) is shown in Figure 5.6. It calls for differentiating between natural recharge, effects of engineered barriers, and effects of uncertainty on analyses. The approach recommended by Ho et al. (2002; 2004) is a scenariobased framework for modeling alternative events and processes that could affect cover performance. The two approaches to evaluating cover performance are viewed as complementary. When a disposal facility is constructed with engineered barriers, including an engineered cover, an analysis of infiltration must consider the natural hydrologic system and alterations of natural infiltration caused by an engineered system. These are separate problems that require different approaches. Since natural infiltration is a characteristic of an undisturbed setting, it can be measured or estimated using site-specific data. However, infiltration through a cover system can only be estimated by modeling prior to its construction. Unfortunately, the latter information is of greater importance than the former when engineered barriers are assumed to provide some benefit. Thus, modeling of infiltration in performance assessment involves, to a significant extent, development of a credible link between measurable undisturbed conditions and conditions that might apply when a disposal facility is in place. An often-used assumption is that a worst-case infiltration rate through a cover is equal to the natural infiltration at a site prior to

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Fig. 5.6. Conceptual approach to estimating infiltration into waste developed by Meyer et al. (1996).

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construction of a facility. This assumption must be evaluated for the specific facility design and hydrological conditions at a site. Subsidence of a cover system following degradation of an underlying disposal vault could lead to infiltration rates in excess of natural infiltration. Furthermore, it is necessary to take account of the episodic behavior of infiltration when estimating annual infiltration. Despite the episodic, transient nature of all these processes, almost all performance assessments assume steady-state flow behavior as abstracted from transient data. Care must be taken that this abstraction process leads to credible estimates of infiltration into a facility. As in other analyses conducted in performance assessment, a large body of literature that describes research on infiltration is available. Rather than exhaustively reviewing that literature, the purpose of this Section is to describe how the current understanding of infiltration mechanisms is translated into a format useful to performance assessment. From the perspective of modeling an entire disposal system in a performance assessment, an estimated infiltration rate into waste usually provides an input to an analysis of releases of radionuclides (i.e., the source term). An estimated infiltration rate at locations away from a disposal facility provides input to an analysis of flow in the vadose zone and in groundwater. 5.2.2

Types of Covers

Several types of cover designs that employ different kinds of barriers to infiltration have been considered for use in near-surface disposal facilities. The various covers can be categorized as resistive, conductive, or biological (Schulz et al., 1990; 1997). The third type uses engineered barriers and vegetation, and is sometimes referred to as a “bioengineered” cover. A resistive cover is the simplest. As shown in Figure 5.7 (top), a resistive cover consists of a layer of low-permeability material, which limits infiltration below the cover by virtue of its low hydraulic conductivity. This type of cover most frequently consists of a thick clay layer. A conductive cover is considerably more complex than a resistive cover. Conductive covers consist of multiple layers of materials of differing textures. This type of cover is designed to impede infiltration by using layers that provide low vertical permeability but high lateral hydraulic conductivity under both saturated and unsaturated conditions. In this way, water is channeled laterally around waste. In its simplest form, a cover system consists of a soil

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Fig. 5.7. Schematic representations of cover systems for near-surface waste disposal facilities (Schulz et al., 1990): resistive cover consisting of single clay layer above waste (top); conductive cover consisting of layer of relatively fine material providing high lateral conductivity above a layer of coarser material providing a capillary break (middle); biological (vegetative) cover used in conjunction with impermeable material (bottom).

layer above a capillary break, as shown in Figure 5.7 (middle). The capillary break consists of a layer of coarse material, such as gravel. Under unsaturated conditions, coarse materials conduct water poorly, in which case infiltrating water will tend to channel through the finer soil above the capillary break and be conducted around waste. More complicated conductive covers consisting of many layers of materials with contrasting properties have been proposed (CILLRWCC, 1997; NE, 1997; RAEC, 1988). Those covers incorporate engineering experience on constructability of layers of various thicknesses and overburden, an ability to compact soils at field scale, and use of filters to limit clogging of coarse material by fines that are transported into them. A considerable body of literature exists on construction of such covers (EPA, 1989b; Hauser et al., 2001). An example of a multilayer cover, which has a resistive

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layer (clay) barrier over a conductive layer barrier, is shown in Figure 5.8. This type of cover design represents an attempt to build in some redundancy, so that a failure of one part of the cover does not result in failure of the cover as a whole. Biological covers consist of vegetation that is intended to optimize transpiration of water (Schulz et al., 1990) or minimize the potential for biotic intrusion (Salvo and Cook, 1993). A test facility developed to illustrate the potential usefulness of a biological cover (Schulz et al., 1990) is shown in Figure 5.7 (bottom). A relatively impermeable engineered cover is used to enhance runoff of precipitation. By keeping the vegetation under a water-stressed condition, the plant root system efficiently scavenges infiltrating water. 5.2.3

Degradation of Covers

Over the long time periods considered in performance assessment, there is considerable uncertainty about the behavior of covers. Each type of cover can be expected to evolve differently over time, and to be degraded by different mechanisms. All covers may be subject to aeolian and fluvial erosion processes. The extent to which erosion plays a factor in performance assessment depends on site conditions and cover design. It is generally expected that aeolian erosion will be easier to characterize, evaluate and design against. This type of erosion is not considered further in this Report. Fluvial erosion, as a process, has been extensively studied (e.g., Flanagan et al., 2001), but specific approaches that are applicable to evaluating covers for near-surface disposal facilities are not easily specified, because conditions in covers differ from typical conditions in the broader erosion literature. Methodologies that are intended to evaluate cover performance from a broad scenario-based viewpoint (e.g., Ho et al., 2002; 2004) omit consideration of erosion in example applications. Related landfill covers at RCRA and uranium mill-tailings disposal sites have relatively short required design lifetimes (EPA, 1989b; Hauser et al., 2001). Consequently, uncertainties inevitably exist concerning the range of erosion rates for a particular cover design. Within these constraints, erosion typically is considered from a design perspective, with sufficient depth of a cover used to accommodate expected erosion during an assumed period of performance. Erosion rarely is included in performance assessment as an explicit pathway for transport of radionuclides from a disposal facility. Clay covers are subject to desiccation fracturing, which can occur under drought conditions but also in moist climates. This process can lead to large increases in hydraulic conductivity. Multilayer

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Fig. 5.8. Schematic representation of multilayer cover consisting of conductive and resistive materials (Meyer, 1993).

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conductive covers must retain their structural integrity in order to function as intended. Consequently, subsidence or other mechanical disturbances can have severe effects on functional capability (Schulz et al., 1990). Resistive and conductive covers will begin to degrade when deep roots disrupt their structure and, as noted previously, those covers also can degrade by clogging. Cover designs usually call for planting of shallow-rooted species (often grasses) over a cover. However, those species cannot be relied on to persist after the period of institutional control, since native plants can be expected to invade a site at that time. Over hundreds to thousands of years, it is reasonable to presume that deep-rooted plants will disrupt a cover to some extent. Biological covers also can be expected to experience a progression toward other plant species, with the result that such covers cannot be expected to behave as intended after the period of institutional control. Bioengineering of cover systems should be regarded as a suitable technology for remedial actions and maintenance of structures, but they normally will not be important in long-term performance assessments. 5.2.4

Approaches to Estimating Infiltration

There is no generally applicable method of estimating infiltration of water into natural soils or engineered cover systems (Balek, 1987; Gee and Hillel, 1988; Meyer and Taira, 2001; Meyer et al., 1996; Smyth et al., 1990). Water balance models described below have been found to be appropriate under moist conditions, but that approach can lead to large relative errors when applied at arid sites (Gee and Hillel, 1988). The report by Meyer et al. (1996) provides an excellent summary of advantages and disadvantages of various approaches. In water balance models, hydrologic processes that occur at the ground surface are divided into precipitation, evaporation, transpiration and runoff, which are estimated independently, with the remainder of the water budget applied to infiltration (Figure 5.5). Thus, infiltration at steady-state (I) is estimated as: δ( I = P – E – T – R , ) where: P E T R

= = = =

precipitation evaporation transpiration runoff

(5.1)

126 / 5. PERFORMANCE ASSESSMENT MODELS Forms of Equation 5.1 that represent transient behavior include a storage term to account for the storativity of soil (Meyer et al., 1996). Parameters in Equation 5.1 generally are averaged over some period of time, and it generally is found that the resulting estimate of infiltration depends on the averaging time. That is, use of yearly averages of precipitation, evaporation, transpiration and runoff to estimate infiltration tends to produce different results than use of monthly, daily or hourly averages. An approach to applying a dynamic model to obtain average infiltration rates has been developed and used by EPA (2000c). Except for precipitation, each component on the right-hand side of Equation 5.1 is difficult to measure, and each depends on many effects that are not easily quantified. Furthermore, each component of the water balance is highly variable, either spatially, temporally or both. Those variabilities generally are not taken into account in performance assessment. Instead, a spatially uniform and temporally constant infiltration rate usually is assumed. An assumed infiltration rate should be based on credible site-specific data. At sites with abundant precipitation, errors introduced by use of simplifying assumptions described above tend to be small (Gee and Hillel, 1988). That is, when infiltration is high, small errors in parameters in Equation 5.1 result in relatively small errors in estimated infiltration. However, when infiltration is low, as at arid sites, the error in an estimated infiltration can be comparable to or greater than the infiltration itself. As a result, other approaches have been proposed for use at arid sites. Of particular importance at arid sites are so-called isotope methods (Muller, 1982), which often involve measurements of concentrations in soil water of radionuclides that are produced in the atmosphere. For example, measured concentrations of 3H or 14C in soil water at different depths and knowledge of production rates in the atmosphere over time combined with the known half-life can be used to estimate an infiltration rate of water. Measurements of stable isotopes also can be used. For example, age dating of water using the stable decay product 3He is an extension of the method using 3H (Torgersen et al., 1979), and measurements of 4He can be used to estimate the length of time that water has been in contact with soil that contains naturally occurring, alpha-emitting radionuclides (Marine, 1979). Isotope methods require knowledge of historical production rates in the atmosphere and other sources in soil that could affect measured concentrations. For example, contact of water with carbonate materials can affect the ratio of 14C to total carbon in soil water (Fontes and Garnier, 1979). The particular

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isotope method that would be most useful depends on the infiltration rate of water. For example, age dating of very old groundwater may require measurements of long-lived radionuclides (e.g., 36Cl). Lysimeter testing also can be used to determine infiltration at arid sites. However, this option likely will be impractical, owing to its cost and the length of time over which testing would need to be conducted at sites with low infiltration rates. For the purpose of performance assessment, Kozak et al. (1993) recommended that more than one method should be used to evaluate infiltration at a site. As many methods as practicable should be applied in an attempt to develop an estimate of uncertainty associated with the approximation technique used in an analysis. The different approaches described above have been developed to determine natural infiltration at a site. An estimate of the infiltration rate through an engineered cover system and into a waste disposal unit also is needed. The quantity of interest in performance assessment is the infiltration rate through the bottom of a cover as a function of time. Infiltration through a cover system can be described using standard vadose-zone flow models. Many modeling approaches have been proposed to carry out an analysis (Meyer et al., 1996; Simmons and Meyer, 2000; Smyth et al., 1990). The utility of such approaches must be evaluated on a design-specific basis. Scanlon et al. (2001; 2002) have demonstrated the difficulty of obtaining consistent results using different detailed modeling approaches. The maximum infiltration rate through an intact resistive barrier frequently can be estimated on the basis of a one-dimensional, unit-gradient approach represented by: q = K sat , where: q = Ksat =

(5.2)

Darcy velocity through a cover effective (macroscopic) saturated hydraulic conductivity of a resistive layer

This approach assumes (1) steady-state flow through the cover, (2) negligible ponding of water over the cover, (3) negligible pressure-head gradient (capillarity) under the cover, and (4) lateral extension of the cover sufficiently far from the edge of the waste to prevent significant multidimensional flow effects. These assumptions usually must be violated severely before significant deviations from a unit-gradient flow rate are calculated. Thus, this approach usually is reasonable for purposes of performance assessment.

128 / 5. PERFORMANCE ASSESSMENT MODELS Once the conductivity of a resistive cover is assumed to be altered from its initial condition, use of Equation 5.2 often is still appropriate, but the saturated conductivity will change over time. A common assumption in performance assessment is that a cover functions as designed for a relatively short period of time and then fails completely. The infiltration rate into an underlying disposal unit therefore is assumed to equal Ksat for a specified period of time, after which it is assumed to undergo a step change to the natural infiltration rate. An assumed period of full functioning of a cover must be justified in performance assessment, and the final infiltration rate must be assessed because, depending on the facility design and mode of degradation of a cover, the final infiltration rate can exceed the natural infiltration rate. Modeling of intact conductive covers tends to be a more formidable task than modeling of resistive covers. Flow through conductive covers is intrinsically multidimensional, and soil properties of adjoining layers intentionally are quite different. Furthermore, conductive covers frequently are designed such that the theoretical flow rate through a cover is effectively zero. It is unlikely that such ideal projections can be achieved in the field, although remains found in some ancient tombs in Japan (Watanabe, 1989) suggest that good functioning of a cover over many thousands of years may be possible. Flow through conductive covers is most commonly modeled using the Richards equation of vadose-zone flow (Section 5.5.5). Scanlon et al. (2002) concluded that methods founded on that equation were superior to other approaches. However, the model requires considerable information about characteristic properties of soil that are difficult to measure, are highly variable, and are sensitive to techniques used to construct a cover. As a result, this type of flow analysis relies heavily on data that will, for the most part, be unavailable or unreliable. As in the case of resistive covers, failure of conductive covers usually is assumed to occur instantaneously and completely. As a result, the infiltration rate after cover failure is commonly assumed to be the same as the natural infiltration rate on the basis of an additional assumption that an underlying vault or trench experiences only minimal subsidence. If massive subsidence could occur (e.g., due to collapse of a vault roof), infiltrating water may be funneled into a disposal unit and the infiltration rate may be considerably greater than the natural infiltration rate. This effect can only be prevented by appropriate design of disposal units. In particular, a design should include adequate backfill or grout to prevent this type of failure.

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Summary and Conclusions

When near-surface disposal facilities are constructed using engineered cover systems, performance assessments require assumptions about the infiltration rate of water through an intact cover as it differs from the natural infiltration rate at a site and assumptions about changes in infiltration rate as the cover fails over time. At any site, many physical processes may occur that would tend to produce gradual changes in infiltration over time. However, those processes and resulting changes in infiltration are difficult to assess quantitatively. Therefore, as shown in Figure 5.9 (left), gradual changes in flow rates through waste are commonly represented by simple step changes in flow. Given the difficulties in modeling changes in infiltration through an engineered cover system over time, an assumption of multiple step changes, as represented in Figure 5.9 (left), is difficult to justify. Therefore, a single step change shown in Figure 5.9 (right) often is assumed in performance assessment. In this simple approach, the infiltration rate is assumed to increase instantaneously from its initial value for an intact cover system to a value representative of undisturbed soil at a site. Another assumption that is sometimes used in performance assessment is that the flow rate increases linearly with time from an initial value that applies to an intact cover system to the natural infiltration rate at the time the cover system is presumed to fail. Such an assumption also is difficult to justify.

Fig. 5.9. Representations of changes in flow through waste over time: step changes used to represent gradual failure of a cover system over time (left); single step change commonly assumed in performance assessments of engineered cover systems (right).

130 / 5. PERFORMANCE ASSESSMENT MODELS As in other aspects of current designs of near-surface disposal facilities, an engineer’s ability to design and construct an elaborate cover system may exceed an analyst’s ability to project its behavior with confidence. There is no generally accepted way to represent cover behavior for the purpose of estimating infiltration, and definitive guidance on the best way to describe infiltration in a performance assessment cannot be given. However, a cover system is an important part of a multiple-barrier approach to developing confidence in results of performance assessment, even if it is difficult to demonstrate beneficial effects of a cover on the consequences of waste disposal over long time periods using quantitative analysis. 5.3 Performance of Concrete Barriers Concrete is an important component of many near-surface low-level waste disposal facilities (Anderson and Pearson, 1993; Seitz and Walton, 1993). There also has been considerable interest in use of concrete as a barrier in decommissioning activities (Seitz, 2002; Snyder and Philip, 2002). Especially at sites with abundant precipitation, disposal facilities that include engineered concrete structures are expected to provide a greater degree of waste containment than facilities that are constructed entirely of earthen materials.25 Planned disposal systems frequently incorporate solid concrete monoliths in the form of vaults (Figure 5.10, top), modular canisters (Figure 5.10, bottom), or bunkers to enclose disposed waste. Such engineered concrete structures enhance containment of low-level waste by several physical and chemical means, including that they (1) provide structural support for an earthen cover, (2) delay and inhibit inflow of water to disposed waste, (3) supply additional adsorbing materials to retard movement of radionuclides into the surrounding environment, and (4) delay and inhibit release of leachate from a facility. Even though concrete barriers are used in many designs of near-surface disposal systems, there is very little information about the performance of concrete over periods of several hundred to thousands of years. There is evidence that, under certain conditions, some types of concrete can maintain their physical integrity and load-bearing capabilities for thousands of years. For example, 25Use of concrete structures may not be the best option in arid environments, where degradation by sulfate attack (Section 5.3.3.1) can be a greater concern than in wetter environments (Seitz and Walton, 1993; Walton et al., 1990).

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Fig. 5.10. Example disposal systems that include concrete monoliths: disposal in a concrete vault (top); disposal in modular concrete canisters (bottom).

the Great Dome of the Parthenon was constructed by the Romans with a lime and fly ash concrete more than 1,800 y ago (Cohen and Heun, 1979) and still stands today. The Romans also built aqueducts using similar cementitious materials, and remnants of those aqueducts can still be seen at many locations (Chang and Hasan, 1990). Although the types of historical concrete structures noted above have maintained a high degree of physical integrity for centuries, they have provided limited data of relevance to the problem of predicting the longevity of structures that have been built in the last 200 y using modern concrete formulations (e.g., Portland cement). Furthermore, they have provided little information on parameters that affect predictions of the long-term performance of modern concrete with respect to its ability to inhibit flow of water and transport of radionuclides from a disposal facility into the environment under different environmental conditions, and it is those properties of concrete that are of primary concern to performance assessment.

132 / 5. PERFORMANCE ASSESSMENT MODELS 5.3.1

General Approach to Modeling of Concrete Barriers

Models of degradation rates and times to failure of concrete barriers and resulting water infiltration rates through the barriers normally used in performance assessment are not intended to represent the actual behavior of concrete at future times. Rather, degradation of concrete barriers is described by assuming that presently known degradation mechanisms will be active for several hundred years and that hydrologic and chemical conditions at a site and facility will not change significantly in the future. Models of concrete degradation do, however, take into account changes in concrete characteristics, such as compressive strength, over time and changes in conditions that induce cracks in concrete. 5.3.2

Water Flow Through Concrete

Water can flow through pores, cracks, and joints in concrete. That flow is analogous to pore and fracture flow through earthen materials. Therefore, flow through concrete can be analyzed by attributing hydrogeologic parameters of earthen materials, such as hydraulic conductivity or permeability and specific moisture capacity, to concrete (Shuman et al., 1988). Furthermore, water flow through cracks and joints in concrete can be modeled with the same equations used to calculate flow through fractures in natural formations. However, potentially complicating factors that could substantially alter flow conditions include that cracks that develop in concrete over time may be filled by hydration products or may become connected depending on local stress conditions. Modeling of saturated and unsaturated flow through bulk concrete and surrounding soils is based on Darcy’s Law, as extended to unsaturated systems by Richards (1931). In unsaturated systems, capillary suction generally forces water to flow preferentially through smaller pores (Section 5.5). As moisture content increases, larger pores are filled and conductivity increases. When pores become nearly saturated, cracks begin to conduct water. Thus, the conductivity of a fractured concrete system is a complex function of the number of fractures, their size and distribution, the degree of moisture saturation, and properties of unfractured concrete. Flow through cracks and joints in concrete occurs primarily under saturated conditions and can be represented as flow between two parallel plates. In this approach, the opening (aperture) between plates is assumed to be well defined, and its relation to flow rate is well known. A shortcoming of this approach, however, is that it does not account for a variable crack thickness through concrete.

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Flow through each crack under saturated conditions can be described in a manner consistent with the notation of Darcy’s Law (Neuzil and Tracy, 1981)26 as:

where: Ksat q γ µ b dh----dz

dh q = K sat -------- , dz

(5.3)

γ 2 K sat = --------- b 12µ

(5.4)

= = = = =

equivalent saturated hydraulic conductivity (m s–1) Darcy flow velocity of water through crack (m s–1) specific weight of water (kg m–2 s–2) dynamic viscosity (kg m–1 s–1) segment aperture (meters)

=

gradient of hydraulic head (dimensionless)

Flow of water through the roof of a concrete vault depends on many factors, including properties of any overlying earthen cover, the amount of precipitation, the moisture tension above the vault, and the hydraulic conductivity of the vault roof. High-quality concrete without cracks has a very low hydraulic conductivity, typically ~10–10 to 10–12 m s–1 or less. However, even a single crack will cause the effective saturated hydraulic conductivity to increase by many orders of magnitude. For example, a crack of width 1 mm has a saturated hydraulic conductivity of ~1 m s–1. An equivalent hydraulic conductivity then can be estimated as an area weighting of hydraulic conductivities of fractures and the bulk matrix. The width and spacing of fractures can be estimated from standard fracture theory for concrete. If infiltration of water is not sufficient to support saturated flow through cracks in concrete, the hydraulic conductivity of the cracks decreases significantly. Under unsaturated conditions, and depending on such factors as the degree of saturation, width of the cracks, and hydraulic head, flow may be dominated by bulk flow through the concrete pores even when concrete is cracked. Such conditions can occur even in moist environments if an earthen cover over a vault restricts infiltration to the roof of the vault. However, water can accumulate above a vault relatively rapidly, even at infiltration rates as low as 0.01 m y–1 (Seitz et al., 1993), although 26Rogers,

V.C., Shuman, R. and Chau, N. (1994). “Modeling the longterm performance of concrete,” unpublished paper.

134 / 5. PERFORMANCE ASSESSMENT MODELS proper design of the vault roof and cover can provide for drainage away from the roof. Thus, an accurate description of the performance of concrete barriers requires an analysis of the entire system that takes into account the precipitation rate and the performance of an earthen cover, as well as concrete barriers. While an important function of a concrete roof is to delay and inhibit infiltration of water into disposed waste, the function of a vault floor is somewhat counterintuitive. A concrete floor separates waste from soils in the foundation, and it provides structural stability and inhibits release of radionuclides to the surrounding environment. However, given the potential for the so-called bathtub effect, it may be preferable to include a drain in the floor and to make the floor more permeable than the roof, in order to prevent accumulation and overflow of water that enters a vault (Dolinar et al., 1994; Seitz and Walton, 1993; Shuman et al., 1988) . Water flow through concrete generally increases as cracks develop. Development of cracks and their depth, width and spacing are time-dependent, so flow through concrete also is time-dependent. In most performance assessments, concrete failure is defined as occurring when water can flow through cracks and joints. Long-term degradation of a waste form also is considered when it is a concrete-type material, such as grout. 5.3.3

Degradation of Concrete

Several physical and chemical processes cause concrete barriers to crack over time. Differential settling of foundation materials, the floor, or materials such as waste packages that may be supporting the roof can cause cracking. Such settling can be minimized by proper design and construction of a facility. Chemical reactions or processes that affect concrete or reinforcing steel may weaken concrete, and applied loads and moments may induce cracking of weakened material. The five most important chemical reactions or processes that affect the performance of concrete (Baird et al., 1987; Snyder, 2001; 2003; Snyder and Clifton, 1995)27 are shown in Figure 5.11. In general, those reactions or processes can be divided into three categories: (1) those that degrade the surface of concrete, thus decreasing the amount of concrete that can support applied loads; (2) those that degrade bulk properties of concrete, thus reducing its compressive strength; and (3) those that corrode reinforcing steel in concrete, thus leading to 27Rogers,

V.C., Shuman, R. and Chau, N. (1994). “Modeling the long-term performance of concrete,” unpublished paper.

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Fig. 5.11. Depiction of different mechanisms of degradation of concrete and their effects on performance.

further degradation. Excellent reviews of these mechanisms are given by Clifton and Knab (1989) and Walton et al. (1990). Various reactions or processes, including sulfate attack, freeze/ thaw cycling, calcium leaching, alkali-aggregate reaction, and corrosion of reinforcing steel, are discussed in the following sections. 5.3.3.1 Sulfate Attack. Sulfate attack usually is the primary mechanism of surface degradation of concrete, even in sulfate-resistant formulations. In an empirical approach to modeling sulfate attack (Atkinson and Hearne, 1984), the thickness of degraded concrete is assumed to be proportional to the amount of tricalcium aluminate [(CaO)3Al2O3] in concrete, the concentration of sulfate (SO4) and magnesium ions in water that contacts concrete, the diffusion coefficient of sulfate ions in concrete, and time. Since the diffusion coefficient of sulfate ions is directly related to the water/cement ratio, some formulations of the model substitute that ratio for the diffusion coefficient.28 Although sulfate attack is linear with time, 28For example, Walton et al. (1990) estimate the depth of concrete degradation (x) over time from magnesium and sulfate attack as x(t) = 0.55 CS[(Mg)2+ + (SO4)2+]t, where x is in centimeters, CS is the weight percent of tricalcium aluminate in unhydrated cement, (Mg)2+ and (SO4)2+ are the concentrations of magnesium and sulfate ions in bulk solution (mole L–1), and t is the time in years.

136 / 5. PERFORMANCE ASSESSMENT MODELS it still shows diffusion characteristics because the surface of concrete spalls off after a short time, thus creating a clean surface for further reactions. The weight percentage of (CaO)3Al2O3 influences degradation caused by sulfate attack because its hydration products are the principal reactants with sulfate ions. Laboratory tests indicate that the amount of (CaO)3Al2O3 in concrete typically varies by ±30 %. Field experiments in which sulfate penetration profiles in concrete are measured have been used to analyze sulfate attack. Data collected from those experiments are consistent with relationships discussed above, in that samples showed penetrations of 8 to 20 mm over a 5 y period compared with a predicted penetration of 11 mm. As sulfate attack progresses into concrete, the affected layers spall off, thus effectively reducing the thickness of the remaining intact structure. As the structure degrades, it reaches a point where it can no longer support loads imposed on it. At this point, cracking may be assumed to occur. A more mechanistic diffusion approach to modeling sulfate attack is given by Atkinson and Hearne (1990). However, the empirical approach described above provides better agreement with experimental data. 5.3.3.2 Freeze/Thaw Cycling. Freeze/thaw cycling occurs in cold climates and can be eliminated by use of a sufficient thickness of earthen material (~1 m or greater) above the concrete. This process mainly degrades the surface of high-quality concretes proposed for use in disposal facilities, due to their very low permeability and, thus, very low penetration of water into concrete before freezing occurs. Effects of freeze/thaw cycling are greatly reduced when air is entrained in concrete. Freeze/thaw degradation occurs only in regions with 70 to 80 % saturation, and several significant freezing and thawing periods must occur before degradation begins. Effects of freeze/thaw cycling generally are threshold in nature, meaning that a concrete member can undergo at least 50 freeze/thaw cycles before degradation occurs (Baird et al., 1987). Once degradation begins, it increases approximately linearly with time when averaged over several cycles, because degraded concrete spalls off, thus exposing a new surface. 5.3.3.3 Calcium Leaching. Changes in bulk properties of concrete mainly occur when calcium in concrete forms calcium hydroxide and is leached from the concrete, thus changing its density and reducing its compressive strength. In some analyses, however, calcium hydroxide leaching is considered to be a form of surface

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degradation. One analysis of degradation of a concrete vault due to calcium hydroxide leaching showed that every percent loss of calcium causes a 1.5 % loss in yield strength of the concrete.29 5.3.3.4 Alkali-Aggregate Reaction. The alkali-aggregate reaction also affects bulk properties of concrete. In this degradation mechanism, the aggregate reacts with alkaline materials throughout the bulk of a concrete member. Products of this reaction form in pore spaces, thus reducing the porosity of concrete. This process initially reduces the permeability and increases the strength of concrete. However, after pore spaces are filled, the alkali-aggregate reaction causes increased internal stresses that create internal cracks, which can reduce the yield strength of concrete and increase its permeability. While alkali-aggregate reactions and their related equations are known, their reaction rates and products can vary greatly. 5.3.3.5 Corrosion of Reinforcing Steel. Corrosion of steel that is used to reinforce concrete is an important cause of concrete degradation. For reinforcing steel to corrode, the surface of the steel first must be depassivated. Depassivation is caused by a carbonation reaction, which reduces the pH of concrete, or by an increase in chloride ion concentration in the vicinity of the surface of steel. In general, chloride ion concentrations in water are sufficiently high that the chloride ion mechanism almost always depassivates steel before a carbonation reaction can reduce the pH to a level that would have the same effect. Once steel is depassivated, oxygen that diffuses through concrete to reinforcing steel begins to react with iron, forming iron oxide (rust). Because iron oxide occupies a larger volume, it creates internal pressures that initiate cracking and eventual spalling of concrete that surrounds reinforcing steel. A key factor in corrosion of reinforcing steel in concrete is the time to onset of corrosion. An empirical relationship that can be used to estimate this time was developed on the basis of regression analyses of concretes that are appropriate for use in highway construction (Clear, 1976). The time to onset of steel corrosion is given by: 1.22

129 xc -, t c = -------------------------------------0.42 ( WCR ) [ Cl ] 29Rogers,

(5.5)

V.C., Shuman, R. and Chau, N. (1994). “Modeling the long-term performance of concrete,” unpublished paper.

138 / 5. PERFORMANCE ASSESSMENT MODELS where: = tc xc = WCR = [Cl] =

time to onset of corrosion (years) thickness of concrete surrounding the reinforcing steel (inches) water-to-cement ratio chloride ion concentration in water (parts per million)

Thus, the time to onset of corrosion increases with increasing thickness of concrete that surrounds reinforcing steel and increases with decreasing water-to-cement ratio and chloride ion concentration in water that contacts the concrete. Damage to concrete that results from corrosion of steel reinforcement manifests itself in expansion, cracking and spalling of a concrete member. A reinforced concrete member may suffer structural damage due to loss of the bond between steel and concrete and consequent reduction in reinforced cross-sectional area. Steel reinforcement generally is passivated as a result of the alkaline nature of the liquid phase in concrete pores and, hence, does not undergo corrosion. However, this passive layer may be destroyed by means of a lowering of the pH of concrete via carbonation or chloride ion penetration to locations of steel. Using a standard solution of Fick’s First Law of Diffusion, the chloride ion concentration at steel reinforcement as a function of time (t) is calculated as: xc ⎧ ⎫ , [ Cl s ] = [ Cl i ] + ( [ Cl w ] – [ Cl i ] ) ⎨ 1 – erf ----------------------0.5 ⎬ 2 ( D Cl t ) ⎩ ⎭ where: [Cls] = [Cli] = [Clw]= xc = DCl =

(5.6)

chloride ion concentration at steel reinforcement (mole m–3) initial chloride ion concentration in concrete (mole m–3) chloride ion concentration in water (mole m–3) thickness of concrete surrounding steel reinforcement (meters) effective diffusivity of chloride in concrete (m2 s–1).

The concentration of chloride ions at steel reinforcement that is required to depassivate steel has been considered by several investigators. Hausmann (1967) found that the pH of concrete affects the concentration of chloride ions that is necessary to initiate corrosion. In studies carried out using NaOH and Ca(OH)2 solutions, a chloride-ion to hydroxide-ion concentration ratio of 0.61 was found to be sufficient to depassivate steel. Equation 5.5 is used to calculate the time at which depassivation of steel occurs.

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Upon depassivation of steel reinforcement by either carbonation or chloride ion penetration, corrosion is assumed to occur at a rate that is determined by the rate of oxygen diffusion to the location of steel. The molar flow of oxygen to the surface of steel reinforcement is modeled using Fick’s First Law of Diffusion as: d[O 2 ] -, J Ο = – D O A --------------dx where: JO = DO = A =

(5.7)

oxygen flow at steel reinforcement (kg s–1) effective diffusivity of oxygen through concrete (m2 s–1) surface area over which oxygen diffuses to steel reinforcement (m2)

d [ O2 ] -------------- = gradient of dissolved oxygen concentration (kg m–4) dx The rate at which the corrosion reaction consumes oxygen is assumed to be greater than the rate of oxygen diffusion to the reaction surface. Under such conditions, the corrosion rate is limited by the flux of oxygen at the steel reinforcement. An epoxy coating on steel reinforcement may delay the onset of corrosion by isolating the steel from aggressive ions and oxygen. While an epoxy coating is not assumed to delay the time of depassivation of reinforcing steel, it is assumed to prevent corrosion for as long as it remains intact. The few attempts to model the effectiveness and degradation of epoxy-coated steel reinforcement have used a linear failure function. Using the time at which failure of a coating begins and the time required for all epoxy coating to fail, a fraction of the failed reinforcement coating is calculated. This fraction is used to decrease the projected rate of corrosion. 5.3.3.6 Combination of Reactions. All reactions shown in Figure 5.11 lead to progressive structural deterioration. This deterioration reduces the ability of concrete to support physical and thermal loads, which leads to enhanced crack formation. When crack formation is sufficient, water flows freely through cracks, and infiltration and movement of radionuclides increase. The different types of reactions also must be considered collectively, because effects of one degradation mechanism may affect degradation rates of other reaction mechanisms. For example, surface attack increases the rate at which chloride ions can diffuse to reinforcing steel and initiate corrosion.

140 / 5. PERFORMANCE ASSESSMENT MODELS 5.3.4

Application of Models

Models that integrate concrete barrier degradation, structural integrity, water infiltration, radionuclide release, and transport of radionuclides in groundwater are becoming available; such a model (Snyder and Clifton, 1995) is discussed in the following section. However, most performance assessments of disposal facilities that are constructed with concrete barriers analyze concrete degradation separately and use the results as input to standard models of infiltration and leaching of radionuclides from disposed waste. Modeling of the performance of concrete barriers is concerned with (1) physical degradation of concrete with respect to its ability to provide structural support and to delay and inhibit inflow of water to waste and (2) chemical properties of concrete with respect to its ability to retard movement of radionuclides into the surrounding environment. The ability of concrete to provide a chemical barrier to release of radionuclides generally should persist over longer time scales than its physical integrity. Concrete degradation has little effect on infiltration and leaching of radionuclides until a concrete member cracks due to physical and thermal loads. The time to onset of cracking will depend on the design of a facility and local environmental conditions. Once cracking starts, it usually proceeds rapidly, and infiltration of water through a concrete barrier is expected to increase rapidly during this time. The infiltration rate through cracked concrete will depend on the design of a facility. Release of radionuclides from a waste form also increases during the time period of concrete degradation. A simple approximation that is frequently used in performance assessment is to assume that once a concrete barrier cracks, the disposal system, including waste forms, is in a fully degraded condition. The infiltration rate of water and release rates of radionuclides from waste forms that are appropriate to the degraded condition then are assumed to be constant for all subsequent times. An important issue for performance assessment is the need to assess and justify an assumed infiltration rate under degraded conditions compared with the natural infiltration rate under undisturbed conditions at a site. The infiltration rate under degraded conditions could be substantially greater than the natural infiltration rate, due to the presence of preferential flow paths in a degraded system. 5.3.5

Example Analyses of Long-Term Performance of Concrete Barriers

Shuman et al. (1988) conducted an analysis of concrete degradation in a low-level waste disposal facility that is constructed using

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below-ground concrete vaults (Figure 5.10, top). A vault was assumed to be divided into smaller rooms or cells to provide structural support for the roof and to provide shielding of workers during loading of waste. The roof of the facility was designed to maintain its integrity for a longer period of time in order to avoid excessive accumulation of water in the vaults. The analysis assumed that ion concentrations in contact with concrete are the same as the concentrations in groundwater, although infiltrating water may have ion concentrations significantly less than those in groundwater. The effect of an epoxy coating on reinforcing steel was ignored. The analysis of concrete degradation for this system gave a surface degradation rate from sulfate attack of 11 mm per 100 y and a time to onset of corrosion of steel reinforcement of 270 y. As steel reinforcement corrodes, the concrete barrier loses most of its tensile strength. Depending on the degree of conservatism in the design, loss of tensile strength may lead to significant cracking within 500 y. However, the concrete vault in this example was designed to withstand the load even after steel reinforcement corrodes, so that significant cracking was not predicted to occur until ~2,000 y after facility closure. The predicted degradation rate obtained in this example is about an order of magnitude lower than degradation rates estimated based on field experiments (Section 5.3.3.1). The current state-of-the-art of modeling of the performance of buried concrete structures is represented by work at the National Institute of Standards and Technology (NIST) (Snyder, 2001; 2003; Snyder and Clifton, 1995), which is intended to evaluate the service life of a buried concrete vault, including its structural degradation and changes in permeability. In contrast to other models that consider degradation mechanisms individually, the NIST model considers the combined effects of multiple mechanisms and, thus, incorporates synergisms of multiple degradation processes. This is accomplished by using coupled reactive-transport equations for various ion species that degrade concrete by different mechanisms. As various ions diffuse into concrete, transport properties change, resulting in changes in the rate of ion diffusion. As the pH of the pore solution changes, available salts either precipitate or go into solution. Precipitation and dissolution processes change the porosity, which in turn changes transport properties. The model thus solves coupled processes of flow, solute transport, and chemical reactions, and synergisms of different degradation mechanisms are taken into account by maintaining chemical equilibrium in the system (Snyder and Clifton, 1995). During a simulation, the model computes the rate of infiltration of water through a degrading concrete roof.

142 / 5. PERFORMANCE ASSESSMENT MODELS Snyder (2001) performed limited comparisons of the NIST model with data and found satisfactory agreement with observations for individual aspects of the model (e.g., porosity and chloride penetration). Those comparisons provide evidence that the model is correct and has a reasonable physical basis. Snyder (2003) has provided guidance on how the current state of a concrete structure can be used to project its remaining service life. The NIST model can be run probabilistically by taking into account uncertainties in various input parameters (Snyder, 2001). In an example calculation that involved 10 iterations (Snyder, 2001), the predicted service life of a concrete structure varied from 80 to 500 y, and the median service life was ~250 y. Modeling of the width and spacing of cracks also can be performed probabilistically by taking into account uncertainties in material and structural properties of concrete. In the NIST model, the most important factor that affects ionic transport through concrete is the presence of cracks, which control the quantity of ions transported and whether there will be convective transport in the absence of an external hydraulic pressure. Even though models that consider combined effects of multiple degradation mechanisms, such as the NIST model described above, are becoming available, there still can be considerable value in understanding individual mechanisms and an ability of particular design features to enhance the performance of concrete (Seitz and Walton, 1993; Walton et al., 1990). For example, a useful design feature would be to provide a relatively high permeability pathway for flow of water around a concrete structure that has a lower permeability material immediately adjacent to the concrete, in order to limit the maximum flow rate of water into cracks in the concrete. An ability to consider degradation mechanisms individually should continue to be important in performance assessment to achieve an adequate understanding of the long-term performance of concrete. Predictions that the service life of concrete will exceed a thousand years, as obtained in the analysis by Shuman et al. (1988) described above, may be difficult to justify in performance assessments. However, it also is important to recognize that a much shorter time to loss of physical integrity of concrete barriers would still be sufficient to provide a substantial benefit in limiting releases of radionuclides that have half-lives of ~30 y or less, and that times to failure beyond a thousand years would provide little benefit in limiting releases of important radionuclides that have half-lives of thousands of years or greater. Therefore, efforts to justify assumptions that concrete barriers will inhibit infiltration and outflow of water for time periods beyond a few hundred years may

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not be warranted in regard to expected improvements in facility performance. An ability of concrete to provide a chemical barrier to release of radionuclides should be maintained beyond the period of its physical integrity. 5.3.6

Summary

Saturated and unsaturated flow through concrete can be modeled in the same way as flow through earthen materials. Water flow through cracks in concrete can be modeled as fracture flow, and significant infiltration through concrete occurs only after significant cracking. Processes that degrade concrete used in low-level waste disposal facilities can be grouped into three categories: (1) those that degrade the surface of concrete, (2) those that affect bulk properties of concrete, and (3) those that corrode reinforcing steel in concrete. Processes that degrade the surface of concrete proceed linearly in time. Sulfate attack is the most important mechanism of surface degradation of concrete, and it results in a continued thinning of a concrete member that can support a load. Leaching of calcium hydroxide from concrete is the mechanism that most affects bulk properties of concrete, such as yield strength and permeability. Corrosion of reinforcing steel in concrete has been shown to be a threshold reaction that requires depassivation of steel before oxygen in the vicinity of steel can react with iron, thus forming iron oxide, which creates internal pressures that cause concrete cracking. Once corrosion of reinforcing steel commences, it proceeds quite rapidly and causes a significant decrease in the overall ability of concrete facility to withstand loads. Because concrete is a manufactured material, it is subject to problems of quality control. To ensure satisfactory long-term performance, concrete must be carefully mixed and poured to specifications. Small deviations from specifications can significantly affect long-term performance. Thus, it is critical that strict quality control and quality assurance procedures be implemented and enforced when using concrete barriers in low-level waste disposal facilities. 5.4 Source Term In this Report, “source term” denotes the rate of release of radionuclides from a disposal facility as a function of time after waste emplacement and facility closure. There is frequent confusion over the meanings of “source term” and “inventory,” with the latter term

144 / 5. PERFORMANCE ASSESSMENT MODELS denoting quantities of radionuclides (usually activities) in disposed waste, and the two terms are often, and mistakenly, used synonymously. The definition of “source term” used in this Report is consistent with its traditional meaning (e.g., in analyses of nuclear reactor accidents). The inventory of radionuclides in disposed waste is just one of many factors that determine the source term. Other important factors are the physical and chemical properties of disposed waste, physical and chemical processes that influence releases from waste forms and transport from a facility, including processes that degrade engineered barriers, waste containers, and waste forms, and processes that alter chemical conditions within a facility. Releases of radionuclides from a disposal facility can be in the form of liquids or gases. Releases in water often are assumed to be the most important, but releases in the gas phase (e.g., 3H, 14C, and radon) should not be ignored without proper justification. Estimates of releases in the liquid phase are used as input to vadose-zone and groundwater transport models, and estimates of gaseous releases are used as input to atmospheric transport models. The source term is strongly coupled to the flow rate of water through a facility and, therefore, to the performance of the cover and other engineered barriers discussed in Sections 5.2 and 5.3 and to properties of the surrounding environment at a site.

5.4.1

Inventories of Radionuclides

Sources of low-level radioactive waste are many and diverse (Section 3.2), and these wastes exhibit a wide range of physical and chemical properties. Many radionuclides may occur in low-level waste, and in concentrations that may vary by many orders of magnitude. In addition, volumes, radionuclide concentrations, and physical and chemical properties likely will change as the types of waste being generated change over time. For example, as more decontamination and decommissioning activities are conducted at nuclear facilities, volumes of contaminated building materials and soils that require disposal should increase substantially. Given the wide range of physical and chemical properties of low-level waste, development of a detailed conceptual model of releases of radionuclides from a disposal facility is a challenging task in performance assessment. In most cases, it is not possible to model every single waste container and waste form. The usual approach is to group wastes into distinct categories that are assumed to have common performance characteristics.

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When generators send low-level waste to disposal facilities that are licensed by NRC or an Agreement State, they are required to estimate activities of specific radionuclides and describe the types of waste and waste containers. Descriptions of waste types provide information on processes used to generate the wastes. Descriptors of types of low-level waste used in the United States include compactible trash, noncompactible trash, cation ion-exchange media, soil, and filter media (Dehmel et al., 1994; Yim, 1994). NRC has developed a uniform shipping manifest for use by low-level waste generators (NRC, 1995). This manifest includes the 25 waste classifications listed in Table 5.1. Use of descriptors of waste forms in estimating release rates of radionuclides from a disposal facility is discussed in Section 5.4.5. In some cases, accurate records of past disposals have not been maintained. Wastes often were described as consisting of “mixed fission products” or “transuranics,” or they were characterized only by measurements of gross alpha, beta or gamma activity. In such cases, estimates of radionuclide inventories usually can be TABLE 5.1—Waste forms listed on shipping manifests for non-DOE low-level waste in the United States (NRC, 1995). Charcoal

Demolition rubble

Evaporator bottoms, sludges, concentrates

Incinerator ash

Cation ion-exchange media

Compactible trash

Soil

Anion ion-exchange media

Noncompactible trash

Gas

Mixed-bed ion-exchange media

Animal carcass

Oil

Contaminated equipment

Biological materials (except animal carcasses)

Aqueous liquid

Organic liquid (except oil)

Activated material

Filter media

Glassware or labware

Other

Mechanical filter

Sealed source or device

EPA or state hazardous

Paint or plating

146 / 5. PERFORMANCE ASSESSMENT MODELS improved by review of processes used to generate the waste, review of other documentation (e.g., process records) to help identify the wastes, and discussions with people involved in waste generation and disposal activities at the time. Those types of review processes have been used at DOE facilities to determine waste disposal practices and quantities of radionuclides disposed up until the late 1980s. In the absence of reliable information, conservative estimates of radionuclide inventories need to be made and carefully documented in assessing the long-term consequences of past disposal practices. However, care should be taken to avoid use of extreme conservatism in estimating inventories in past disposals. The goal should be to use sufficient conservatism to provide reasonable assurance that potential impacts of past disposals on the public would not be underestimated. Hundreds of radionuclides can occur in low-level waste, but only a few normally contribute significantly to projected doses. Almost all of the activity in non-DOE low-level waste is due to shorter-lived radionuclides (i.e., radionuclides with half-lives less than ~30 y), including 137Cs, 60Co, 90Sr, 3H, and 55Fe (Cowgill and Sullivan, 1993). The short half-lives of these radionuclides combined with their expected travel times to locations beyond the boundary of a disposal facility and the expected performance of waste containers and waste forms normally preclude them from being an important concern in regard to projected doses due to releases in groundwater. However, their large inventories may result in high doses in scenarios that involve exposure of future inadvertent intruders (Section 6). The most important contributors to projected doses due to releases in groundwater tend to be long-lived, soluble, and mobile radionuclides, including 129I, 99Tc, 14C, and 36Cl. Under certain chemical conditions, isotopes of uranium can be important. Other radionuclides that are considered by NRC to be important in low-level waste, especially in regard to protection of future inadvertent intruders, are listed in Table 3.3 (Section 3.4.2.1). Justification for not considering specific radionuclides for detailed analysis in a performance assessment must be provided on a site- and scenario-specific basis. DOE sites also have produced substantial amounts of low-level waste in carrying out their missions. The total volume of waste generated annually at DOE sites is substantially greater than the volume of other low-level waste (DOE, 1997a). The types of waste generated at DOE sites and their physical properties are variable and consist of many of the same types of waste as are generated in non-DOE activities (e.g., dry solids, activated metals, ionexchange resins). However, because of differences in processes of

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waste generation, radionuclide compositions of DOE and non-DOE low-level wastes often differ substantially. Nonetheless, the principal contributors to projected doses due to releases to the environment are long-lived, soluble, and mobile radionuclides. Inventories of radionuclides in disposed waste provide an initial condition for estimating releases from a disposal facility. Accurate representation of the source term requires knowledge of physical and chemical forms of radionuclides, their incorporation into specific waste forms, and their disposal in specific containers. However, information in disposal records often is not sufficiently detailed to assign the activity of a radionuclide to a particular container and waste type. Estimation of activities of radionuclides in each waste type or container generally requires an examination of each shipping manifest. This process can be facilitated by developing computerized databases to store this information (Sandwina, 1991). In modeling the source term, radionuclide inventories often are represented as averages over a disposal facility or, for nonDOE wastes, as averages over all Class-A, -B or -C wastes (Section 3.4.2.1). This homogenization greatly simplifies an analysis of the source term, compared with calculating releases from each waste package or disposal cell. However, this approach may result in an underprediction of releases from regions with above-average activities. Effects of an assumed homogenization of radionuclide inventories in a disposal facility normally are treated on a case-bycase basis, because a generally accepted approach has not been developed (Section 5.4.8.3.4).

5.4.2

Radionuclide Release Rates (Source Term)

The rate of release of radionuclides from a waste disposal facility is expressed in units of mass per unit time. Since releases usually are assumed to occur in water, an important design objective of many disposal facilities is to reduce the flow of water to wastecontaining regions. Many design concepts have been used in disposal facilities for low-level waste. The principal components of various facility designs are described below. • Engineered structures typically include reinforced concrete walls, floors and roofs used to isolate waste from water (Section 5.3) and to prevent intrusion by plants, animals and humans. Other engineered concepts include mined cavities and liners in augured shafts.

148 / 5. PERFORMANCE ASSESSMENT MODELS • Covers (or caps) are engineered systems placed above disposed waste for the purpose of diverting water away from waste-containing regions. Covers also may be designed to function as a biointrusion barrier. Different types of cover systems are discussed in Section 5.2. • Containers are external packaging used when disposing of waste. Different types of containers include wooden crates, cardboard boxes, carbon steel drums or boxes, stainless steel drums, high-density polyethylene (HDPE), and concrete. Some wastes are not placed in containers but are placed directly into a disposal facility. Wastes that are frequently disposed of without a container include equipment components (e.g., large activated metal materials, such as reactor pressure vessel parts) and soil or rubble from decontamination activities. • Waste forms are materials in which radionuclides are embedded, the physical and chemical properties of which are important to an analysis of the source term. Many wastes are treated prior to disposal, but some wastes are not treated. Untreated waste is not altered substantially from its form at the time of generation. Untreated waste includes all waste disposed of without a container, such as the types of waste noted above, as well as many wastes that are placed directly in a container (e.g., contaminated laboratory trash and filter media). Treated waste is waste that has undergone further processing prior to disposal to reduce the potential for release. Waste treatment may include solidification in cement, bitumen, polyethylene or other encapsulating agents, dewatering or addition of sorbents to remove free liquids, and volume reduction by incineration or compaction. Estimation of the source term involves considerations of the degradation of engineered structures to an extent sufficient to permit fluids (air and water) to contact waste containers, degradation of waste containers to permit access of fluids to waste, release of radionuclides from a waste form, and transport of released radionuclides in fluids from a disposal facility. A flow chart of processes that are involved in modeling the source term is shown in Figure 5.12. This diagram indicates the principal components of a source-term analysis and couplings between them. The central core that consists of the inventory, container performance, waste-form performance, and transport components represents intrinsic factors that lead to release of radionuclides. External factors, which are represented by infiltration and chemistry, influence the intrinsic

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Fig. 5.12. Depiction of different components of source-term analysis. Solid lines denote strong coupling between components, and dashed line denotes a weak coupling. Arrows pointing in each direction indicate that two components influence each other.

factors to control release. Arrows leading from one factor to another denote a coupling between the factors, with arrows pointing in both directions indicating that each factor can influence the other. A dashed arrow indicates a relatively weak coupling of factors. The remainder of this Section provides a discussion of factors that control container and waste-form performance and transport of radionuclides from a disposal facility, including discussions of couplings between components of a source-term analysis indicated in Figure 5.12. Infiltration and radionuclide inventories are discussed in Sections 5.2 and 5.4.1, respectively. A typical near-surface disposal facility consists of a system of disposal units, sometimes referred to as cells. A diagram of a disposal cell is shown in Figure 5.13. In this example, the cell includes a cover system to divert water away from disposed waste, an engineered barrier system to further protect waste from access by infiltrating water and inadvertent human intrusion, and the waste-containing region. The dimensions of a cell are such that

150 / 5. PERFORMANCE ASSESSMENT MODELS

Fig. 5.13. Depiction of release of radionuclides (source term) from a low-level waste disposal unit due to infiltrating water.

hundreds to thousands of containers can be emplaced within each cell. Each container may have a waste form with unique physical and chemical characteristics. The inventory in Figure 5.12 represents the mass of radionuclides in a disposal facility at the time of facility closure, which is when an analysis begins. Ideally, distributions of inventories by waste container and waste form are known and this information is used in estimating releases from the waste forms. This relationship is represented by arrows leading from the inventory to the waste form and container components in Figure 5.12. In general, the chemical form of waste can influence container performance by means of interactions that cause degradation of a container, but these effects normally are not considered explicitly in source-term modeling. The physical and chemical form of waste will directly impact waste-form performance. Container performance is described in terms of the ability of a container to control access of air and water to a waste form. This relationship is represented in Figure 5.12 by the arrow leading from the container to the waste-form performance component. Except for wooden crates or cardboard boxes, an intact container will provide a barrier to inhibit contact between wastes and aqueous fluids. Container performance can be influenced by chemistry and the infiltration rate. Chemical reactions (e.g., corrosion) are the primary means by which containers degrade. As container degradation occurs,

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those reactions can alter the local chemical environment, especially redox and pH conditions. Although container performance typically is assumed to be independent of infiltration rate, subsidence of containers and wastes following degradation can influence the infiltration rate into a waste-containing region. Waste-form performance is described in terms of the ability of a waste form to control release of radionuclides to contacting fluids. Waste-form performance can depend on the initial physical and chemical form of waste, the time of container failure, the local chemical environment, the flow rate of fluids, and the rate of transport away from a waste form. Release of radionuclides and other materials in waste may alter the local chemical environment. Releases from a waste form provide a source of radionuclides for transport from a disposal facility. Transport processes are strongly influenced by the chemical environment, especially the solubility and sorption properties of radionuclides or their carriers, as well as flow rates of liquids and gases. Transport of materials released from engineered barriers, containers, and waste forms as a result of chemical reactions will alter the chemical environment. External factors termed “fluid infiltration” and “chemistry” in Figure 5.12 are each combinations of many processes. Infiltration is the rate of fluid flow into a disposal facility. Infiltration in the aqueous phase is a function of precipitation, runoff, transpiration, erosion, cover performance, and engineered barrier performance, including the presence of localized cracking. Exfiltration in the gas phase is a function of cover performance and atmospheric pressure (e.g., changes in barometric pressure). Chemical reactions can influence infiltration through precipitation and dissolution reactions, which plug or unplug flow paths and alter hydraulic properties of engineered barriers. The external factor termed “chemistry” in Figure 5.12 consists of all chemical reactions, including microbially facilitated reactions, that can alter the mobility of radionuclides. Chemical controls on transport include solubility limits and sorption reactions. Solubility limits may restrict amounts of radionuclides in solution, with excess amounts forming an immobile precipitate. Chemical reactions may be reversible or irreversible. Such reactions as surface adsorption and ion exchange act to reduce solution concentrations and retard movement of radionuclides. Complexation of radionuclides with ligands and chelating agents can enhance the mobility of radionuclides by increasing their solubility and reducing their tendency to adsorb. In addition to solution reactions, colloidal particles may form. Colloids are naturally occurring suspended particles with

152 / 5. PERFORMANCE ASSESSMENT MODELS diameters <10 µm. Colloids have different transport properties than aqueous solution species and, in some cases, can serve as an important mechanism for increasing the mobility of radionuclides. In most source-term analyses, solution chemistry is modeled using a few global parameters, such as solubility limits and solid/solution distribution coefficients. Chemistry also can have a significant impact on releases of radionuclides from waste forms. Those releases will depend on the chemical composition of waste. For example, if a radionuclide is incorporated into waste as a compound with low solubility in the environment of a disposal facility, its release rate will be very low; in contrast, if a radionuclide is essentially free to be released on contact with water, the release rate will be relatively high. Often, releases from a waste form are controlled by solubility limits or sorption reactions. The preceding discussion indicates the complexity of analyzing the source term using detailed, mechanistic models. In performance assessment, however, it is not necessary, nor even desirable, to incorporate a detailed model into an analysis. Performance assessment is used to obtain defensible projections of dose for the purpose of demonstrating compliance with regulatory requirements (Section 2). Therefore, in developing a suitable source-term model for use in performance assessment, it is the responsibility of an analyst to examine the various physical and chemical processes that may occur in a disposal facility, decide which of those processes are most important in controlling releases of radionuclides from the facility, select appropriate models and data to represent the source term, and justify the choices. The source term for each radionuclide may be represented as a point estimate that varies over time or, if uncertainties are addressed explicitly, as a range of values. Justification of data and models may be provided by more detailed analyses, experimental data, or properly documented judgment. Sections 5.4.4 and 5.4.6 provide guidance on selection of models of container performance, waste-form performance, and transport components of the source term, including discussions of required information, sources of information, and the state-of-the-art of modeling. Interfaces between those models and other models used in performance assessment are considered in Section 5.4.7. Unresolved issues in source-term modeling are discussed in Section 5.4.8. Prior to these discussions, different disposal concepts that have been proposed and their impact on analyses of source terms are discussed in the following section.

5.4 SOURCE TERM

5.4.3

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Disposal Facility Concepts

Several disposal concepts for low-level waste have been used or are under consideration, including above-ground vaults, belowground vaults, earth-mounded concrete bunkers, mined cavities, shafts or augured holes, and shallow trenches (Bennett, 1985; Bennett and Warriner, 1985; Bennett et al., 1984; Miller and Bennett, 1985; Warriner and Bennett, 1985). Each disposal concept has different source-term characteristics and issues. Therefore, analysis of a source term must address the role of individual components in a disposal concept that are relied upon to mitigate releases. In general, those components include engineered features of a disposal facility, waste containers, and waste forms. The objective of a source-term analysis is to obtain justifiable projections of the performance of system components that control release and transport of radionuclides from a facility. • Above-ground vaults are engineered structures, typically reinforced concrete, that are intended to provide an infiltration and intrusion barrier. An important source-term consideration for this design is an almost exclusive reliance on a vault to control releases at all times, including times beyond a period of institutional control. Extreme events, such as earthquakes, tornadoes, and unusually heavy rain or snow storms, may need to be addressed, because relatively large releases could be caused by a single event. For example, a tornado could weaken or breach an engineered structure and lead to an immediate increase in releases. This reliance of performance on an absence of extreme events may suggest the need for a probabilistic analysis, in which the likelihood of an event is estimated and projected consequences of that event are calculated. However, given that performance objectives for low-level waste disposal facilities in the United States are expressed in terms of dose (Section 3.4.2) and that those criteria presumably apply only to reasonably likely processes and events, it is not clear how extreme events would be addressed in demonstrations of compliance. In addition, long-term chronic effects due to natural variations in weather need to be addressed in evaluating the performance of an above-ground structure (e.g., temperature, humidity, precipitation, freeze/thaw cycles). Another consideration that is likely to be more important for above-ground vaults than for other disposal concepts is shielding from external radiation. External exposure is not expected to be a concern for off-site members of the public,

154 / 5. PERFORMANCE ASSESSMENT MODELS but it may be an important consideration in some scenarios for inadvertent intrusion. Finally, compared with other disposal concepts, runoff of waste materials to streams or other surface water and releases to the atmosphere may be important considerations in estimating source terms. • Below-ground vaults are engineered structures, typically reinforced concrete, that are covered by several meters of soil and other materials to further isolate waste. A vault often is covered by an engineered cap that is designed to divert water away from waste (Section 5.2). The cap and vault provide independent controls on water flow through waste-containing regions. Some vault designs use a concrete floor with a system of drains to monitor releases (Lopez et al., 1993), whereas other designs are free draining (Dolinar et al., 1994). The latter designs aim to prevent a vault from filling with water in case drains fail. In all below-ground disposal systems, releases to groundwater usually are found to be the most important pathway to the environment. However, releases to the atmosphere still must be considered for radionuclides that can exist in the gas phase (e.g., radon, 14C, and 3H). • Earth-mounded concrete bunkers are a special case of below-ground vaults. This type of bunker is constructed partially above grade and then covered with soils or other materials to form a cap. Depending on the thickness of the cover, erosion may be a long-term consideration. Otherwise, source-term considerations are similar to those for below-ground vaults. This type of disposal facility has been used in the French tumulus design (Pittiglio and Shaffner, 1986; Van Kote, 1983), as well as at DOE facilities (ORNL, 1997a). • Mined cavities have been used for disposal of radioactive and hazardous chemical wastes in Germany and have been investigated in The Netherlands. This technology also has been selected by Sweden (Carlsson, 1993) and The Netherlands (van den Broeck and Moonen, 1994) for disposal of low- and intermediate-level radioactive wastes, and is being investigated for disposal of intermediate-level wastes in the United Kingdom (Beale and Mogg, 1993). The depth of such mined facilities provides improved isolation of waste from human intrusion and biointrusion by plants and animals. In addition to the generally important source-term issues for shallow-land disposal, a source-term analysis for disposal in mined cavities should consider releases through

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holes and shafts between a mine and the ground surface. Even though those openings will be sealed when a facility is closed, they may provide preferential transport pathways compared with the unaltered host medium. • Augured holes and shafts typically are several meters in diameter and tens of meters deep. This disposal concept has been used at DOE facilities and several places around the world (Trevorrow et al., 1985), and it is receiving increased attention as an option for disposal of spent sealed sources (IAEA, 2003). Advantages of this concept include reduced worker exposure, low susceptibility to erosion, minimization of biointrusion, and ease of fabrication and closure. Untreated waste (e.g., rubble, soil, equipment components) can be placed directly in a shaft or, if greater confinement is required, placed in a container prior to disposal. In some cases, shafts are lined to enhance potential retrievability and remediation procedures. Shafts can be closed by backfill with native soil or with clay or concrete caps to reduce water infiltration and release of gases (e.g., radon). This concept is similar to trench disposal, except the geometry of a disposal unit is different. Use of augured holes and shafts provides a simpler means of achieving greater depth of disposal for relatively small volumes of waste than deeper excavation of trenches (e.g., when deeper burial is desired to mitigate concerns about inadvertent human intrusion). • Trench disposal involves excavation of a region as large as 5 to 10 m deep, 20 to 30 m wide, and hundreds of meters long. Waste, with or without containers, is placed directly in a trench. A trench can be covered with a cap to divert water from the waste. At disposal sites where there is a deep water table, a trench can be located several meters below grade, which helps alleviate concerns about intrusion by humans, plants and animals. Trenches have been widely used for disposal of low-level wastes throughout the world. 5.4.4

Waste Containers

A wide variety of sizes, shapes and materials of construction have been used in containers for low-level waste. Past practices included use of glass bottles, cardboard and wooden boxes, polyethylene bags, and metal drums. More recently, improved waste containers, such as those described below, have been used. A review of shipping manifests for non-DOE low-level waste in the United States indicates that in the period 1987 to 1989, approximately 80,000 waste containers were disposed of annually.

156 / 5. PERFORMANCE ASSESSMENT MODELS Containers ranged in size from 0.00026 to 20 m3 (Roles, 1990). More than 75 % of all containers were 55-gallon (0.2 m3) carbon steel drums. Other containers that were used to meet specific needs of waste generators include carbon steel drums of various sizes (0.02 to 0.3 m3) and metal boxes (0.4 to 1.1 m3).30 However, more highly radioactive wastes (Class-B and -C waste) often are disposed of in high-integrity containers (HICs). Materials used to construct HICs include type-316 stainless steel, Ferallium (a duplex stainless steel), and HDPE. Because of concerns about long-term creep of HDPE, placement of this type of HIC in a steel or, more commonly, concrete overpack usually is required. Concrete overpacks also are used in many other countries (Hauser and Koester, 1989; Marque, 1994). A waste container provides a barrier to prevent water from contacting waste. An important consideration in modeling the source term is the time over which containers remain completely intact and isolate waste from water. In general, different container sizes, shapes and materials, as well as locally different environmental conditions, would require a separate analysis of each container. However, that task would be overwhelming. To simplify an analysis, containers normally are grouped into a few categories, including carbon steel drums, carbon steel liners, stainless steel HICs, Ferallium HICs, and HDPE HICs. Within each category, the performance of individual containers is assumed to be identical. Mechanisms of container failure may be categorized as catastrophic or local. In catastrophic failure, a container is assumed to no longer provide a barrier to water contacting waste. Catastrophic failure typically is described using either a time of failure specified by an analyst or a failure time calculated from the thickness of a container and an assumed corrosion rate. For example, the time to failure of a metal container can be calculated as: d t f = ------ , r where: tf = d = r =

30Large

(5.8)

time to failure (years) corrosion allowance thickness (meters) general corrosion rate (m y–1)

metal boxes also are used at some DOE low-level waste disposal sites (McDowell-Boyer et al., 2000; ORNL, 1997a).

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The corrosion allowance thickness (d) often is assumed to be less than the container thickness, to account for the likelihood of mechanical failure when the thickness of uncorroded metal is no longer sufficient to support the weight of overlying material. Localized failures of waste containers may result from such processes as stress corrosion cracking, weld failures, or pitting corrosion. The most common approach to modeling localized failure in metallic containers is to consider complete penetration of a wall by pitting. With localized failure, the container still provides a barrier to water flow, but the area penetrated by a pit allows water to contact waste. Localized failure thus results in release of radionuclides in the aqueous solution phase at earlier times than would be estimated by assuming a failure time based on a general corrosion rate. In addition, if a container is intended to provide a barrier to releases in the gas phase, such releases can occur when the container is first breached. Extensive data have been collected on corrosion of carbon steels and stainless steels in soil environments (Gerhold et al., 1981; Romanoff, 1957). Testing was conducted on bare samples (unpainted metal) in more than 100 soils over a period of 17 y. Those data often are used as a basis for estimating container lifetimes. A review of data collected by Romanoff (1957) indicates that 55-gallon carbon steel drums with a nominal wall thickness of 1.27 mm (50 mils) should last between 10 and 140 y, depending on the soil environment (Sullivan and Suen, 1989). However, carbon steel also is susceptible to pitting corrosion, and models of pitting corrosion indicate that the time of first breach should be on the order of a few years for wall thicknesses typical of 55-gallon drums (Mughabghab and Sullivan, 1989). For this reason, carbon steel drums often are assumed to provide no protection of a waste form. That is, drums are assumed to fail completely at the time a disposal facility is closed. Carbon steel liners are rectangular boxes that often are used to ship activated metals and other high-activity wastes. On the basis of their wall thickness, which typically is 3.2 mm or greater, this type of container often is assumed to protect wastes from contact by water for periods of 25 to 50 y. Data on pitting of carbon steels (Romanoff, 1957) suggest that such containers may experience complete wall penetration by pitting at times <25 y. In order to be approved for use as a HIC at NRC-licensed lowlevel waste disposal facilities, a container must be able to maintain a positive water seal for at least 300 y (NRC, 1991c). Studies have shown that for container lifetimes in excess of 200 y and less than

158 / 5. PERFORMANCE ASSESSMENT MODELS a few thousand years, projected maximum annual doses to off-site individuals are essentially independent of container lifetime (Kozak et al., 1993). For this reason, and because it is difficult to justify longer-term extrapolations on the basis of limited data that are available, HICs often are assumed to have a lifetime of 300 y. If more credit for the performance of a stainless steel container is desired, data of Gerhold et al. (1981) could be used to estimate a general corrosion rate. There are few data on corrosion properties of Ferallium in soil systems. However, that alloy was developed for use in extremely corrosive chemical environments, and its corrosion performance should be superior to that of stainless steels. Therefore, data of Gerhold et al. (1981) could be used to estimate container lifetimes. Mechanisms for degradation of the performance of HDPE have been reviewed by Cowgill (1992). Degradation mechanisms include corrosion, autocatalytic oxidation, low-creep rupture, environmental stress cracking, irradiation-induced embrittlement, biodegradation, and water uptake. Under typical disposal conditions, autocatalytic oxidation with biodegradation, radiation damage, and creep rupture appear to be the most likely mechanisms of failure (Cowgill, 1992). However, data are incomplete and most credible failure mechanisms cannot be ruled out. Although data indicate that the lifetime of HDPE containers should be well in excess of 300 y, data are sparse and, thus, there are large uncertainties in predicted lifetimes. HICs generally are designed with a passive gas vent to prevent pressurization of containers. Therefore, HICs do not provide a barrier to release of radionuclides in gaseous form. In most performance assessments, the performance of waste containers is evaluated by assuming that catastrophic failure of a container occurs at a time that is estimated on the basis of the thickness and material properties of a container wall. Since limited data are available for predicting container failure caused by localized failure mechanisms, those processes are ignored in most assessments. However, consideration should be given to the effect on the source term of container failure caused by local mechanisms that occur earlier than catastrophic failure.

5.4.5

Waste Forms

Low-level waste may be placed in a waste container without treatment, or it may be treated with sorbents or solidification

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agents, compacted or incinerated to reduce waste volume, dewatered, or surrounded with sand to minimize void space in a container. The term “waste form” refers to the form of contaminated material as placed in a waste container. If waste is untreated, the waste form can be identical to waste as generated. Low-level waste normally consists of a heterogeneous mixture of wastes. An appropriate conceptual model for release of radionuclides from waste packages will depend on the physical and chemical form of wastes. Furthermore, within each category listed in Table 5.1, wastes may have different physical and chemical forms. In solidified wastes, for example, solidification agents may include several different types of cement, bitumen and vinyl-ester styrene. Evaluating release of radionuclides from individual waste forms in a performance assessment would be impractical. Therefore, as in an analysis of container performance described in the previous section, waste forms generally are grouped into a few categories. Different waste forms within a category are assumed to experience the same release mechanism, although release rates could be different. For example, releases from all solidified waste forms are assumed to be controlled by diffusion, but the diffusion coefficient may vary for different materials as well as for different radionuclides within a single type of waste form. The first step in defining categories of waste forms to be used in performance assessment is to determine which waste forms contain most of the activity. Inventory data on shipping manifests for non-DOE waste (Table 5.1) indicate that most of the activity is found in activated metals, wastes solidified with cement, dewatered resins, dry solids (e.g., laboratory trash, papers, plastics, glassware), evaporator bottoms, filter sludges, and solid noncombustible materials (Cowgill and Sullivan, 1993; Roles, 1990). These seven waste-form categories contain more than 95 % of the total activity in non-DOE low-level waste. Even with an understanding of the most important waste forms, releases from a given type of waste form must be evaluated on a radionuclide-specific basis. The inventory of a radionuclide in waste forms in which it is found often differs significantly from the average inventory in all wastes. For example, all reported 232Th in non-DOE waste that was sent to a disposal site in Richland, Washington, during a single year was contained in sorbents (Sullivan and Cowgill, 1991). Thus, when estimating releases of any radionuclide, a detailed characterization of the inventory in terms of its distribution in various waste forms is needed for more accurate modeling of the source term.

160 / 5. PERFORMANCE ASSESSMENT MODELS 5.4.5.1 Waste-Form Performance: Aqueous Phase. In modeling waste-form performance, the first step involves obtaining information on distributions of radionuclide inventories in different waste forms and the distribution of waste forms in different waste containers. Information on containers is needed to estimate the time when water contacts waste and releases begin. Distributions of radionuclide inventories are needed to form a conceptual model of processes that will control release. Aqueous releases of radionuclides from a waste form frequently are assumed to occur as a result of one of the following four physicochemical processes: • • • •

surface rinse with partitioning diffusion dissolution solubility-limited release

The types of models that are frequently used to represent these processes in assessments of radionuclide source terms are described in the following sections.

5.4.5.1.1 Surface rinse with partitioning. A frequently-used model for surface rinse with partitioning assumes that the entire inventory of radionuclides is released immediately upon contact of a waste form with water, subject to equilibrium partitioning of radionuclides between the waste form and contacting solution. The partition coefficient, which is the ratio of the contaminant mass in the waste form per unit mass of the waste form to the concentration in solution at equilibrium, is used to define the distribution of a radionuclide in the waste form/solution system. In the extreme of a partition coefficient of zero, instantaneous release of the entire waste-form inventory upon contact with water is assumed. This assumption frequently is used to provide an upper bound on release or in situations where there are few data to support other assumptions about release mechanisms. Leaching data are lacking for many waste forms, including dry-active wastes, de-watered resins, filter sludges, and evaporator bottoms. Given an assumption of instantaneous release (i.e., a partition coefficient of zero), the total mass of a radionuclide released from a waste form is given by: ∞

M r = M 0 ∫ δ ( t – t f )e 0

– λt

dt ,

(5.9)

5.4 SOURCE TERM

where: Mr M0 δ tf λ

= =

= = =

/ 161

total mass of radionuclide released (kilograms) initial mass of radionuclide in waste form (kilograms) Dirac delta function time at which waste is first exposed to aqueous solution (seconds) radioactive decay constant (s–1).

The Dirac delta function, which represents an assumption of total instantaneous release at time tf, is defined by the relations: δ ( x ) = 0 if x ≠ 0,

(5.10)

∫ δ ( x ) dx = 1, where the region of integration includes the point x = 0. In Equation 5.9, the reduction in mass of a radionuclide in a waste form due to radioactive decay prior to the time of release is taken into account, but ingrowth due to decay of other radionuclides is not. When a radionuclide is produced by decay of another radionuclide in a waste form, the mass available for release at time tf should be obtained using the Bateman equations (Evans, 1955). If an analytical solution procedure is used, Equation 5.9 can be integrated directly. If the equation is solved numerically, the mass release rate during the time step when failure occurs is obtained by dividing the available mass at the time of failure by the numerical time-step size. When partitioning of a radionuclide between a waste form and contacting solution is assumed, a description of the release over time becomes more complex. Models that include a nonzero partition coefficient are characterized by a relatively large release at the time of failure of a waste container followed by a slower release over time until the inventory in a waste form is depleted. The initial release into an uncontaminated solution is such that the concentration in solution is in equilibrium with the concentration in the waste form, as defined by the partition coefficient. The slow release that follows is controlled by the rate at which a radionuclide is transported away from the waste form and the equilibrium partition coefficient. Therefore, the amount of mass released depends on the volume of the contacting solution, and the conceptual model for this type of release needs to include an approach to estimating the volume of the solution contacting the waste and the volume of the waste form. Sullivan and Suen (1991) have estimated the mass of a radionuclide released at time t when equilibrium partitioning occurs as:

162 / 5. PERFORMANCE ASSESSMENT MODELS Kp M a ( t ) – ⎛ ------⎞ M s ( t ) ⎝ γ⎠ M r ( t ) = --------------------------------------------------K 1 + -----pγ where: Mr = Ma = Kp

=

Ms = γ =

(5.11)

mass of radionuclide released (kilograms) mass of radionuclide available for release on waste form (kilograms) equilibrium waste form/solution partition coefficient for radionuclide (kg m–3) mass of radionuclide in contacting solution (kilograms) ratio of volume of water contacting waste form to mass of waste form (m3 kg–1)

When using a numerical technique (finite-difference or finiteelement) to calculate radionuclide transport away from a waste form, the volume of the contacting solution is selected to be the volume of the computational cell multiplied by the moisture content in that cell. The mass available for release is the mass remaining on the waste form after correcting for radioactive decay, ingrowth and releases at earlier times. As the partition coefficient in Equation 5.11 approaches zero, the entire mass is released instantaneously and the simple rinse-release model in Equation 5.9 is recovered. It also should be noted that there is a time when the release is zero (i.e., the system is at equilibrium) and a range of times when the release is negative (i.e., the waste form adsorbs contamination from solution). Release rates from a waste form are estimated by determining the mass released from Equation 5.11 and dividing by the computational time step. 5.4.5.1.2 Diffusion-controlled release. A model for diffusioncontrolled release assumes that radionuclides are initially distributed uniformly throughout a waste form and that the process that controls their release is diffusion through the waste form. The concentration at the edge of a waste form generally is assumed to be zero, which corresponds to a case where transport away from the waste form is much faster than diffusion within the waste form. This assumption gives the highest concentration gradient and, thus, the highest release rate. With these assumptions, the governing diffusion equation and initial and boundary conditions are given by:

5.4 SOURCE TERM

∂C(x,y,z,t) ---------------------------- = D∇ 2 C(x,y,z,t) – λC(x,y,z,t), ∂t

/ 163

(5.12)

C ( x,y,z,0 ) = C 0 C(x b ,y b ,z b ,t) = 0, where: C = D = λ = C0 = (xb,yb,zb) =

radionuclide concentration in waste form at coordinates (x,y,z) and time t (kg m–3) diffusion coefficient of radionuclide in waste form (m2 s–1) radioactive decay constant (s–1) initial concentration of radionuclide in waste form (kg m–3) coordinates of boundary of waste form

The diffusion equation is solved and the release per unit area per unit time (i.e., flux density) at the surface of a waste form [J (kg m–2 s–1)] is calculated as: J = – D∇C.

(5.13)

The flux density then is integrated over the surface area of the waste form to obtain the mass release per unit time (kg s–1). Analytical solutions of a diffusion-controlled release model have been developed for several waste-form geometries, including a semi-infinite medium and finite-sized spherical, cylindrical or rectangular source regions (ANSI/ANS, 1986; Pescatore, 1991). For a semi-infinite source region, the solution of Equation 5.13 in the presence of radioactive decay is: D 0.5 – λt J(t) = C 0 ⎛ -----⎞ e , ⎝ πt⎠

(5.14)

where: t

=

time since onset of a diffusion release

It should be noted that the flux density becomes infinite as t approaches zero. This behavior, which is characteristic of all diffusion models, results from a discontinuity between the concentration at the boundary of a waste form (C = 0) and the initial concentration immediately adjacent to the boundary (C = C0). To

164 / 5. PERFORMANCE ASSESSMENT MODELS circumvent this problem, Equation 5.14 may be integrated over time and evaluated at times t and t + ∆t. The integral of the flux density approaches zero as t approaches zero, and the release rate can be estimated from the difference in the calculated amount released at the two times and the magnitude of the time step (∆t). An assumption of a semi-infinite source region provides a reasonably accurate representation of release from a waste form when <20 % of the original mass would have been released in the absence of radioactive decay (ANSI/ANS, 1986). After this amount has been released, depletion of the inventory in a waste form causes the semi-infinite approximation to overpredict release rates. More accurate representations of diffusion-controlled release can be obtained by taking into account the finite size of waste forms. Analytical solutions of Equation 5.12 for finite waste forms are expressed as infinite series (Pescatore, 1991). Series solutions converge slowly at small times, and special techniques are used to enhance the convergence rate of solutions. Diffusion models provide a time-dependent release rate that is independent of processes that occur outside a waste form. Other models that incorporate solution feedback effects have been developed (Sullivan and Suen, 1991). In practice, however, radionuclide concentrations in a waste form generally are much greater than in a contacting solution, and solution feedback effects thus are unimportant. Those effects become important only when radionuclide concentrations in solution approach solubility limits, but a solubility-limited release model is more appropriate in that case (Section 5.4.5.1.4). A diffusion model is used most often for wastes that are solidified in a binder material, such as cement or vinyl-ester styrene. The diffusion coefficient should be determined experimentally on a radionuclide-specific basis for the particular combination of waste and binder used for disposal. If relevant experimental data are not available, the diffusion coefficient often is assumed to be the maximum value that has been approved for an acceptable solidification agent by NRC (1991c). That value, which is D = 10–10 m2 s–1, probably provides a considerable overestimate of diffusion from intact binder materials for most radionuclides. Predicted diffusion-controlled cumulative fractional releases of a contaminant over time from a waste form with dimensions of a 55-gallon (0.2 m3) drum (i.e., radius of 0.28 m and height of 0.85 m) are shown in Figure 5.14 for three diffusion coefficients (D = 10–10, 10–12, and 10–14 m2 s–1). For larger waste forms, cumulative fractional releases are lower for a fixed diffusion coefficient. For the largest diffusion coefficient of 10–10 m2 s–1, more than 40 % of

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Fig. 5.14. Calculated cumulative fractional release of contaminants from cylindrical waste form of radius 0.28 m and height 0.85 m assuming diffusion-controlled release and different diffusion coefficients.

the initial inventory is released within the first year and the release is essentially complete within 20 y. For the smallest diffusion coefficient of 10–14 m2 s–1, the release is spread out over tens of thousands of years. In that case, an analyst must demonstrate that the waste form will perform at least as well as the modeled result for extremely long time periods. 5.4.5.1.3 Dissolution (constant) release. The dissolution-release model assumes congruent release of all radionuclides in a waste form at a rate that is controlled by the dissolution velocity of the waste form. This model often is applied to releases from activated metals and glass waste forms. The dissolution velocity normally is assumed to be constant, and its value is based on experimental data for metallic corrosion or glass dissolution, as appropriate. Given the geometry of a waste form and an assumption of a constant dissolution velocity, the cumulative mass release of a radionuclide over a time t following onset of congruent dissolution is given by: S M r ( t ) = u ∫ M a ( τ ) ---- dτ, V

(5.15)

166 / 5. PERFORMANCE ASSESSMENT MODELS where: τ Mr u Ma

= = = =

S V

= =

dummy variable of integration total mass of radionuclide released (kilograms) dissolution velocity of waste form (m s–1) mass of radionuclide available for congruent dissolution (kilograms) surface area of waste form (m2) volume of waste form (m3).

The release rate from a waste form [Rw (kg s–1)] at time t is the time-derivative of the cumulative release in Equation 5.15, or: S R w ( t ) = u ⎛ ----⎞ M a ( t ). ⎝ V⎠

(5.16)

Many performance assessment models simulate release from a waste form using a constant fractional release rate. From Equation 5.16, that release rate is given by u(S/V). This model can be generalized to permit a radionuclide-specific release rate by defining a unique dissolution velocity (or fractional release rate) for each radionuclide. 5.4.5.1.4 Solubility-limited release. In a solubility-limited release model, an instantaneous release of radionuclides into solution is assumed to occur until the solubility limit in the fluid contacting a waste form is reached, and further releases are controlled by the rate of migration of radionuclides away from the waste form. A balance is achieved between the release rate and migration rate such that the concentration in solution at the edge of the waste form remains at the solubility limit. Therefore, there is a strong coupling in this model between transport away from a waste form and the release rate from the waste form. The solubility-limited release model can be applied to specific radionuclides in all waste forms. The solubility limit in the fluid contacting a waste form frequently controls the rate of release when a radionuclide has a low solubility limit in groundwater (e.g., uranium and other actinides). A solubility-limited release model often is applied in conjunction with models of other release mechanisms. Other mechanisms are assumed to control release unless a solubility limit is exceeded, in which case precipitation is assumed to occur such that the concentration in solution is equal to the solubility limit. The precipitated mass of a contaminant will begin to dissolve when transport mechanisms carry the contaminant in solution away from a waste form and the solute concentration falls below the solubility limit.

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A solubility-limited release model can be difficult to apply to chemical elements that have several radioactive isotopes (e.g., uranium). Care must be taken to insure that a solubility limit is applied to the total mass of all radioisotopes of an element of concern. 5.4.5.2 Waste-Form Performance: Gas Phase. Under certain environmental conditions, 3H, 14C, 85Kr, 129I, and 222Rn can exist in the gas phase. These radionuclides may be produced in gaseous form in a low-level waste disposal facility by production of hydrogen in anaerobic corrosion of metals, production of methane and carbon dioxide in microbial degradation of organic materials, failure of sealed sources of 85Kr, radiolysis to produce gaseous 3H and 129I, radioactive decay of 226Ra to produce 222Rn, and volatilization of liquid 3H. Formation of radionuclides in gaseous form may lead to releases to the environment through gas-phase diffusion and advective transport. Estimation of releases of gaseous radionuclides from a waste form requires consideration of each process listed above. Estimation of releases of 14C in carbon dioxide or 14C and 3H in methane as a result of biodegradation requires a detailed analysis of the types and amounts of organic materials present in a disposal facility and the rate of production of these gases in those materials. An analysis of this process that assumed current amounts of biodegradable material in waste has indicated that ~5 to 10 % of the inventory of 14 C could be released in the gas phase (Yim, 1994). Smaller fractional amounts of tritiated gas are expected to be released. Releases from sealed sources usually are described by assuming a time-to-failure of a seal followed by total release of the radionuclide inventory at that time. Radiolysis has been found to produce limited amounts of gas and can be neglected as a mechanism for gas release even at the maximum radiation levels expected in low-level waste forms, which are absorbed doses of <106 Gy (Yim, 1994). Release of 222Rn can be calculated from the decay rate of 226Ra and knowledge of the waste form. Release of radon from a surface contaminated with radium would occur immediately upon decay. In solidified wastes, however, gaseous 222Rn would have to diffuse through a waste form. Diffusion of gases through solidified waste forms is likely to be much more rapid than diffusion of liquids. 5.4.5.3 Ingrowth of Radionuclides. Decay of many radionuclides produces other radionuclides. These decay products generally will have different physical and chemical properties and, thus, different release and transport characteristics. Radioactive decay products

168 / 5. PERFORMANCE ASSESSMENT MODELS generally must be addressed in performance assessment. For example, decay products of isotopes of thorium often are more important than the parents, due to their greater solubility and mobility. This is particularly the case in a decay chain that includes 230 Th, 226Ra, and 222Rn gas. Another important consideration is the long time period that may be required to reach activity equilibrium in a decay chain. In the 238U decay chain, for example, the activity of 226Ra and its shorter-lived decay products, including 222Rn, increases for more than 106 y. Several models have been developed to account for ingrowth of radioactive decay products during transport outside a waste form. However, few models account for ingrowth in a waste form prior to release. Models to address ingrowth prior to release have been developed by MacKinnon and Sullivan (1994) and Sullivan et al. (1996). 5.4.6

Transport in Disposal Facility

Following release from a waste form, radionuclides will migrate out of a disposal facility. Processes that influence movement of radionuclides within a facility include advection, dispersion, diffusion, sorption and radioactive decay. Those processes apply to transport in aqueous and gaseous phases. Water is expected to flow through the cover and any other engineered barriers in a disposal facility, and water that enters a disposal cell will flow around waste. Most disposal facilities contain many different materials and obstructions to flow. Those heterogeneities, along with the potential for such localized effects as cracking of an engineered barrier, may result in highly nonuniform flow through a facility. In performance assessment, however, the flow field through a facility normally is assumed to be constant over specified time periods and spatially uniform at all times. 5.4.6.1 Aqueous-Phase Transport. The simplest approach to modeling releases from a disposal facility in the aqueous phase is to assume that once a radionuclide is released from a waste form, it also is released from the facility. This method is appealing because it is simple and inherently conservative, and because it bypasses the need to consider transport within a disposal facility. However, this approach will predict earlier releases from a facility than methods that consider transport within the facility and, thus, may substantially overestimate release rates. For example, consider a case where the average time for a nonsorbing radionuclide to move through a disposal facility is 10 y and the inventory of a radionuclide is assumed to be uniformly distributed throughout the facility

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and released immediately following failure of a waste container. If release from a waste form is assumed to be equivalent to release from the facility, all of the inventory would be released instantaneously. However, accounting for transport through the facility would spread the release over the 10 y transport time, thus reducing the peak release rate by a factor of 10 compared with assuming an instantaneous release. In the next level of modeling complexity, transport of released radionuclides through a disposal facility is simulated by considering advection but ignoring dispersion and diffusion. In this type of model, a disposal facility is divided into a number of mixing cells. A mass balance calculation is performed for each cell, and movement between adjacent cells is estimated by taking advection, sorption and decay into account (Kozak et al., 1990). Under appropriate conditions for the release rate from a waste form (i.e., rinse release with partitioning or a uniform release over time), analytical solutions can be obtained for an arbitrary distribution of sources in a disposal facility (Sullivan and Suen, 1991). Diffusion may be an important transport process within a disposal facility when engineered barriers are performing as designed and advection is very low. Dispersion during transport, especially longitudinal dispersion, may be important when the radionuclide travel time to a receptor location is comparable to its half-life, because this process tends to cause spreading of contamination in space, which leads to an earlier arrival of some radionuclides at the receptor location and in correspondingly greater amounts. When dispersion and diffusion are important, solution of the advective-diffusion transport equation is required (Section 5.7). A one-dimensional transport model that assumes spatially uniform flow through a disposal facility often is used in performance assessment. When localized effects are important (e.g., flow around containers or infiltration barriers, flow through cracks in grout backfill and engineered barriers), two- or three-dimensional models may be necessary. However, more detailed models may not lend themselves to evaluating thousands of different cases in simulating releases over long time periods. In modeling long-term releases, multidimensional models may be used to provide justification for selection of a uniform flow rate that bounds nonuniform flow effects for use in a one-dimensional model. Short-term transient effects on flow generally are not taken into account in modeling radionuclide transport in a disposal facility, but long-term alterations in flow rates often are considered by using step or ramp changes. Those changes are used to represent degradation of engineered barriers over time. Releases from a

170 / 5. PERFORMANCE ASSESSMENT MODELS facility are quite sensitive to flow rate and changes in flow rate. A ramp change slowly alters the flow rate and, thus, spreads a release out in time. Step changes in flow rate, which often are assumed to be as large as an order of magnitude, lead to predictions of large changes in release rates. In an alternative conceptualization of changes in flow rates, a disposal cell is assumed to have two flow regions. The first region receives water at a rate determined by the performance of intact barriers to flow. In applications of this type of model, radionuclide releases in this region often are assumed to be diffusion-controlled and the flow velocity is assumed to be zero (Selandar et al., 1991). The second region receives water at a rate determined by an absence of any barriers to flow. Fractions of a disposal cell that are assigned to the two regions then are assumed to change in time to simulate degradation of engineered barriers. 5.4.6.2 Gas-Phase Transport. Similar to the treatment of aqueous-phase transport described in the previous section, the simplest approach to modeling transport of radionuclides in the gas phase within a disposal facility is to assume that a gas is released immediately to the atmosphere above the facility once it is formed. If a more detailed analysis is required, transport in the gas phase will depend on the design of a disposal facility and the condition of engineered barriers. If a facility remains intact and fractures are not present, effects of changes in barometric pressure are expected to be small and diffusion should be the dominant transport process. In this case, an analysis would involve estimating the rate of diffusion of gases through engineered barriers to the ground surface. Monitoring wells, which are incorporated in the design of many disposal facilities, can provide a direct pathway for release of radionuclides in gaseous form. Releases will be caused by barometric pumping effects and will occur in discrete pulses that correspond to times when the ambient pressure is less than the gas pressure inside a facility. Under those conditions, the transport time through a facility will be <1 y and releases will be rapid (Yim, 1994). When monitoring wells are located outside a disposal facility, gases must migrate through the facility and engineered barriers to reach the surface. This process also will be influenced by barometric pumping, albeit to a lesser extent than in cases where monitoring wells are located within a facility. For example, transport of radon through cement slabs in basements often is dominated by barometric pumping effects (Nazaroff, 1992). Therefore, if fractures form in engineered barriers, their influence on advective transport of gases may need to be analyzed.

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In general, processes of advection, dispersion, diffusion, sorption and decay will control movement of gases within a disposal facility. However, when barometric pumping effects are important, it may not be reasonable to perform a steady-state transport analysis. 5.4.7

Interfaces with Other Performance Assessment Models

Modeling of the source term is one component of a performance assessment (Figures 5.3 and 5.4). The source-term model is directly coupled to a number of other models used in an analysis. Important interfaces between the source-term model and models of other processes are summarized as follows. • Models of aqueous infiltration provide water flow rates that are used as input to a model of the source term. In some cases, the flow rate is calculated as part of the source-term model (i.e., the model calculates flow and transport simultaneously). • Gas flow rates must be obtained using appropriate models for a disposal facility. For example, barometric pumping through monitoring wells must be considered in some cases. • Chemical controls on radionuclide releases may be provided by engineered barriers in a disposal facility and in waste containers and waste forms. Values of parameters that depend on chemical conditions (e.g., solid/solution partition coefficients, solubility limits) may be obtained from geochemical speciation codes or experimental data. The latter should be site-specific. • Release rates of radionuclides in the aqueous phase obtained from a source-term model provide input to models of transport in water (vadose zone and aquifer). • Release rates of radionuclides in the gas phase provide input to models of atmospheric transport. 5.4.8

Source-Term Issues

A number of issues need to be considered when developing conceptual models to represent release of radionuclides from a waste form and their subsequent transport from a disposal facility. These issues are concerned with descriptions of radionuclide inventories, waste containers, waste forms, and radionuclide transport. Issues in describing infiltration, cover performance, and concrete degradation are discussed in Sections 5.2 and 5.3.

172 / 5. PERFORMANCE ASSESSMENT MODELS An issue common to all aspects of an assessment of the source term is specification of initial conditions for an analysis. Performance assessment is concerned with releases during the postclosure time period of a disposal facility, but initial conditions for an assessment are determined during the preclosure period. In particular, processes that degrade engineered barriers, waste containers, and waste forms can occur throughout the period prior to facility closure. Thus, consideration should be given to processes that occur during the preclosure period that can alter performance in the postclosure period, including radioactive decay. Inclusion of radioactive decay requires information on the time history of waste emplacements, unless a conservative assumption is made that all disposals occur at the time of facility closure. Another issue common to all aspects of a source-term analysis is treatment of multiple sources. In conducting a performance assessment, a disposal facility frequently is treated as a single, homogenized source that is intended to represent average waste in the facility. However, each disposal cell in a facility can act as a unique source. If it is believed necessary to treat each cell individually, a source-term analysis would be performed for waste containers and waste forms in each cell. The predicted release rate from each cell would act as a source for transport at discrete spatial locations, and a transport calculation would consider migration of radionuclides from multiple sources to a receptor location. Modeling of each disposal cell individually also allows consideration of multiple closure times to represent closure of each cell, as opposed to a single time for closure of an entire disposal facility. Treatment of each disposal cell as a unique source greatly complicates an analysis of the source term and should be considered only as necessary in demonstrating compliance with regulatory criteria. An assumption of a single, representative disposal cell is much easier to visualize and analyze. Through proper choice of performance parameters, it should be possible to justify an analysis of a representative cell for most disposal facilities. 5.4.8.1 Radionuclide Inventory Issues. Two issues that involve radionuclide inventories are discussed in this Section. The first is use of a unit source term in assessing the performance of a disposal facility, and the second is the problem of inaccurate estimates of inventories. 5.4.8.1.1 Unit source term. Many source-term models assume that the release rate of a radionuclide is directly proportional to its inventory in a disposal facility. This assumption permits an analysis

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of the source term for a unit inventory. This approach is particularly useful when an analysis is intended to be used in establishing waste acceptance criteria for a disposal facility in the form of limits on inventories of radionuclides. A unit source-term approach often is used when data needed to predict the source term, especially radionuclide inventories, are lacking. Any conceptual model that assumes that release and transport from a waste form are independent of radionuclide concentrations in waste will lead to a linear response to changes in inventory (i.e., the source term will be proportional to the inventory). However, this condition will not be met if release is controlled by solubility, or if sorption processes depend on concentrations of radionuclides in solution (i.e., if sorption is not proportional to concentrations in solution). An assumption of nonlinear sorption generally is not used in performance assessment, but solubility limits of radionuclides often are assumed in source-term analyses. The nonlinear dependence of the source term on solubility makes it difficult to determine waste acceptance criteria when radionuclide releases are controlled by solubility. In such cases, however, considerations of protection of inadvertent intruders (Section 6) usually are more restrictive in determining disposal limits. 5.4.8.1.2 Inaccurate estimation of inventories. Waste generators are required to estimate quantities (activities) of individual radionuclides in their wastes. Many wastes contain several potentially important radionuclides with inventories that differ by orders of magnitude. Measurement of quantities of some radionuclides with relatively low activity levels is a difficult problem. A common approach is to use scaling factors, which assume that the quantity of an unanalyzed radionuclide is a multiple of a measured quantity of an easily detectable radionuclide, such as 137Cs or 60Co. Lower limits of detection, rather than measured values, may be reported as estimated inventories, especially when radionuclides do not influence waste classification and there is no financial or regulatory incentive to determine inventories more accurately. In non-DOE low-level waste, radionuclides that usually result in the highest projected doses to off-site individuals, such as 14C, 129I, and 99Tc, usually have much lower activities than other radionuclides and are difficult to detect because their primary emissions are low-energy beta particles. Scaling factors for those radionuclides that were developed for use in the nuclear power industry beginning in the mid-1980s were believed to overestimate actual quantities in low-level waste from power reactors by as much as a few orders of magnitude (Robertson et al., 2000). Substantial overestimation of

174 / 5. PERFORMANCE ASSESSMENT MODELS activities by waste generators continued because it did not impact disposal costs, requirements on waste transportation, or classification of wastes as Class A, B or C (Section 3.4.2.1) and, thus, there was no advantage to generators in reporting lower activities. However, performance assessments of low-level waste disposal facilities have indicated that assumed inventories of these radionuclides, especially 129I, could result in projected doses to the public that exceed performance objectives in 10 CFR Part 61 (Table 3.2) at some sites (NRC, 1982b; 1993). In response, efforts were undertaken to develop more realistic scaling factors (Robertson et al., 2000). That work also included development of scaling factors for other potentially important radionuclides that are not included in NRC’s waste classification system in 10 CFR Part 61 (Table 3.3), including 10Be, 36Cl, 93Mo, 93mNb, 108mAg, 113mCd, and 121mSn. Reporting of radionuclide inventories on the basis of lower limits of detection, which was incorporated, for example, in early scaling factors for 99Tc and 129I (Robertson et al., 2000), generally can result in large overestimates of inventories of radionuclides that commonly occur only in trace amounts in low-level waste. Since many thousands of waste containers may be sent to a waste disposal facility, a claim that there is a small quantity of important radionuclides in each container that is estimated on the basis of lower limits of detection can lead to erroneous estimates of significant quantities in an entire facility. If this is a potential problem at a site, use of laboratory analytical methods on selected wastes may be required. Reported inventories of uranium and thorium in non-DOE low-level waste also are believed to be exaggerated (Dehmel et al., 1994; DOE, 1993; Roles, 1990). Substantial overestimates of inventories can occur when a measured specific activity in waste is applied to the total mass of a waste form, rather than to a much smaller mass of uranium or thorium in a waste form. In addition, an inventory of thorium can be overestimated when a measured activity of 232Th and its decay products is assigned to 232Th only. Even though overestimation of inventories of uranium and thorium does not affect classification of non-DOE low-level waste or disposal costs and, thus, there is no advantage to generators in performing a more accurate analysis, excessive pessimism should be avoided because of potential impacts of unreasonably high inventories on projected doses to off-site members of the public or inadvertent intruders. 5.4.8.2 Waste-Container Issues. Two issues that involve the performance of waste containers are discussed in this Section. The first is insufficient data to characterize containers, and the second is use of a time-distributed failure function for container performance.

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5.4.8.2.1 Insufficient characterization of containers. Data on waste containers for non-DOE low-level waste normally are limited to the number of containers, their volume, and a designation as to the type of container (drum, box, cask liner, or other). Since data on materials of construction and wall thickness are not reported, it is difficult to estimate container lifetimes. In some cases, needed information can be inferred from reported data. For example, since the volume of a 55-gallon (0.2 m3) drum is known, waste containers with the same volume could be assumed to be a carbon steel drum with known wall thickness. In addition, only ~15 % of the data on containers for non-DOE low-level waste has included information on the distribution of radionuclides by container (Dehmel et al., 1994). The remaining 85 % of the data has provided this distribution only for an entire shipment of waste containers. This type of data does not allow an accurate assignment of radionuclides to a particular waste container or waste form. 5.4.8.2.2 Distributed failure of containers. Many models of container failure used in performance assessment assume that all containers fail simultaneously. This assumption generally results in predicted release rates of radionuclides that are more localized in time and are greater than would be predicted if containers were assumed to fail over a period of time. Similar containers should fail over a period of time, rather than all at the same time. Attempts have been made to account for distributed failures when modeling the source term (Dolinar et al., 1994). However, it is difficult to defend an assumed failure rate over time. If the actual failure rate exceeds an assumed value for only a brief period (a few years), the maximum individual dose at locations beyond a facility boundary could be underpredicted. For example, an assumption that containers fail at a uniform rate for 100 y starting at 200 y after facility closure can, in some instances, result in a lower projected dose than an assumption that all containers maintain their integrity for 300 y and only 10 % of the containers fail at that time. 5.4.8.3 Waste-Form Issues. Issues of concern to the performance of waste forms discussed in this Section include changes in waste types and characteristics over the operational life of a disposal facility, insufficient characterization of waste, insufficient data on release rates from waste forms, an assumption of homogenization of radionuclide inventories over an entire disposal facility, and the role of geochemistry in determining releases from waste forms.

176 / 5. PERFORMANCE ASSESSMENT MODELS 5.4.8.3.1 Changes in waste types and characteristics. As practices that generate low-level waste and methods of waste management evolve over time, significant changes in waste types and characteristics, including radionuclide inventories, are likely. Radionuclide inventories at the end of the period of waste emplacement at a disposal site of ~20 to 30 y may differ significantly from initial projections. For example, there has been an increasing tendency to dewater ion-exchange resins and place them in HICs, instead of using solidification agents, and large volumes of wastes from decontamination and decommissioning activities and cleanup of old waste disposal sites will be produced in the future. As such changes occur, source-term modeling may need to be repeated to better represent the distribution of wastes in a disposal facility at the time of closure. 5.4.8.3.2 Insufficient waste-form characterization. In the past, disposed wastes have been identified on waste manifests using such descriptors as “mixed fission products,” “uranium wastes,” or “thorium wastes.” These types of labels do not provide specific information about radionuclide inventories or the physical and chemical characteristics of wastes. Even today, with better record keeping, it often is not possible to correlate the inventory distribution of a specific radionuclide with various waste forms. Information on the chemical form of radionuclides in waste is almost always lacking. If more accurate and defensible source-term modeling is required, better characterization of physical and chemical forms of wastes is needed. Alternatively, facility designs could be enhanced to provide better performance (e.g., sorbent barriers could be added). Current information on waste characteristics generally is insufficient to support realistic estimates of release rates. Information on physical and chemical characteristics of waste, such as the type of activated metal, composition of dry solids, and types of cements used to solidify particular wastes, often is lacking. Some of this information can be inferred; for example, activated metals often are stainless steel reactor components, and by determining the source of components, information on the type of stainless steel could be obtained. In most cases, however, physical and chemical compositions of waste forms are not identified. Lack of information on physical and chemical characteristics of waste forms often leads to use of conservative release rates of radionuclides in performance assessment. In many cases, essentially no credit is given to a waste form (e.g., as in the surface rinse model discussed in Section 5.4.5.1.1). In such cases, water flow and chemistry (partition coefficients and solubility limits) are relied upon to control releases from a facility.

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5.4.8.3.3 Insufficient data on release rates. Release rates of radionuclides from each waste form will depend on physical and chemical properties of waste and the chemical environment near the waste form, which generally is site-specific. Release rates from apparently similar waste forms (e.g., wastes solidified using different formulations of cement) may differ substantially (Sullivan and Suen, 1989). Also, releases from identical waste forms would be expected to differ in different chemical environments. At the present time, data on release of radionuclides from most of the important waste forms (e.g., activated metals, dewatered resins, dry-active wastes, filter media, evaporator bottoms) are sparse. There are some data on releases from solidified waste forms, but measurements have tended to focus on high-activity, shorter-lived radionuclides that generally are unimportant in regard to projected doses to the public resulting from release and transport in groundwater (e.g., 137Cs, 90Sr, and 60Co). Also, experiments often are conducted under controlled conditions that do not necessarily represent conditions in a disposal facility. In the absence of defensible data, a source-term model that gives the highest release rates must be used in performance assessment (e.g., instantaneous release of an entire inventory of radionuclides from a waste form following container failure). Obtaining defensible data requires knowledge of waste forms and chemical conditions (solubility limits and partition coefficients) at a disposal site. Literature values developed for other applications under different environmental conditions cannot be used without sufficient justification that they are appropriate to a disposal site of concern. 5.4.8.3.4 Homogeneity of wastes. A common assumption in modeling the source term is that all radionuclides are distributed uniformly throughout a disposal facility. This condition never occurs in practice, and there are always regions where concentrations are higher or lower than an average value. This variability is a particular concern for important radionuclides that are present only in relatively small amounts (e.g., 129I, 99Tc). Such radionuclides may occur in only a few waste packages, and their placement in a facility will be characterized by zero concentration over most of the facility with a few discrete regions having nonzero concentrations. In one case, 129I was reported in only 79 of 649 shipments to a disposal site, and 25 of the shipments contained more than 90 % of the reported activity (Dehmel et al., 1994). Thus, if 129I is assumed to be uniformly distributed within a single shipment, homogenization of the activity over an entire facility could reduce peak concentrations in releases by a factor of ~25 (649 shipments divided by

178 / 5. PERFORMANCE ASSESSMENT MODELS 25 shipments containing 129I). Depending on a number of factors, including distance to a receptor location, the flow regime in a disposal facility and the local environment, and chemical effects, the concentration at a receptor location could exceed a projected value that is based on an assumed homogenization of activity in a facility. Another cause of inhomogeneity in radionuclide inventories in disposal facilities is radioactive decay during the operating period of a facility. Over the period of waste emplacement of ~20 to 30 y, shorter-lived radionuclides may undergo substantial decay. Therefore, even if the same inventory of those radionuclides is placed in a facility during each year of operation, the concentration at the start of the postclosure period will not be uniform and will depend on the spatial pattern of waste emplacements over time. 5.4.8.3.5 Issues of geochemistry and solubility. The particular chemical form (and phase) of a radionuclide directly influences its solubility and, therefore, the potential for release. For example, if a soluble species exists in a dry-active waste form as a solid phase simply because there is no free water present, its release by dissolution in infiltrating water will likely be rapid once water contacts the waste form. However, if a species is insoluble in water, its release will take place only after chemical reactions occur that either transform the radionuclide to a more soluble form or modify chemical conditions so that the radionuclide becomes soluble. Single values of solubility limits of chemical elements generally are used in source-term modeling. In actuality, however, the solubility limit of a particular element will vary in time and space due to changes in the chemical environment. For example, if chelating agents are released from a waste form, radionuclides can form chemical complexes with them, thereby substantially increasing solubility. When a single solubility limit is used, an analyst must demonstrate that the assumed value is an appropriate representation of the actual solubility in a disposal facility (e.g., that an assumed solubility limit is a conservative representation of likely values). Another complication arises with the formation of colloids, which are suspended solids that often incorporate radionuclides from solution. In this case, an apparent solubility, which is the concentration of a radionuclide incorporated into colloids plus a thermodynamically defined solubility in the aqueous phase, will exceed the thermodynamic solubility. Therefore, an assumption of solubility-limited release on the basis of aqueous-phase solubility would not adequately represent a release of radionuclides. If colloids are present and could incorporate substantial fractions of the

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inventory of a radionuclide, they must be treated as a separate phase in an analysis of release and transport. However, colloids generally are immobile in soil systems and are not considered in most performance assessments (see also Section 5.4.8.4.5). 5.4.8.4 Issues of Radionuclide Transport. Issues that involve transport of radionuclides through a disposal facility discussed in this Section include use of steady-state and uniform flow-field approximations and the role of geochemistry in determining transport. 5.4.8.4.1 Steady-state flow. The time frame for performance assessment (typically 1,000 y or more) often leads to an assumption of steady-state flow of water throughout a disposal facility. A justification for this assumption is that seasonal fluctuations in flow should be less important, in the long term, than an average flow over long time periods (Kozak et al., 1990). Numerical modeling studies have indicated that this should be the case for nonreactive transport in a spatially uniform flow field (Duguid and Reeves, 1977; Wierenga, 1977). However, sorption processes often exhibit hysteresis and transitory effects, which can influence long-term transport in a nonlinear manner (Jones and Watson, 1987). In addition, time-dependent flow rates may need to be used if there are localized effects that favor movement of radionuclides over large distances in short periods of time as a result of transitory flow events. In a disposal vault, for example, localized short-term effects may result from formation of channels as a consequence of improper placement of backfill, subsidence of wastes, or localized fracturing of engineered barriers. If steady-state flow rates throughout a disposal facility are used to estimate the source term, an analyst must demonstrate that selected flow rates lead to projected release rates that are reasonable bounds of release rates that could result from transitory flow events. 5.4.8.4.2 Uniform flow fields. The velocity field of water through a disposal facility often is assumed to be spatially uniform. Even in more sophisticated analyses that account for flow effects in two or three dimensions, spatially uniform flow normally is assumed. In many natural systems, however, local heterogeneities permit flow to proceed at rates far above average (i.e., preferential flow and fingering phenomena can occur). Various containers in a disposal facility provide large-scale heterogeneities such that flow properties and, therefore, flow through a facility will not be uniform. An analyst should justify that the flow field selected for use in an assessment adequately accounts for nonuniform flow effects.

180 / 5. PERFORMANCE ASSESSMENT MODELS 5.4.8.4.3 Role of geochemistry in transport. Because wastes often are poorly characterized, waste forms usually are given little credit in controlling releases of radionuclides from a disposal facility. Instead, reliance is placed on geochemical processes (solubility and sorption) to control releases from waste forms and transport from a facility. However, geochemical processes of concern may be difficult to describe realistically. Adsorption/desorption reactions depend on the chemical forms of radionuclides. Those reactions also depend on solution chemistry (Eh, pH, presence of competing ions, and presence of available sorption sites), the distribution of adsorbents on soil mineral surfaces, and concentrations of compatible adsorbate species in solution (Serne et al., 1990). Changing environmental conditions along a flow path can substantially alter solubilities and sorption of radionuclides. The most important parameters that determine the chemical form and distribution of a radionuclide among its various phases include: (1) concentrations of chemical constituents, especially complexforming ligands, in solution and on soil grain surfaces; and (2) prevailing pH, redox potential (Eh), and temperature conditions (Serne et al., 1990). Redox conditions are particularly important because many radionuclides, such as isotopes of plutonium, uranium and technetium, have several redox states, and their speciation (and, therefore, solubility and sorption) depends on redox conditions. Effects of pH on solubility and sorption also are important in determining releases of radionuclides, and many radionuclides are susceptible to a change in speciation when the pH changes. For example, many metals tend to form carbonate precipitates at higher pH values. The capacity to adsorb radionuclides also may be influenced by the pH. In most performance assessments, the role of sorption in radionuclide release and transport is described by a single parameter, the solid/solution distribution coefficient (Kd), and solubility is described by a single solubility limit. An analyst must insure that those parameters adequately represent the geochemical behavior of radionuclides under a range of conditions that will occur after waste disposal. Proper justification will require site-specific experimental data, more detailed geochemical modeling, or documented scientific judgment. The issue of uncertainty in chemical parameters can be addressed formally by defining ranges of values that can be used to generate distributions of calculated radionuclide release rates.

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5.4.8.4.4 Role of microbial processes. Microbially-mediated processes, including generation of byproducts by microorganisms, can substantially alter the geochemical environment and affect releases of radionuclides from a disposal facility. Bacterial catalysts can enhance actions of organic reducing agents, and the pH can be lowered as concentrations of CO2 and H2S increase, particularly under stagnant or ponding-water conditions (Manahan, 1991; Stumm and Morgan, 1981). Microbially-produced organic chelates can significantly enhance the mobility of some heavy metals by forming soluble complexes that do not readily adsorb (Francis and Dodge, 1993; Manahan, 1991; NAS/NRC, 1993; Spor et al., 1993; Toste et al., 1984). Some heavy metals can adsorb directly onto microorganisms and thus be immobilized (NAS/NRC, 1993; Spor et al., 1993). Most performance assessments have not considered effects of microbially-mediated processes on radionuclide release and transport. However, an analyst may need to consider such effects in environments that may occur in a disposal facility after closure. The potential importance of microbial processes should be reflected in the selection of chemical parameters (distribution coefficients and solubilities) used in a source-term analysis. 5.4.8.4.5 Role of colloids. Radionuclide migration may be significantly enhanced by the presence of mobile colloids. Colloids are particles with diameters <10 µm that are suspended in solution, and they commonly occur in natural and engineered soil and water systems. Colloidal particles include microorganisms, humic substances, mineral precipitates, clay minerals, and iron oxides (McCarthy and Zachara, 1989). Potential sources of colloids in a disposal facility include: (1) corrosion of metal containers, which can give rise to colloids in the form of amorphous iron oxides, and (2) degradation of cementitious waste forms and concrete barriers, which may lead to formation of colloidal-sized calcium carbonate particles. Colloids have high surface areas per unit volume, which can make them highly reactive adsorbents for radionuclides. As a result, colloids can substantially enhance migration of radionuclides that might otherwise adsorb to the host medium. However, colloids may be removed from suspension during transport by mechanical filtration, adsorption onto soil, and neutralization of their surface charge by acid-base and complexation reactions, thus allowing them to coagulate. Those processes frequently prevent substantial movement of colloids. Therefore, colloid-facilitated transport generally is believed to be unimportant and is usually

182 / 5. PERFORMANCE ASSESSMENT MODELS neglected in performance assessment. The significance of colloid-facilitated transport is discussed by McCarthy and Zachara (1989) and Mills et al. (1991). 5.4.9

Summary

The source term is defined as the rate of release of radionuclides from a disposal facility. The source term for any radionuclide is a function of time after disposal. An analysis of the source term requires consideration of several factors including: • radionuclide inventories by waste form and container; • the performance of waste containers in controlling access to waste forms by water and release of radionuclides in gaseous form; • the performance of waste forms in controlling releases of radionuclides in aqueous or gas phases; and • transport of radionuclides in aqueous and gas phases through a disposal facility. Container and waste-form performance and transport of radionuclides through a disposal facility are influenced by two external factors, infiltration and the chemical environment, whose effects must be included in analyzing the source term. In performance assessment, the objective of source-term modeling is to obtain a defensible projection of release rates of radionuclides from a disposal facility for the purpose of demonstrating compliance with applicable performance objectives. To that end, an analyst should consider all processes and events that significantly influence release and transport of radionuclides. However, it is not necessary, nor even desirable, to model all potentially important processes and events. The required degree of sophistication in modeling depends on the overall performance of a disposal system, and the source term should be described using the simplest and most defensible model that allows compliance with performance objectives to be demonstrated for wastes that are intended for disposal in a facility. For example, if an assumption of instantaneous container failure and complete release of radionuclides from waste forms at that time is sufficient, more sophisticated models that rely on the performance of waste forms or chemical controls on releases are not needed. Section 5.4 has discussed modeling approaches that can be used to estimate container and waste-form performance and radionuclide transport through a disposal facility. Modeling starts with the

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simplest and most defensible assumptions (instantaneous failure of containers and waste forms, with instantaneous transport from a facility) and increases in complexity by incorporation of mechanistic features that describe the physical and chemical behavior of a disposal system. Use of more sophisticated models must be supported by experimental data, more detailed modeling, or documented judgment. Container performance frequently is described using a container lifetime. An estimate of container lifetime should be based on experimental data relevant to the container material and environmental conditions in a disposal facility. Modeling of waste-form performance depends on the physical and chemical form of wastes. For example, solidified wastes often are modeled assuming diffusion-controlled release. Chemical controls on release are modeled by means of solubility limits and sorption onto waste forms. Under certain conditions (e.g., glass or metal waste forms), releases frequently are modeled assuming a uniform fractional release rate. Again, selection of a particular model and required parameters should be supported by experimental data that are relevant to waste forms of concern and site-specific environmental conditions. Radionuclide transport within a disposal facility in either the gas or liquid phase often is modeled by taking into account processes of advection, dispersion, diffusion, sorption and radioactive decay. As more sophisticated models are required to demonstrate compliance with regulatory criteria, more detailed justification of relevant parameters to describe container performance, release from waste forms, and transport is needed. For example, if solubility-limited releases are described by a single parameter, an analyst must show that a selected solubility limit is representative of the system over the duration of releases from a facility. This will entail an analysis to demonstrate that the solubility used to describe releases is representative of a wide range of environmental conditions that may occur within a disposal facility as it evolves over time (e.g., pH, Eh, presence of ligands). Similarly, if solubility is represented as a distribution of possible values, the choice of parameters to define a distribution must be justified. 5.5 Unsaturated Zone Flow and Transport 5.5.1

Introduction

Near-surface disposal of low-level waste generally occurs above ground or in the unsaturated zone. Thus, the unsaturated zone

184 / 5. PERFORMANCE ASSESSMENT MODELS provides the link between release of radionuclides from a disposal facility and migration in an underlying aquifer (zone of saturation), and water flow and radionuclide transport under variably saturated conditions generally must be addressed in performance assessment. Discussions in this Section focus mainly on methods of estimating flow rates in the unsaturated zone. Transport normally is modeled as described in Section 5.7. The importance of flow and transport in the unsaturated zone depends on a number of factors. At moist sites, a disposal facility generally lies within a few meters or tens of meters of an underlying aquifer, and flow and transport over such a short distance may not be important in assessing overall performance. At dry sites, however, an aquifer may lie hundreds of meters below a facility. In this case, the time required for radionuclides to reach an aquifer or the amount of dispersion that may occur during transport in the unsaturated zone can significantly reduce concentrations in water before radionuclides reach an aquifer, especially when radionuclides are shorter-lived. This is a desirable feature of any site with a deep water table. In order to take credit for those processes in performance assessment, a defensible model of the behavior of water and radionuclides in the unsaturated zone is needed. The simplest approach to addressing the unsaturated zone in performance assessment is to take no credit for delay and dilution in transport between a disposal facility and an underlying aquifer. This may be a reasonable approach at sites where the water table is shallow. Alternatively, models of unsaturated flow and transport may be included in an assessment. This Section extends discussions of flow and transport in groundwater in Sections 5.6 and 5.7 to highlight issues specific to variably saturated flow and transport modeling that affect the defensibility of a performance assessment. More detailed discussions of flow and transport in the unsaturated zone are provided, for example, by Bear (1979), Bear and Verruijt (1987), Domenico and Schwartz (1998), Evans et al. (2001), Faybishenko et al. (2000), Freeze and Cherry (1979), Hillel (1980), Looney and Falta (2000), NAS/NRC (2001), Nielsen et al. (1986), and Todd (1980). NAS/NRC (1990) and NCRP (1984a) discuss unsaturated flow and transport in the context of models used for regulatory compliance. Several reports sponsored by NRC discuss characterization of the unsaturated zone (Meyer et al., 1999; Rockhold, 1999; Wierenga et al., 1993) and a strategy for performance confirmation monitoring in the unsaturated zone (Young et al., 1999a; 1999b). In modeling flow and transport in the unsaturated zone, it generally is desirable to start with a relatively simple representation of

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a system. The objective should be to identify modeling approaches that are supported by available data and can be applied in different climates and to various disposal concepts. Most discussions in this Section are concerned with unsaturated zone flow under homogeneous conditions, which in many cases would be a simplifying assumption. When unsaturated media are highly structured or fractured, the task of modeling becomes more complex. Some problems of unsaturated flow modeling in such media are briefly mentioned. 5.5.2

Interfaces with Other Performance Assessment Models

Inputs to unsaturated zone flow and transport models include the infiltration rate through the cover of a disposal facility, flow rate through the facility, and release rates of radionuclides from the facility. The infiltration rate is obtained from models of cover performance discussed in Section 5.2; its value normally is estimated at a point below an engineered cover and is assumed to be at steady-state. The flow rate through a trench or vault and release rates of radionuclides are obtained from models of the source term discussed in Section 5.4. Outputs from unsaturated zone flow and transport models that are used as inputs to other models include the flow rate to an unconfined aquifer (recharge rate) and rate of radionuclide migration from the unsaturated zone into an aquifer. These two outputs, along with any additional direct recharge to an aquifer that may occur due to runoff from a cover, are the primary inputs to aquifer flow and transport models discussed in Sections 5.6 and 5.7. As with all interfaces between different models used in performance assessment, care should be taken to account for mass balance and consistency of units. Improper matching of models can be an important source of error in performance assessment. The interface between unsaturated and saturated zones must be evaluated with particular care, because assumptions that appear conservative may produce nonconservative results. For example, an increase in flow velocity in an aquifer would appear to yield higher concentrations of radionuclides at off-site locations, because of the resulting decrease in travel times. However, as the flow velocity in an aquifer increases, the volume of water available to dilute radionuclides that slowly percolate into the aquifer from the unsaturated zone also increases. Thus, a higher flow velocity in an aquifer yields a decrease in concentrations of radionuclides in the aquifer. The presence of such behavior indicates that intermediate results, such as concentrations and fluxes of radionuclides, should be calculated and carefully reviewed at all interfaces in a flow system.

186 / 5. PERFORMANCE ASSESSMENT MODELS 5.5.3

General Discussion of Unsaturated Zone Flow and Transport

Flow in the unsaturated zone depends greatly on soil characteristics and moisture content. Thus, it is necessary to understand how hydraulic conductivity changes as a function of the amount of moisture, as well as the suction pressure or matric potential, in soil. This added complication does not occur in describing saturated flow (Section 5.6). Properties of the unsaturated zone may be distinguished from those of the saturated zone in the following ways (Freeze and Cherry, 1979): • the unsaturated zone occurs above the water table and above the capillary fringe; • soil pores are only partially filled with water (i.e., the moisture content is less than the porosity); • fluid pressure is less than atmospheric pressure (i.e., the pressure head is negative and suction occurs); • hydraulic conductivity and moisture content are both functions of pressure head. Features of the unsaturated and saturated zones are illustrated in Figure 5.15, including the concepts of variable saturation and pressure head (suction pressure) in the unsaturated zone. One of two state variables, volumetric moisture content or saturation, is commonly used to characterize the amount of water in a given volume of unsaturated soil. Volumetric moisture content is the volume of water per unit volume of soil and pore space, and saturation is the fraction of the pore space that is occupied by water. The amount of water in a given volume of soil can change with time depending on soil properties and the infiltration rate. As described in the third property listed above and shown in Figure 5.15d, an important difference between saturated and unsaturated zones is that fluid in the unsaturated zone exists in a state of tension or negative (suction) pressure head because of capillary forces. Tension on soil surfaces attracts water in a manner similar to forces that cause water to rise in a capillary tube. As depicted in Figure 5.16, the smaller the spacing between soil grains, the greater the attractive force and suction pressure and, thus, the greater the capillary rise. Since suction is greatest in fluid immediately adjacent to soil, suction decreases as pores fill with water. Likewise, as moisture content increases, water will flow more readily (i.e., hydraulic conductivity increases). Flow in the

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Fig. 5.15. Groundwater conditions near the ground surface: (a) saturated and unsaturated zones; (b) profile of moisture content versus depth; (c) pressure-head and hydraulic-head relationships, with insets showing water retention under pressure heads less than (top) and greater than (bottom) atmospheric pressure; (d) profile of pressure head versus depth; (e) profile of hydraulic head versus depth (Freeze and Cherry, 1979).

188 / 5. PERFORMANCE ASSESSMENT MODELS

Fig. 5.16. Relationship between pore size (r) and pressure head (h) in unsaturated soil (NAS/NRC, 1990); pressure head is denoted by ψ in this Report.

unsaturated zone occurs under the influence of suction, which draws water from one void to the next, and gravity. Even in moderately dry soils, the suction potential can be more important than gravity, and upward flow can occur as in a capillary tube. Hydraulic conductivity and hydraulic potential (head) in unsaturated soils both depend on the degree of saturation. Relationships between moisture content (θ) and pressure head (ψ) (θ versus ψ curves), between hydraulic conductivity (K) and pressure head (K versus ψ curves), and between hydraulic conductivity and moisture content (K versus θ curves) are referred to as moisture characteristic curves. Knowledge of these curves for a range of conditions to be considered for a selected soil and climate is essential in modeling unsaturated flow. Empirical approximations of moisture characteristic curves are discussed in the following section. Examples of moisture characteristic curves for different soils are given in Figure 5.17. The top set of curves gives examples of moisture content as a function of (negative) pressure head, and the bottom set of curves gives examples of hydraulic conductivity as a function of pressure head. These curves indicate, for example, that hydraulic conductivity and moisture content in unsaturated sand both decrease rapidly with increasing suction, and that hydraulic conductivity in unsaturated sand can be less than in other soils

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with the same moisture content. At saturation, soils with smaller pores (e.g., silts, clays) typically are less conductive than soils with larger pores (e.g., sands). However, under unsaturated conditions at the same moisture content, soils with smaller pores can be more conductive than soils with larger pores. Curves on the top of Figure 5.17 illustrate that the suction required to begin draining a soil with smaller pores (fine sandy loam) is greater than the suction required to begin draining sand, which has larger pores. These curves also illustrate that sand drains abruptly with a decrease in suction, while a soil with smaller pores drains more slowly over a larger decrease in suction. In clay, there is an additional electrostatic force that attracts and bonds water to soil particles (Hillel, 1980), and this force contributes to the larger suction required to begin draining water from clay. Of great importance to modeling of unsaturated zone flow is the highly nonlinear relationship between moisture content and hydraulic conductivity. Curves in the bottom of Figure 5.17 illustrate that hydraulic conductivity can vary over a large range for a small change in suction head. The effect of this nonlinearity is that model calculations are strongly dependent on the choice of moisture characteristic curves. Thus, site-specific data are needed to support a choice of these curves. This choice is a particularly important concern when attempting to model unsaturated flow in fractured or structured media. Domenico and Schwartz (1998) and NAS/NRC (1990) discuss flow in fractured systems, but those discussions apply mainly to saturated systems. There has been recent work on modeling in unsaturated fractured rock (Evans et al., 2001; Faybishenko et al., 2000; NAS/NRC, 2001), but it is not yet clear how this work can be used in performance assessment. For the purpose of performance assessment, it probably is not worthwhile to attempt to model flow in an unsaturated, fractured system and defend the results. Rather, defensible bounding calculations normally should be used in such cases. Flow in the unsaturated zone is highly dependent on the infiltration rate, which generally varies with time as a result of the intermittent nature of precipitation. This time dependence is most pronounced at the surface and tends to dampen with increasing depth in the unsaturated zone or at locations beneath an intact engineered cover on a disposal facility. The transient nature of infiltration generally is not taken into account in long-term calculations used in performance assessment, but potential impacts of transient changes in infiltration should be acknowledged when defending models of flow in the unsaturated zone.

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Fig. 5.17. Examples of moisture characteristic curves for different soils (NAS/NRC, 1990): moisture content (θ) versus pressure head (h), with vertical arrow at h = –100 cm indicating soil water content at field capacity (top); hydraulic conductivity (K) versus pressure head (h) (bottom); pressure head is denoted by ψ in this Report.

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The foregoing discussions can be summarized by emphasizing that difficulties presented by nonlinear relationships among parameters, obtaining moisture characteristic data for fractured or structured media, and potential impacts of assuming steady-state conditions are three issues that must be addressed in defending models of flow in the unsaturated zone. Processes that govern transport of radionuclides in the unsaturated zone [i.e., advection, dispersion, and sorption (retardation)] essentially are the same as those that govern transport in the saturated zone. These processes are discussed in Section 5.7 and in such texts as Bear and Verruijt (1987), Domenico and Schwartz (1998), and NAS/NRC (1990).

5.5.4

Data Requirements

The primary data requirements for calculating unsaturated flow are the moisture characteristic curves for soils of concern. A number of empirical models have been developed to represent moisture characteristic curves (Gillham et al., 1976; King, 1965; Laliberte, 1969; Mualem, 1976a; van Genuchten, 1980; Varallyay and Mironenko, 1979; Visser, 1968). NAS/NRC (1990) has summarized many of the empirical relationships found in the literature; some of these relationships are given in Tables 5.2 and 5.3. Use of empirical models to describe moisture characteristic curves in performance assessment should be based on data obtained in laboratory or field experiments. Mualem (1976b) provides a number of tabulated data sets of moisture characteristics for a variety of soils. Such data, including statistical parameters that can be used in Monte-Carlo uncertainty and sensitivity analyses, also are provided by Meyer et al. (1997) and Meyer and Taira (2001). Data in the literature can be used in initial calculations and in comparisons with data collected at a given site. Many computer codes allow use of one or more empirical models or tabulated data sets of suction pressure, moisture content, and hydraulic conductivity. It is important to recognize that different empirical models yield different moisture characteristic curves. Given the potential importance of the highly nonlinear nature of models, curves developed using empirical formulae should be compared with data on which they are based. Data on infiltration rates and the corresponding moisture content in the unsaturated zone as a function of time are needed in order to consider the importance of transient behavior. In natural systems, transient infiltration events often have an effect near the

192 / 5. PERFORMANCE ASSESSMENT MODELS TABLE 5.2—Some empirical relationships between moisture content (θ ) and pressure head (ψ ).a Functionb

[ a ( θs – θr ) ] θ ( ψ ) = θ r + ---------------------------if ψ < 0, b [a + ψ ]

Source

Brutsaert (1966)

θ ( ψ ) = θ s if ψ ≥ 0 [ a ( θs – θr ) ] - if ψ < – 1, θ ( ψ ) = θ r + -------------------------------b [ a + ( ln ψ ) ]

Haverkamp et al. (1977)

θ ( ψ ) = θ s if ψ ≥ – 1 ( θs – θr ) θ ( ψ ) = θ r + ----------------------------------m- if ψ < 0, b [1 + a( ψ ) ]

van Genuchten (1980)

θ ( ψ ) = θ s if ψ ≥ 0 1 m = 1 – --b a b θ ( ψ ) = θ r + ( θ s – θ r ) ⎛ -------⎞ if ψ < 0, ⎝ ψ⎠ θ ( ψ ) = θ s if ψ ≥ 0 a See NAS/NRC (1990); pressure head is denoted by h in Table 3.1 of that report. b θs is saturated water content, equal to porosity; θr is residual moisture content; ψ is pressure head; and a, b, and m are empirical constants.

surface but not in deeper soils. An engineered cover on a disposal facility normally is expected to effectively dampen effects due to transient infiltration events (Section 5.2). Thus, as long as the cover functions as intended, it may be sufficient for the purpose of performance assessment to assume steady-state flow in the unsaturated zone below a cover. Use of steady-state approximations is discussed by Wierenga (1977).

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TABLE 5.3—Some empirical relationships between hydraulic conductivity (K) and pressure head (ψ ) or moisture content (θ ).a Functionb

K(ψ ) = a ψ

–b

K(ψ ) = K s e

Source

Wind (1955)

–a ψ

Gardner (1958)

ψ b K(ψ ) = ⎛ -------⎞ + 1 ⎝ a⎠

–1

Gardner (1958)

ψ - if ψ < ψ cr , K(ψ ) = -------------–b ψ cr

Brooks and Corey (1966)

K(ψ ) = K s if ψ ≥ ψ cr θ–θ K(θ ) = K s ⎛ --------------r-⎞ ⎝ θ s – θ r⎠ K(ψ ) = e

a( ψ – ψ

cr

)

b

, b = 3.5

if ψ 1 ≤ ψ ≤ ψ cr ,

Averjanov (1950) Rijtema (1965)

K(ψ ) = K s if ψ > ψ cr , ψ K(ψ ) = K ( ψ 1 ) ⎛ ---------⎞ ⎝ ψ1 ⎠ a K(ψ ) = K s ⎛ -----------------b-⎞ ⎝a + ψ ⎠

–b

K(θ ) = a e

b[θ (ψ)]

if ψ < ψ 1

if θ < θ s ,

Haverkamp et al. (1977) van Genuchten (1980)

K(θ ) = K s if θ = θ s , K(ψ ) = K s e

–b ψ

if ψ < 0 ,

K(ψ ) = K s if ψ ≥ 0 , K(ψ ) = a [ θ ( ψ ) ]

b

aSee NAS/NRC (1990); pressure head is denoted by h in Table 3.2 of that report. b Ks is saturated hydraulic conductivity; ψ is pressure head; θs is saturated water content, equal to porosity; θr is residual moisture content; and a, b, ψcr, and ψ1 are empirical constants.

194 / 5. PERFORMANCE ASSESSMENT MODELS 5.5.5

Modeling of Unsaturated Flow

The fundamental relation that is generally used to describe fluid flow in porous media is Darcy’s Law, which assumes that water flows in response to a gradient in hydraulic head (Section 5.6). This relation can be modified for unsaturated conditions in one dimension (vertical flow), defined by the coordinate z, to reflect the dependence of hydraulic conductivity (K) on pressure head (ψ) as: ∂h ∂ψ q = – K ( ψ ) -------- = – K ( ψ ) ⎛ ------- + 1⎞ , ⎝ ∂z ⎠ ∂z where: q = h

=

(5.17)

flow rate through a unit cross-sectional area (Darcy velocity) hydraulic head given by h = ψ + z

A pore velocity (vp) that represents the approximate velocity of a given molecule of water in unsaturated soil can be estimated as: q v p = ------ , θ where θ =

(5.18)

soil moisture content

A modified form of the flow equation for saturated media, referred to as the Richards equation (Richards, 1931), is used in calculating unsaturated flow. The model for unsaturated flow takes into account the dependence of hydraulic conductivity and the storage term on suction potential. The storage term is discussed, for example, in Bear (1979). A general form of the Richards equation is: ∂ψ C ( ψ ) --------- = ∇ ⋅ [ K ( ψ ) ⋅ ( ∇ψ + ∇z ) ] , ∂t where: C =

(5.19)

specific moisture capacity

Again, for unsaturated flow, pressure head (ψ) is negative, and a key difference between unsaturated and saturated flow is the dependence of hydraulic conductivity on pressure head. Equation 5.19 also can be written in terms of moisture content (θ) rather than pressure head.

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In performance assessment, steady-state and homogeneous conditions generally are assumed, although different conditions may be assumed in different soil layers in the unsaturated zone. At steady-state and noting that ∇z = 1, Equation 5.19 describing flow simplifies to: 2

K ( ψ ) ⋅ ∇ ( ψ + 1 ) = 0.

(5.20)

Given the nonlinear nature of unsaturated flow and difficulties associated with defending model calculations, it is often desirable to simplify the problem of modeling flow. The easiest simplification, noted in Section 5.5.1, is to ignore the unsaturated zone and assume that radionuclides are released from a disposal facility directly into an aquifer. If credit for unsaturated flow and transport is taken in an assessment, the next level of simplification involves a unit-gradient model, which assumes steady-state flow and an equality between the flow rate in the unsaturated zone and the infiltration rate. This model is intuitive in the sense that water is assumed to flow through the unsaturated zone at the rate at which it is introduced into the system. Using Darcy’s Law for unsaturated flow in Equation 5.17, the unit-gradient model can be derived by assuming that all flow is gravity driven and that suction gradients are not significant. Given those assumptions, Darcy’s Law becomes: q = K( ψ ) ,

(5.21)

and the pore velocity (vp) in Equation 5.18 is given by: K(ψ) v p = ------------- . θ

(5.22)

In steady-state flow, the bulk flow rate through the unsaturated zone is equal to the infiltration rate, and q and the hydraulic conductivity thus are equal to the infiltration rate. The moisture content for a given infiltration rate (i.e., the hydraulic conductivity, when assuming a unit gradient) can be obtained from a characteristic curve of K versus θ (Table 5.3). As indicated in Equation 5.22, the pore velocity can be obtained by dividing the infiltration rate by the moisture content obtained from such a curve. In this way, a unit-gradient model can be used to estimate a steady-state travel time through the unsaturated zone.

196 / 5. PERFORMANCE ASSESSMENT MODELS Several approaches that involve analytical or numerical solutions are available for more detailed modeling of flow in the unsaturated zone. Analytical solutions tend to be restricted in the range of conditions that can be considered, but they can be useful for a selected set of problems. Examples of analytical solutions for unsaturated flow are discussed by Bear (1979), Nielsen et al. (1986), and Philip (1955; 1957). Numerical solutions are more flexible than analytical solutions, because heterogeneous systems can be modeled using a variety of zones in a grid structure. However, numerical solutions also require adequate data sets to justify the conceptual model. Different solution techniques are discussed by Huyakorn and Pinder (1983), Istok (1989), and NCRP (1984a). Numerical modeling of unsaturated flow can involve complex techniques, given the dependence of hydraulic conductivity on suction head. The nonlinearity of the flow equation and resulting need for iterative approaches make it more difficult to obtain stable solutions than in numerical modeling of saturated flow. Thus, the potential for numerical errors in modeling is an important concern in defending results. A second issue of concern is the accuracy of moisture retention curves and the extent to which conclusions of an analysis depend on those data. There has been considerable recent work on modeling the unsaturated zone, especially in relation to waste disposal at arid sites with thick unsaturated zones below emplaced waste, such as the geologic repository at the Yucca Mountain Site in Nevada (Bodvarsson et al., 2000; 2001; Flint et al., 2001) and near-surface disposal facilities in the western United States (Khaleel, 2004; Magnuson and Sondrup, 2005; Mann et al., 2001). However, modeling of the unsaturated zone remains a challenging problem, especially in fractured or highly structured media, in part because of uncertainties in developing an appropriate conceptual model at a site (NAS/NRC, 2001). Simplified approaches that are more easily defended have proven to be adequate at some low-level waste disposal sites (e.g., Shott et al., 2000), and a combination of detailed and simplified modeling approaches has been used at other sites (e.g., McCarthy et al., 1998). 5.5.6

Summary

The primary areas of concern in modeling unsaturated flow are the need for site-specific data, the potential for numerical errors in flow calculations due to their difficulty, and effects of large nonlinearities that can occur in moisture characteristic curves.

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Section 5.5 has discussed a graded approach to modeling unsaturated flow and transport. The simplest option is to ignore time delays and dispersion in the unsaturated zone and thereby assume that radionuclides are released directly from a disposal facility into an aquifer. If unsaturated flow and transport are considered, the next level of detail involves a unit-gradient model, which assumes that the hydraulic conductivity (and Darcy flow rate for a unit gradient) through the unsaturated zone is equal to the infiltration rate. In this model, the water travel time through the unsaturated zone can be estimated on the basis of the Darcy flow rate divided by the moisture content. More complex methods of modeling unsaturated flow, including analytical and numerical techniques and incorporation of Monte-Carlo techniques of uncertainty analysis, are available. However, great care must be exercised when attempting those types of calculations, as the potential for error is much greater than in modeling saturated flow. At the present time, models of unsaturated flow through fractured or highly structured media are not considered practical for use in performance assessments of low-level waste disposal facilities. If site-specific data are available, it may be possible to defend results of complex models, but a large amount of data and experimental evidence (and resources necessary to obtain those data) would be required to support conclusions. 5.6 Aquifer Flow Flow in groundwater systems (saturated rock or soil) has been treated extensively in the literature (e.g., Anderson, 1979; Bear, 1972; Dagan, 1989; de Marsily, 1986; Freeze and Cherry, 1979). It is not the purpose of this Section to reiterate or summarize that body of information. Rather, this Section considers groundwater flow in the context of performance assessment. Hydrological aspects of performance assessment also have been treated extensively (NAS/NRC, 1990). However, other treatments lack a focus on the near-surface environment and an emphasis on regulatory decision making that are the primary concerns of this Report. 5.6.1

Modeling of Aquifer Flow

In performance assessment, the purpose of an analysis of aquifer flow in saturated media is to generate a velocity field, which then provides input to an analysis of radionuclide transport in groundwater (Section 5.7). In general, a velocity field can be expected to be a function of location and time. However, groundwater velocities are

198 / 5. PERFORMANCE ASSESSMENT MODELS not measurable in the field, nor can they be deduced directly from other field measurements. Instead, a flow field must be generated using a model that is conditioned using data on hydraulic head obtained from monitoring wells, together with data from pump tests or core hydraulic conductivity tests. The basis for virtually all analyses of groundwater flow is Darcy’s Law, which represents an assumption of a linear relationship between a pressure driving force and the Darcy velocity of a fluid (i.e., specific discharge given by the volume of fluid passing through a unit cross-section area per unit time) as: q = – K s ∇h, where: q = Ks = h =

(5.23)

Darcy velocity saturated hydraulic conductivity hydraulic head

Darcy’s Law fails at both low and high flow rates (for differing reasons) and when other driving forces become important, such as natural convection (density-driven flow) or osmotic phenomena. Such conditions are rarely encountered in performance assessment and are not considered further in this Report. In the intermediate range of groundwater flow rates in porous media, Darcy’s Law is generally accepted by practitioners of performance assessment. The range of applicability of Darcy’s Law has received considerable attention in the literature. It can be derived from fundamental arguments of momentum transport coupled with assumptions about spatial averaging (e.g., Anderson, 1979; Bear, 1972; Dagan, 1989; de Marsily, 1986; Freeze and Cherry, 1979). The flow equation in the saturated zone is obtained by combining Darcy’s Law with the principle of conservation of mass (continuity equation) to give: ∂h C s -------- = ∇ ⋅ ( K s ∇h ) + S ( x,y,z,t ) – R ( x,y,z,t ) , ∂t where: Cs = S = R

=

(5.24)

specific storage (fluid capacity) at saturation distribution of fluid sources at spatial coordinates (x,y,z) and time (t) distribution of fluid removal

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In this Equation, the saturated hydraulic conductivity (Ks) has been expressed as a second-order tensor. Solution of Equation 5.24, subject to appropriate boundary and initial conditions, estimates of spatial distributions of sources and sinks, and estimates of the spatial variability of the hydraulic conductivity tensor, constitutes the foundation of groundwater modeling in the saturated zone. The problem of characterizing a groundwater flow field in performance assessment typically proceeds as follows. • Site characterization activities that involve the sinking of wells are undertaken to characterize the geology and lithology of the subsurface environment, to observe the piezometric surface of an aquifer, and to conduct pump tests in which the response of an aquifer is observed under differing kinds of hydraulic stresses. • Results of site characterization activities are used to generate a data set of spatially and temporally varying hydraulic heads and estimates of hydraulic conductivity that represent a spatially averaged conductivity over an assumed, but indeterminate, volume of an aquifer. That is, hydraulic conductivities used as input to the flow equation are not fundamental quantities but are determined through use of a model. • A spatially complete estimate of hydraulic conductivity required to solve Equation 5.24 is generated by interpolating between points at which there are estimates based on measurement. A spatially continuous boundary condition around the entire domain of simulation also is needed. That requirement generally is met by making logical, but not directly testable, arguments about the influence of natural terrain features on the flow field. • Equation 5.24 is solved to produce a spatially continuous hydraulic head. • The Darcy velocity field is obtained from Equation 5.23 by calculating the gradient in head at all locations and then multiplying by the spatial estimate of hydraulic conductivity. This velocity field generally is three-dimensional. One additional step remains before an estimated velocity field can be used in an analysis of radionuclide transport. The velocity that a solute particle experiences is not the specific discharge (Darcy velocity) (q) but, rather, is the pore-scale or interstitial velocity (vp). Issues associated with converting specific discharge to interstitial velocity are discussed in Section 5.7.1.2.

200 / 5. PERFORMANCE ASSESSMENT MODELS In performance assessment, numerous additional assumptions are commonly invoked in solving Equation 5.24. Groundwater flow rates are almost universally assumed to be time invariant, even in assessments over hundreds to thousands of years or more, in which case the left side of the equation is set to zero. An argument for this assumption is that time frames of interest are much longer than typical periods of temporal variability in a flow field. However, the transient behavior of a system can influence projected annual doses resulting from releases in groundwater. For example, a flow system can be significantly different in winter and early spring than in summer, in which case projected doses resulting from irrigation of food crops may differ substantially from those obtained in a comparable steady-state analysis. Nevertheless, possible transients in flow generally are ignored. Additional common assumptions involve the spatial distribution of the source function in the flow equation. Recharge to an aquifer system often is assumed to be uniform over the spatial domain, since available information usually is not sufficient to support other assumptions. 5.6.2

Issues in Solving Flow Equation

Issues related to use of Equation 5.24 of particular importance to performance assessment include: (1) development of a steadystate analysis from intrinsically transient data, (2) effects of spatial scale and heterogeneity on the constitutive parameters, (3) determination of boundary conditions, and (4) flow in fractured media. These issues are discussed below. 5.6.2.1 Development of Steady-State Conditions. In assuming a time-invariant flow field in solving Equation 5.24, particular steady-state conditions must be justified. The process of extracting steady-state conditions from transient data is necessarily interpretive and subjective; there is no “correct” method that is appropriate to time periods of concern to performance assessment. In traditional hydrogeological practice, this subjectivity is mitigated by calibrating a steady-state model on the basis of observations. Such calibrations also are possible in performance assessment, but time periods of concern generally are long and steady-state conditions usually are assumed to persist for the duration of an analysis. A steady-state condition that is chosen for an analysis over a long time span may not be one that can be calibrated based on present-day conditions. 5.6.2.2 Scale and Heterogeneity. Issues of spatial scale and heterogeneity are almost always important in models of groundwater

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flow used in performance assessment. Issues of scale generally relate to interpretation and use of measurements that are directed toward establishing a spatial distribution of hydraulic conductivity (Dagan, 1986). Pump or slug tests yield an average conductivity over an unknown volume of an aquifer that is representative of the scale of the formation. Bench-scale conductivity measurements on cores frequently are used as inexpensive and convenient sources of point data (Taylor et al., 1987). However, to use such point measurements, they must be extrapolated to the spatial scale of the formation, which involves assumptions about variations in conductivity between point measurements. An approach called “kriging” is commonly used to estimate a conductivity field by interpolating between measured points (de Marsily, 1986). Kriging is an interpolation technique which assumes that the function being modeled has certain prescribed statistical properties. The significance of those assumptions for performance assessment is that kriging treats a geologic setting as though it can be represented by a relatively well-behaved statistical function. Thus, kriging is unlikely to provide satisfactory representions of discontinuities and complex geological settings, which may be important to an assessment at a particular site. The uncertain nature of the scaling process also can be represented by using a generalization of the kriging method to obtain multiple realizations of conductivity fields that fit available point measurements (de Marsily, 1986). In that approach, multiple randomly generated hydraulic conductivity fields, each of which is compatible with measured data, are analyzed. The method allows a broader representation of geological uncertainty than does the less general approach, but it still is limited due to the fact that it is a linear estimator. Some nonlinear estimators, such as indicator kriging (Journel and Huijbregts, 1978), are available but have rarely been used in performance assessment (Zimmerman and Gallegos, 1993). In indicator kriging, a geologic setting is subdivided into classes, each of which is assigned a probability of occurrence as a function of location. For application to flow modeling, each class would in turn have to be assigned a distribution of conductivities that could be sampled. The result is a complicated approach that permits alternative realizations of a geologic structure and its associated hydraulic conductivity field. Use of such approaches in performance assessment can be expected to be computationally intensive. A simpler approach that is frequently used in performance assessment is to assume that a hydraulic conductivity field is uniform, but uncertain. If uncertainty is treated adequately to include

202 / 5. PERFORMANCE ASSESSMENT MODELS adverse conditions in an analysis, that approach can represent uncertainty in a system for purposes of performance assessment without invoking any assumptions associated with geostatistical techniques. The approach assumes that a site has uniformly bad (or good) characteristics. Alternatively, more detailed assumptions can be invoked as needed during development of models used in performance assessment. Another issue related to development of an approximation to a hydraulic conductivity field is that of incompatibility of different sources of information. Site-specific information likely will consist of a combination of point estimates (perhaps with their associated interpolations) and well tests, which average information over an unknown volume of an aquifer. Those sources of information are not necessarily compatible, and they may even provide conflicting indications of the behavior of a system. Furthermore, traditional field-scale hydrogeological methods, such as pump and slug tests, produce information about mean flow behavior, but transport of radionuclides tends to occur along a path of least resistance. In performance assessment, the required scale of a groundwater flow analysis is dictated by regulatory performance objectives. Current performance objectives for low-level waste disposal discussed in Section 3.4.2 require an analysis of dose to an individual who normally is assumed to use water from a well located at or near the boundary of a disposal site. This requirement suggests that groundwater analyses at low-level waste disposal sites need not be conducted on a large (e.g., regional) scale, but that a scale of a few hundred meters at most will suffice. In contrast, regulations for disposal of high-level and transuranic wastes in geologic repositories (EPA, 1993a; 2001a; NRC, 2001) specify that the accessible environment where members of the public could be exposed is located at a distance of 5 or 18 km from emplaced waste, which imposes different requirements on the scale of groundwater analyses. 5.6.2.3 Boundary Conditions. Related to issues of spatial scale and heterogeneity is the issue of establishing boundary conditions to be used in analyzing groundwater flow in a performance assessment. As with most models of physical systems, boundary conditions imposed on an analysis have important effects on results of calculations. Boundary conditions of importance to analyses of groundwater flow have uncertainties that may result from several factors. Since the subsurface environment contains spatial heterogeneities, the

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location of any imposed condition, such as a no-flow boundary, is uncertain. This problem is particularly important in groundwater analyses, due to the paucity of available data for underground systems. Another factor is that performance assessments are invariably conducted using a steady-state flow model. Extracting steady-state boundary conditions for intrinsically transient phenomena clearly introduces additional uncertainty. An example of this kind of uncertainty is the difficulty in estimating recharge from transient rainfall and evapotranspiration data. Groundwater analyses used in performance assessment frequently treat the vadose and saturated zones as separate units, with a boundary condition of zero pressure head at the interface (Figure 5.15d). The problem is subdivided in this way for conceptual and numerical convenience, but the approach can lead to inconsistencies in flow behavior. At a fundamental level, the only condition that applies at the boundary between the vadose and saturated zones is continuity of mass, and boundary conditions should only be applied at the soil surface and at aquitards. However, the more rigorous approach generally is too cumbersome to be used in performance assessment. As in other aspects of performance assessment, a balance often is sought between rigor and simplicity, with the deciding factor being the degree of defensibility of an approach in the context of decisions to be made on the basis of results of an assessment.

5.6.2.4 Flow in Fractured Media. Site characterization and modeling of flow in the saturated zone are particularly difficult problems in fractured rock systems. Difficulties have been summarized by NAS/NRC (1990), which concluded that such systems “... are often complex and extraordinarily difficult to characterize, especially with the level of effort considered normal for most site investigations.” Reviewers of performance assessments in which modeling of flow in fractured media is attempted are likely to (and should) view an analysis with considerable skepticism. An approach that often may be considered conservative is to treat a fractured medium as a conduit for radionuclides, so that leachate released from a disposal facility is assumed to be essentially undiluted, except by radioactive decay, in transport to a receptor location. That approach, while not resolving technical difficulties associated with analyzing flow in fractured systems, may in some cases allow an adequate decision about regulatory compliance to be made.

204 / 5. PERFORMANCE ASSESSMENT MODELS 5.6.3

Summary

The purpose of an analysis of groundwater flow is to produce a flow field to be used as input to an analysis of radionuclide transport in groundwater. A flow field is derived using an indirect method of solution for a system that is dominated by largely unmeasured, spatially variable, or unknown physical properties and boundary conditions that can only be interpreted subjectively. A velocity field generated using this approach is nonunique, and many velocity fields usually can be found that reproduce data at a particular site equally well. Uncertainties associated with analyses of groundwater flow are too diverse to admit any general solution in the context of performance assessment. Consequently, uncertainties must be resolved on a site- and analysis-specific basis by means of model intercomparisons and corroboration by data. As a performance assessment is developed through successive iterations (Section 4), uncertainties in groundwater flow need to be resolved as required for the purpose of an analysis, which is to support a decision about regulatory compliance. In performance assessments of near-surface disposal facilities, many of the complicating factors in groundwater flow analyses normally are treated in simple, approximate ways. Since regulatory requirements dictate that the spatial scale of concern is relatively small, aquifer flow often is treated as unidirectional and spatially and temporally invariant, with the direction and gradient established on the basis of local monitoring data. That approach has the advantage that it eliminates the need for approximation of boundary conditions. However, it represents only one of several possible ways of accounting for spatial variability of soil properties, and it may result in underestimates of radionuclide transport if a representation is seriously in error. Furthermore, important exposure pathways sometimes involve discharge to surface water, which may be located a considerable distance from a facility boundary. More complex representations of aquifer flow patterns require introduction and justification of boundary conditions, spatial interpolation of material properties, and coupling of a flow analysis with an analysis of radionuclide transport. Assumptions invoked in each step must be carefully justified to produce a credible performance assessment. Uncertainties associated with groundwater flow systems are best treated by an analysis of multiple conceptual models. In that way, effects of different credible assumptions on results of an analysis can be examined. Such an approach is particularly relevant in

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analyzing groundwater flow systems, due to the nonunique nature of any representation of physical properties. At any site, there will be alternative ways of representing spatial variability of properties that are consistent with available data. Each alternative representation can be expected to provide different (and, in some cases, markedly different) flow fields. Thus, it is advisable to consider numerous competing alternatives, to ensure that system behavior is acceptable over a range allowed by available data. Integration of multiple conceptual models in the overall framework of a performance assessment is discussed further in Section 7.

5.7 Radionuclide Transport in Groundwater and Surface Water Performance assessments of low-level waste disposal facilities usually incorporate an assumption that transport of radionuclides in groundwater is the most important pathway for releases to the environment. Transport of radionuclides in groundwater is assessed by combining results of a source-term analysis (Section 5.4) and an analysis of flow and transport in the unsaturated zone (Section 5.5) with results of an analysis of groundwater flow (Section 5.6). Inputs to a transport analysis include a velocity field for the groundwater flow system, which is obtained as output from a flow analysis, and radionuclide-specific release rates into the aquifer, which are obtained as output from source-term and unsaturated zone analyses. The desired output of an analysis of radionuclide transport in groundwater depends on the features of a disposal site and regulatory performance objectives. Because performance objectives for low-level waste disposal are expressed in terms of annual dose to individuals (Section 3.4.2), desired outputs are concentrations of radionuclides in water at assumed receptor locations as a function of time. An exposed individual often is assumed to reside at or near the boundary of a disposal site and to use untreated well water obtained at that location, in which case radionuclide concentrations in well water are calculated. Groundwater also may discharge to local surface water, in which case radionuclide concentrations in surface water may need to be estimated. A transport analysis thus provides the link between source-term and flow analyses and analyses of exposure pathways and radiological impacts (Figure 5.3). This Section is concerned primarily with modeling of radionuclide transport in groundwater. Modeling of transport in surface water is discussed in Section 5.7.5.

206 / 5. PERFORMANCE ASSESSMENT MODELS Modeling of transport of radionuclides (or other contaminants) in groundwater has been treated extensively in the literature, and at different levels of sophistication (e.g., Bird et al., 1960; Codell and Duguid, 1983; Dagan, 1989; de Marsily, 1986; NAS/NRC, 1990; NCRP, 1984a). Consequently, a comprehensive review of the subject is not given in this Report. Instead, discussions in this Section focus on aspects of a transport analysis that are particularly important to performance assessment. In analyses of low-level waste disposal systems, transport of radionuclides must be projected over long periods of time and sometimes over large spatial domains. Furthermore, analyses frequently are based on limited amounts of relevant information. 5.7.1

Phenomena That Influence Transport in Groundwater

This Section provides a review of phenomena that influence transport of radionuclides in groundwater, including sorption, advection, diffusion, dispersion and radioactive decay. In each case, ways in which a phenomenon is commonly represented in performance assessment are described. Methods of combining different phenomena mathematically and application of those methods to modeling of transport in the unsaturated zone are discussed in Section 5.7.2. In transport analyses used in performance assessment, radionuclides generally are considered to be dissolved ionic species in aqueous solution. This assumption is based on current waste acceptance criteria, which often preclude disposal of significant amounts of organic materials (contaminated or otherwise). As a result, phenomena associated with multiphase transport of nonaqueous phase contaminants generally are considered to be unimportant and are not discussed here. 5.7.1.1 Sorption. Sorption, as it is customarily treated in performance assessment, is a general phenomenon that takes into account contributions from all reversible heterogeneous reactions of dissolved contaminants with solid surfaces, including chemisorption, physisorption and ion exchange. Other types of reactions that are essentially irreversible are not accounted for by sorption (NAS/NRC, 2000). In virtually every performance assessment, sorption processes are lumped into a single linear sorption factor, called the distribution coefficient (Kd). That approach is taken because (1) computer codes that couple geochemical effects with transport are rare and computationally intensive, (2) existing geochemical databases and models are not sufficiently reliable to justify the

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additional modeling detail that would be required, and (3) the approach is considered adequate to represent geochemical information that is available for use in performance assessment. Consequently, despite the simplistic nature of the Kd concept, it represents the best available approach to practical modeling of sorption. The distribution coefficient (Kd) generally is defined as the derivative of the concentration (C) of a sorbed contaminant in the solid phase with respect to its concentration in solution (fluid) adjacent to a solid: ∂C solid -. K d = --------------∂C fluid

(5.25)

On the basis of this definition, Kd often is referred to as a solid/ solution distribution coefficient. The most common assumption used in performance assessment is that this derivative is constant across a range of concentrations of interest, in which case the distribution coefficient is a constant given by: C solid -. K d = -----------C fluid

(5.26)

Since concentrations in the solid phase usually are expressed in units of mass or activity per mass of soil or rock, and concentrations in a fluid usually are expressed in units of mass or activity per volume of fluid, Kd is expressed most commonly in units of volume per mass (e.g., m3 kg–1). Given the model of sorption described above, the primary issue for performance assessment is a need to justify values of Kd used in an analysis. Sorption capabilities of soils can be expected to vary both spatially and temporally in unknown ways. Furthermore, many important radionuclides have the potential to change valence states and chemical speciation under differing chemical conditions in groundwater. As a result, care should be taken in choosing Kds, especially when a soil-water system is assumed to be homogeneous, as is often the case. In performance assessment, it often is appropriate to select Kds that result in conservative estimates of concentrations of radionuclides in water at assumed receptor locations beyond the boundary of a disposal site. However, care must be taken in selecting Kds that are intended to be conservative. A common misconception is that selecting low Kds is universally conservative, but there are at least two circumstances in which this is not the case. The first occurs when pathways other than releases to groundwater are important

208 / 5. PERFORMANCE ASSESSMENT MODELS (e.g., when erosion of material above emplaced waste or intrusion into waste by plants, animals or humans is of concern). In such cases, selection of higher Kds can result in higher doses when waste is exposed at the surface or brought to the surface by intrusion. The second occurs when a parent radionuclide of relatively low radiotoxicity decays to radionuclides of higher radiotoxicity. An important example is 238U, which decays to 226Ra, 222Rn, and 210Pb. Assigning a low Kd to uranium could allow 238U to be rinsed from a disposal system before significant ingrowth of its decay products occurs, but an assumption of a high Kd would cause uranium to be retained in a disposal system, resulting in higher calculated concentrations of more highly radiotoxic decay products. 5.7.1.2 Advection. Advection is the forced transport of dissolved radionuclides by movement of water. In performance assessment, movement of groundwater usually is considered to result only from hydraulic forces. Consequently, the gradient of hydraulic head usually is the primary motive force for advectively-driven radionuclide transport in groundwater. The rate of such transport usually is described by a groundwater velocity. The velocity needed in an analysis of radionuclide transport in groundwater is the pore-scale, or seepage, velocity (vp) given by: q v p = ---------- , ηφ where: q = η φ

= =

(5.27)

specific discharge (Darcy velocity) calculated using Equation 5.23 total porosity of soil empirical modifying factor which accounts for the fact that not all pore spaces are available for transport

In theory, the effective porosity (ηφ) may be a tensor quantity, but it is highly unlikely that properties of the subsurface environment would ever be known in sufficient detail to allow a tensor effective porosity to be used in performance assessments for low-level waste disposal facilities. Freeze and Cherry (1979) suggest that φ is in the range of 0.98 to 1.18 in sands, with wider variations occurring in more complex soil types. In unsaturated soils, it is customary to assume that the advective velocity is modified by the reduction in specific area through which flow occurs. Using an assumption that the superficial area fraction of water-filled pore space is equal to the volumetric fraction of filled pores, the seepage velocity is expressed as:

5.7 RADIONUCLIDE TRANSPORT

q v p = ---------- , θφ where: θ =

/ 209 (5.28)

volumetric moisture content (Section 5.5.5)

Owing to a lack of data, φ generally is assumed to be unity. In modeling saturated and unsaturated systems, it should be understood that the effective porosity (ηφ) and effective moisture content (θφ) are empirical constructs that cannot be predicted from a knowledge of soil structure. These parameters can only be determined by means of a tracer test over a spatial scale of interest. 5.7.1.3 Diffusion. Diffusion is a fundamental mechanism for transport of dissolved ionic species in groundwater. The most common representation of transport via diffusion is Fick’s First Law, which states that the diffusive flux (J) of the ith contaminant is linearly proportional to the gradient in concentration: J i = – D i ∇C i , where: D = C

=

(5.29)

a constant of proportionality, known as the diffusion coefficient concentration of the contaminant in solution

For diffusion of conservative tracers (i.e., contaminants that do not chemically interact with soil) in porous materials, D is commonly assumed to represent an “effective” diffusion coefficient that includes alteration of a diffusion rate by the porosity and tortuous diffusion path. When chemical sorption is included in a diffusion coefficient, it is referred to as an “apparent” diffusion coefficient. It can be shown mathematically that an apparent diffusion coefficient includes linear sorption effects when a transport model that incorporates a distribution coefficient (Kd) is used. It is important to realize that different analysts use these terms in different ways, often interchangeably, and that there is no standard terminology. Therefore, care must be taken to understand which concept is being used when diffusion coefficients are obtained from the literature. It must be emphasized that the expression of Fick’s Law in Equation 5.29 is a simplified model that neglects nonideal solution behavior, nonlinear dependencies of diffusive flux on the concentration gradient, and cross-contaminant fluxes. As a practical matter, those influences are universally ignored in performance assessment.

210 / 5. PERFORMANCE ASSESSMENT MODELS Given the form of Fick’s Law in Equation 5.29, effective diffusion coefficients must be defined for each radionuclide and material (soil type) of interest. Intrinsic diffusion coefficients (i.e., diffusion coefficients measured in free water) often are used when data for a soil-water system of interest are unavailable. That approach tends to overestimate the contribution of diffusion to transport, since effective diffusion coefficients generally are less than intrinsic values, frequently by many orders of magnitude. The degree of conservatism in the approach depends on conditions at a disposal site, and no generalization can be made. Intrinsic diffusion coefficients sometimes are used in conjunction with semiempirical equations that incorporate effects of porosity and tortuosity on the effective diffusion coefficient (Price et al., 1993). Use of such approaches can lead to improved estimates of an effective diffusion coefficient, but parameters in those equations must be chosen with care, since they cannot be measured. 5.7.1.4 Dispersion. Dispersion refers to an observed spreading of contaminants in an advective velocity field. In the groundwater transport literature, dispersion often is described as a sum of two physical effects. The first is molecular diffusion, which is described in the previous section. The second is referred to as “mechanical” dispersion and is ascribed to the tortuous flow path that a tracer must follow during transport through soil, which leads to a mixing effect. However, “mechanical” dispersion is not a true physical phenomenon. Rather, it is an approximate representation of velocity variations that are not explicitly taken into account in a flow model. That is, dispersion is an informational effect, and the extent to which it is included in a model depends on the amount of resolution used in the model. To illustrate this concept, consider the original treatment of dispersion in transport through a pipe developed by Taylor (1953) and Aris (1956). For that situation, the velocity field can be derived exactly. In Taylor-Aris dispersion theory, information about a microscopic flow field is used to provide a mathematical link between the macroscopic average flow velocity and the spreading behavior of a contaminant carried by the fluid. If a microscopic representation of a flow field is used, there is no need to invoke the concept of dispersion. It is only when a velocity field is averaged over some volume that dispersion is needed to account for discrepancies between contaminant spreading predicted on the basis of an average velocity and the actual behavior of a system. Similarly, dispersion in porous media represents the relationship between a macroscopic observable velocity and a spreading

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that results from velocity variations at a smaller scale than the one over which an average velocity is defined. The difference between conditions studied by Taylor (1953) and Aris (1956) and conditions of concern in groundwater modeling is that the nature of velocity variations in groundwater systems can never be known. It is of theoretical and practical interest to note that the amount of dispersion used in analyzing transport should decrease as a flow model becomes more complex. For example, Moltyaner and Killey (1988) found that when detailed knowledge of a flow field was obtained from a tracer test, use of small dispersivities could be justified. The most common mathematical representation of dispersion is to treat it identically to molecular diffusion. Thus, the “dispersive flux” (Jdisp) of the ith contaminant is calculated as: J disp,i = – E i ∇C i , where: E =

(5.30)

dispersion coefficient

Theoretical treatments of this approach in the literature suggest that the dispersion coefficient can be a second-order tensor. In practice, however, data to support determination of different components of this tensor are unavailable, and this aspect of the dispersion coefficient is usually neglected in performance assessment. The approach to representing dispersion in Equation 5.30, which sometimes is called Fickian dispersion because of the similarity to Fick’s First Law of Diffusion in Equation 5.29, is equivalent to assuming that unknown velocity variations are randomly distributed about a mean value in accordance with a Gaussian (normal) statistical distribution. The approach represents an extreme extrapolation of Taylor-Aris theory. It is justified to some extent by observations of transport in columns that contain uniform sediments, but it is of dubious applicability in the field. Nevertheless, Fickian dispersion is applied in many analyses, with the justification that no information is available to support any other approach. In a further extrapolation of the form of Taylor-Aris theory represented in Equation 5.30, the dispersion coefficient is assumed to be linearly proportional to velocity: E x,i = D i + α x v x ,

(5.31)

212 / 5. PERFORMANCE ASSESSMENT MODELS where: Ex = D αx

= =

dispersion coefficient for the ith contaminant in the direction of flow (x) diffusion coefficient dispersivity in the x direction

In developing this equation, the tensorial nature of the dispersion coefficient has been neglected; data sufficient to support a multidimensional treatment of dispersion are highly unlikely to be available at near-surface waste disposal facilities. Some experimental evidence for the appropriateness of Equation 5.31 is obtained in column experiments, but there is no evidence that it is appropriate at field scale in natural formations. Dispersivities have been observed to increase with increasing spatial scale of a system to which they are applied. That is, a dispersivity deduced from a bench-scale column experiment can be expected to be less than a dispersivity that will match data at a larger scale, either in the laboratory or in the field. That behavior has important implications for performance assessment, because it means that dispersivities measured at bench scale do not apply at the scale of interest at a disposal site. The only way to ensure that a dispersivity is appropriate to a particular spatial scale is to calibrate dispersivity using an existing plume. Such a calibrated dispersivity can be expected to reproduce similar plumes on the same scale, but it cannot be extrapolated with confidence to other conditions or spatial scales. A measured dispersivity that applies to conditions of concern in a performance assessment is expected to be available only rarely. In the absence of a measured site- and scale-specific dispersivity, a general approximation frequently is used in which the longitudinal dispersivity is set to one-tenth of the spatial scale of the problem (Walton, 1988). That approximation usually is justified by a compilation of existing field-scale dispersion data (Gelhar et al., 1985; 1992) and a related attempt to develop a universal scaling relationship for dispersivities (Neuman, 1990). However, none of those investigators suggested that the one-tenth rule is appropriate. Indeed, Gelhar et al. (1992; 1993) criticized the idea that any universal regression fit would be appropriate. Even a cursory examination of data cited by Gelhar et al. (1992) shows variations in dispersivities about a regression line by orders of magnitude. The foregoing discussions have important implications for performance assessment. Site-specific data on dispersion likely will be scarce. This means that dispersivities must be assumed for use in an analysis, but there is no general way to choose appropriate

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values. Consequently, dispersivity must be considered highly uncertain and should be varied over wide ranges to determine the importance of this parameter to the results of an analysis. 5.7.2

Combination of Phenomena That Influence Transport in Groundwater

There are two primary methods of combining the phenomena of sorption, diffusion, dispersion and radioactive decay in an analysis of transport in groundwater. The first, and most common, approach is to invoke a Fickian description of dispersion in Equation 5.30 and include diffusion as a dispersive-flux term. As discussed in Section 5.7.1.4, the tensorial nature of the dispersion coefficient generally is ignored in performance assessment. Prior to solving the advective-dispersion equation, additional approximations and simplifications are commonly introduced. Invoking the well-known retardation factor allows an incorporation of heterogeneous chemical reactions that tend to retard the movement of radionuclides in an aquifer: ρb R = 1 + -------K , η d where: ρb = R =

(5.32)

dry bulk density of soil retardation factor

The retardation factor represents the water velocity relative to the velocity of a contaminant, and the effective porosity is assumed to be equal to the total porosity (η). It often is assumed that similar dimensional arguments can be used for contaminant transport in unsaturated porous media, in which case the retardation factor takes the form: ρb R = 1 + -------K , θ d where: θ =

(5.33)

moisture content

In contaminant transport in unsaturated media, the conceptual interpretation of the retardation factor is that the product θR represents the total mass in the solid phase divided by the mass in solution. In developing Equation 5.33, the surface chemical behavior of partially saturated soil is assumed implicitly to be identical

214 / 5. PERFORMANCE ASSESSMENT MODELS to that of saturated soil. The appropriateness of this equation is rarely questioned, but it has never been demonstrated experimentally. Indeed, there is some evidence that the retardation factor for some soils and contaminants is directly dependent on moisture content, rather than inversely dependent. Since the retardation factor (R) in saturated or unsaturated media represents chemical behavior, its value generally depends on the chemical element, but not on the isotope of a given element. Thus, all isotopes of a given element will have common retardation factors. Decay and ingrowth of radionuclides provide additional sinks and sources that usually are included in the advective-dispersion equation. With radionuclide concentrations in solution (C) expressed in units of activity per volume of fluid and by invoking the customary assumptions of a linear sorption isotherm (i.e., an assumption that Kd is independent of concentration) and spatially invariant, isotropic dispersion, the advective-dispersion equation for radionuclide i takes the form: ∂C R i -----------i- = ∇ ⋅ ( E k ∇C i ) + v p ⋅ ∇C i + S i ( x,y,z,t ) – λ i C i R i + ∂t

i–1

∑ λi fi,j Cj Rj ,

j=1

(5.34)

where: Ek = S

=

components of dispersion coefficient in x, y, and z directions all sources and sinks other than radioactive decay

In the last two terms that describe decay of the radionuclide and its production by decay of other radionuclides, M is the decay constant, and fi,j is the fraction of decays of precursor radionuclide j that produce radionuclide i. Equation 5.34 forms the foundation of most models of radionuclide transport in groundwater, usually with some additional simplifying assumptions discussed in the following section. The second common approach to combining different phenomena that influence radionuclide transport in groundwater involves treating dispersion as a variability in the flow field, usually as a statistical distribution of velocities. The dispersion term in Equation 5.34 then reduces to a molecular diffusion term, and the variability in velocity is accounted for explicitly. Alternatively, even diffusion is neglected, in which case transport is considered to be driven only by advection, with a variable velocity field used to account for dispersion. Issues associated with use of the Fickian

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form of the dispersion equation in Equation 5.30 (Dagan and Neuman, 1991) have not affected its use in performance assessment to a significant extent. Many other phenomena may affect transport of radionuclides in groundwater systems. Examples include osmosis, electro-osmosis, density-driven advection, colloidal transport, and multicomponent interactions among ionic species. The presence of microorganisms may affect local geochemical conditions or may enhance or retard radionuclides by acting as colloidal material. Those phenomena generally are ignored in performance assessment. Their neglect may be justifiable in most circumstances, owing to the nature of soil-water systems and the dilute concentrations of contaminants that typically are obtained in a transport analysis. Nevertheless, alternative transport mechanisms should be considered for each disposal system and their importance evaluated, especially when they would tend to enhance or facilitate transport of radionuclides (e.g., colloidal transport in fractured rock systems or, in some cases, interactions with microorganisms or organic compounds produced by microorganisms). 5.7.3

Methods of Solution of Groundwater Transport Equation

In principle, a single general technique for solving Equation 5.34 could be used in all performance assessments. Such a technique would involve solving equations in multiple dimensions with large spatial regions coupled through appropriate conditions at interfaces between regions, including considerations of spatial and temporal heterogeneities. Given that the surface dimensions of a near-surface disposal facility usually are comparable to or greater than the distance to an assumed receptor location at the site boundary, modeling of releases to groundwater over an extended area, rather than as a point source, generally is required. However, general solutions typically are cumbersome, and large amounts of computer time are required to obtain results over long time periods of concern. As a result, a variety of simpler numerical, analytical, and semi-analytical solutions are commonly used. The following discussions are based on an assumption that, in practice, calculations of radionuclide transport in groundwater at most sites will be carried out using simplified geometries, often a one-dimensional system with spatially uniform properties. Under such conditions, the transport equations reduce to a set of coupled ordinary differential equations. The particular method of solution used in an assessment often is chosen on the basis of a desire to achieve a reasonable balance between modeling efficiency and complexity and between correctness of results and a conservative bias.

216 / 5. PERFORMANCE ASSESSMENT MODELS 5.7.3.1 Analytical Solutions. Analytical solutions of the transport equation can be broadly classified as one of two kinds: solutions in which dispersion is included, and solutions in which it is not. If dispersion is not included, Equation 5.34 reduces to a first-order partial differential equation (or a coupled system of such equations in the case of decay chains), which can be solved by the well-known method of characteristics (Carrier and Pearson, 1976). In essence, such solutions displace radionuclides in space and time through a groundwater system, with concentrations modified only by decay, ingrowth and reactions (sorption). If dispersion is included in the transport equation, solutions of the resulting second-order partial differential equation can be obtained using a variety of solution methods and approximations. Complex analytical solutions are available in the literature for transport of decay chains containing several members. 5.7.3.2 Green's Function (Semi-Analytical) Solutions. Green’s functions provide a general method of solving inhomogeneous differential equations (Bender and Orszag, 1978). The principle behind this method is that a solution of the differential equation: L(y) = f(x) , where: L = f(x) =

(5.35)

differential operator on the dependent variable (y) an inhomogeneity

The inhomogeneity f(x) can be generated from the solution of the related differential equation: L ( g ) = δ ( x – x' ) , where: δ = g =

(5.36)

Dirac delta function Green’s function

The Dirac delta function is defined in Section 5.4.5.1.1 (Equation 5.10). If Equation 5.36 can solved to obtain the Green’s function, the solution of Equation 5.35 is given by: ∞

y(x) =

∫ f ( x' )g ( x – x' )dx' . –∞

(5.37)

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The Green’s function approach is particularly useful in performance assessment because of the clear manner in which arbitrary and complex forcing functions can be treated. Codell et al. (1982) describe a number of useful Green’s function solutions for groundwater transport analyses. However, the real value of this method is in developing site- and design-specific solutions of new problems. For instance, solutions for transport in unsaturated soils can be developed for a variety of source-term functions; such solutions are related to, but differ from, solutions described by Codell et al. (1982). The main drawback of the method is the simplicity of underlying assumptions used when generating the Green’s function. For example, the Green’s function that forms the basis for solutions given by Codell et al. (1982) assumes uniform, homogeneous, and semi-infinite or infinite media and flow in one dimension. In addition, a Green’s function to describe transport of radionuclides in decay chains is not available. Consequently, decay chains can be analyzed using this method only if a simplifying approximation is used. Two assumptions that allow Green’s function solutions are that (1) a parent radionuclide and its decay products are in activity equilibrium at all times and locations or (2) all members of a decay chain are transported at the same rate [i.e., have the same distribution coefficient (Kd)] (Codell and Duguid, 1983). Those approximations may give conservative results, but such claims should be justified. 5.7.3.3 Finite-Element and Finite-Difference Solutions. Most computer codes that analyze radionuclide transport in groundwater obtain solutions of the transport equation using numerical finite-element or finite-difference methods. In these methods of solution, a physical domain is partitioned into a discrete set of finite-sized regions. Governing equations are written in terms of the finite regions, which leads to a set of discrete linear algebraic equations that are solved by matrix inversion methods. Differences between finite-element and finite-difference methods are well established (Finlayson, 1980), and are not important to this discussion. Practical considerations in applying finite-element and finitedifference methods that are most important to performance assessment are summarized as follows. • They can be adapted to arbitrary and complex modeling geometries. Thus, they are most appropriate for complex groundwater systems.

218 / 5. PERFORMANCE ASSESSMENT MODELS • Their accuracy increases as the discretization in space and in time is made finer. Since performance assessments often are carried out over thousands of years, a conflict between computational effort and accuracy arises. The computational burden can be significant if an uncertainty analysis of a model is conducted. • When solving discretized equations, spurious “numerical dispersion” is introduced into a solution, which makes solutions inaccurate. The accuracy of numerical solutions can be evaluated adequately only by solving the equations using progressively finer discretization meshes, until an analyst is confident that the discretization is sufficiently fine. Thus, addressing numerical accuracy also adds to the computational burden of an analysis. A simplified numerical method of solution, referred to as compartment modeling, was used in performance assessments beginning in the 1990s (EPRI, 1996; SKI, 1996) following development of computer codes that allow flexible implementation by means of object-oriented programming (IAEA, 2004c). Mathematically, the compartment method is formally identical to a finite-difference approach. However, implementations using compartment models tend to have a coarser discretization than are typical in a finitedifference solution, with the result that solutions may be characterized by large numerical dispersion. 5.7.3.4 Stream-Tube Solutions. The stream-tube approach to modeling radionuclide transport in groundwater is a specialized method that has been developed for performance assessment. A stream tube is defined from flow paths of radionuclides, as shown in Figure 5.18 (Kozak and Olague, 1995). Transport is considered to be one-dimensional within a stream tube, and transport is assumed not to occur across the boundary of a stream tube. A key to use of the stream-tube approach is to define an appropriate size of a stream tube. Since performance objectives for low-level waste disposal are expressed in terms of annual individual dose (Section 3.4.2), required outputs of a transport model are concentrations of radionuclides in water. Therefore, given the source term for a disposal facility expressed as the quantity of radionuclides released per unit time, it is necessary to define the spatial extent (area) of a stream tube through which transport is assumed to occur. The spatial extent of a stream tube should be defined on the basis of knowledge of flow paths of groundwater at a site.

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Fig. 5.18. Depiction of stream tube used in modeling radionuclide transport in groundwater (Kozak and Olague, 1995).

The stream-tube approach was developed primarily to provide a method for rapid analyses of transport in groundwater over long time periods of concern in performance assessments of radioactive waste disposal facilities. The method is implemented by defining a one-dimensional region within which the advectivedispersion model in Equation 5.34 can be solved analytically or semi-analytically. An example of a one-dimensional flow model for which an analytical solution can be obtained is a model developed and used by EPA (2000c). A novel numerical method of solving Equation 5.34 in one dimension for use in a stream-tube model is the distributedvelocity method (Campbell et al., 1981). In that method, accuracy increases with increasing time step. Therefore, calculations over long time periods of interest to performance assessment can be performed quickly and efficiently. More recent developments in the distributed-velocity method are reviewed by Olague et al. (1991). Other stream-tube solution methods are available that incorporate dispersion as a distribution of velocities in a tube, thus treating transport as a purely advective process (Simmons et al., 1986). 5.7.4

Boundary Conditions

In performance assessments of near-surface disposal systems, an analysis of flow and transport in water below the ground surface normally is subdivided into analyses of the vadose (unsaturated) and saturated zones. The purpose of this subdivision is to enhance

220 / 5. PERFORMANCE ASSESSMENT MODELS the simplicity and clarity of an analysis, both conceptually and mathematically. However, when vadose and saturated zones are analyzed separately, appropriate boundary conditions must be specified at the interface between the two.31 The following options can be considered. • If dispersion and diffusion are neglected, transport in the vadose and saturated zones can be modeled using a firstorder partial differential equation, and no boundary condition is needed. An argument in favor of this approach is that it would tend to be conservative, at least for long-lived radionuclides. However, when the half-life of a radionuclide is comparable to the travel time in the vadose zone, the approach may not be conservative. • A boundary condition of zero concentration can be specified at the base of the vadose zone. This condition represents discharge into an aquifer with a high flow velocity that rapidly carries radionuclides away from the boundary. It can be argued that this approach is conservative with respect to calculating radionuclide fluxes, because it maximizes dispersive flux, but it is inappropriate for the purpose of evaluating concentrations at the interface. • A boundary condition of zero gradient in radionuclide concentration can be specified. This condition represents cases in which advective transport dominates at the boundary. Many groundwater transport codes use this assumption, and it frequently is not scrutinized. However, for a low-to-moderate Peclet number, which is a quantity proportional to the pore velocity divided by the dispersion coefficient, this condition is not physically appropriate. It can be argued that a zero-gradient boundary condition results in less conservative radionuclide fluxes than a boundary condition of zero concentration, but it is not clear what effect that condition has on the overall conservatism of an analysis. • Transport can be analyzed by assuming an infinite or semi-infinite domain. This approach is tantamount to ignoring the presence of the water table but evaluating what occurs at that plane anyway. The approach most closely represents discharge into an aquifer that has a strong downward component of velocity in the neighborhood of a discharge. 31If the hydrologic system is modeled as an entire unit, a condition of continuity of mass must be applied at the interface between the vadose and saturated zones (Section 5.6.2.3).

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• Unsaturated flow and transport fields can be coupled to aquifer flow and transport fields, and continuity of mass applied. This approach is mathematically correct, and it represents a division of the system in a different manner than described previously. However, as noted previously, this approach frequently is impractical for use in performance assessment, due to the complexity of calculations. Except for the last option, some aspect of each of these approaches is not physically correct. Nonetheless, in individual cases, they may be appropriate for use in performance assessment. There frequently are no natural boundary conditions that can be applied to analyses of radionuclide transport in groundwater. An example of this situation is an analysis of systems with a low Peclet number, in which diffusion and dispersion are comparable in importance to advection. In such cases, it is difficult to assign a boundary condition on radionuclide concentrations anywhere in a reasonably-sized domain, because boundary conditions tend to be influenced by phenomena between the boundaries. 5.7.5

Modeling of Transport in Surface Water

Modeling of transport of radionuclides in surface water is not considered in many performance assessments of low-level waste disposal facilities. However, surface-water transport could be important at some sites, such as sites where surface runoff of radionuclides could be significant or contaminated groundwater could discharge to surface water at locations close to a facility. In both situations, contaminated surface water could be an important source of exposure of individuals for whom projected doses are the highest. Modeling of radionuclide transport in surface water also can be important in analyses conducted to evaluate whether releases to the environment will be ALARA (Section 3.4.2.3), especially when an ALARA analysis takes into account projections of collective dose in an exposed population (DOE, 1999c). Models of radionuclide transport in surface water are reviewed, for example, by Jirka et al. (1983) and Onishi et al. (1981), and models suitable for use in screening analyses are presented in NCRP Report No. 123 (NCRP, 1996a; 1996b). The main difference between NCRP’s screening models and models discussed by Jirka et al. (1983) and Onishi et al. (1981) is that the former do not consider effects of sorption of radionuclides onto sediments and sediment deposition and resuspension on concentrations of radionuclides in water. Transport by attachment to sediments can be

222 / 5. PERFORMANCE ASSESSMENT MODELS important in reducing concentrations of radionuclides in water when the solid/solution distribution coefficient (Kd) is high, as it can be for many chemical elements (Jirka et al., 1983; Onishi et al., 1981). However, sorption onto sediments can cause some radionuclides to be flushed from a surface water body at a slower rate than the rate at which water itself is exchanged, and deposited sediments can provide a long-term source of contamination of a surface water body. Rigorous modeling of radionuclide transport in surface water involves simultaneous solution of coupled time-dependent and three-dimensional advective-diffusion equations with appropriate source and sink terms for all phases of concern (e.g., dissolved radionuclides and radionuclides adsorbed onto sediments). However, such complexity is far beyond the needs of performance assessment, and simplified treatments of surface-water transport usually should be sufficient. For example, it rarely, if ever, should be necessary to consider a time-dependent model or advection and diffusion in three dimensions, and sorption of radionuclides onto sediments often can be ignored. Different models have been developed to describe transport of radionuclides (and other contaminants) in rivers and smaller streams, estuaries, coastal waters, and lakes. Those models differ mainly in their approach to describing dispersion and mixing of radionuclides in water. Given the locations of currently operating low-level waste disposal facilities and likely locations of future facilities, releases to estuaries and coastal waters should not be of concern (i.e., concentrations of radionuclides in such water bodies should be much less than concentrations in surface water into which discharges could occur) and are not considered further. 5.7.5.1 Modeling of Discharges to Surface Water. Discharges of radionuclides to surface water can occur by overland transport or transport in groundwater. Discharges in groundwater can be estimated on the basis of a conservative assumption that all radionuclides distributed in an aquifer in the vicinity of a surface water body are discharged into it (NRC, 2000). With this assumption, the rate of discharge of a radionuclide is given by the product of the concentration in groundwater and the flux of groundwater at that location. The flux of groundwater can be estimated using a one-dimensional stream-tube analysis (Section 5.7.3.4). Discharges in groundwater can be assumed to occur at a point, which should be conservative in some cases, or to be dispersed over a laterallydistributed groundwater plume (NRC, 2000).

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Discharges by overland transport could occur after the cover above disposed waste has been removed by erosion and waste is exposed at the ground surface. Overland transport should be potentially more important at above-grade disposal facilities than at well-designed facilities located below grade, although overland transport that results from the so-called bathtub effect can occur at below-grade facilities that rapidly fill with water and are poorly drained. In modeling overland transport, it is commonly assumed that the concentration of a radionuclide in soil water is determined by the solid/solution distribution coefficient (Kd) in surface soil, and that the annual discharge to a surface water body is given by the product of the radionuclide concentration in soil water and the fraction of the volume of the annual precipitation at the location of a disposal facility that enters the water body by runoff (EPA, 2000c). Such a model assumes that overland runoff of radionuclides attached to soil particles is unimportant, which may not be the case during unusually intense precipitation events. Modeling of discharges of radionuclides attached to soil particles requires an assumption about the amount of soil from the location of a disposal facility that discharges to a water body. Such an assumption could be based on site-specific data on the erosion rate. 5.7.5.2 Modeling of Transport in Rivers and Streams. At many disposal sites, it should be adequate to model transport of radionuclides in rivers and streams on the basis of simplifying assumptions that the activity discharged annually to such a water body is diluted uniformly by the annual flow and that losses from the water column by sedimentation are unimportant (EPA, 2000c; NRC, 2000). With those assumptions, the concentration of a radionuclide at any downstream location is given by the ratio of the annual discharge to the annual flow. If exposure to a contaminated river or stream could occur at a location close to a discharge, an assumption of uniform dilution of annual discharges of radionuclides by the annual flow could result in underestimates of concentrations in water used by a receptor. In such cases, a site-specific model of surface-water transport could be considered. If discharges are assumed to occur along a shoreline, radionuclide concentrations as a function of downstream distance (x) and lateral offshore distance (y) can be estimated by solving a two-dimensional advective-diffusion equation at steady-state. Solutions are given, for example, by Codell et al. (1982), Napier et al. (1988), NCRP (1996a), and NRC (1977a). Lateral and longitudinal dispersion are described by empirical relationships in which

224 / 5. PERFORMANCE ASSESSMENT MODELS dispersion coefficients depend on the depth, width and flow velocity of a river or stream. A useful result of two-dimensional advective-diffusion modeling is that in the absence of removal of radionuclides from the water column by sedimentation, nearly complete and uniform mixing of radionuclides released to a river or stream occurs at downstream distances beyond ~3B2/d, where B is the width of the river and d is the depth (NCRP, 1996a). Use of an annual average width and depth generally is appropriate. This condition should almost always be met in discharges to small streams. Concentrations of radionuclides at the shoreline along the same side of a river where discharges occur (i.e., at y = 0) often are of interest. Those concentrations can be expressed as a function of downstream distance only (NCRP, 1996a). If the downstream distance is less than the minimum distance at which nearly complete and uniform mixing of a discharge occurs, as noted above, a partial mixing correction factor greater than unity can be applied to the concentration obtained by assuming complete and uniform mixing. A graph of this correction factor, which is plotted against the parameter E = 1.5dx/B2, is given in Figure 3.5 of NCRP Report No. 123 (NCRP, 1996a). 5.7.5.3 Modeling of Transport in Lakes. Models of transport of radionuclides in lakes often distinguish between large lakes with a surface area greater than ~400 km2, in which the residence time of water is large and lake flow and mixing are dominated by wind-induced currents, and small lakes, in which concentrations of radionuclides often can be assumed to be uniformly mixed (NCRP, 1996a). When an assumption of uniform mixing in a small lake is reasonable, the concentration of a radionuclide at steady-state can be calculated as Wo/(Q + λV), where Wo is the annual radionuclide discharge to the lake, Q is the annual inflow and outflow of water at the lake, λ is the radionuclide decay constant, and V is the volume of the lake. In larger lakes, concentrations of radionuclides at near-shore locations are of interest, and those concentrations can be much higher than would be obtained by assuming uniform mixing. Concentrations at steady-state can be estimated using an advective-diffusion model similar to the model for transport in rivers and streams (Napier et al., 1988; NRC, 1977b). By taking into account that even in very large lakes, such as the Great Lakes, nearly complete mixing can occur within several weeks of a release, NCRP (1996a) recommended that the concentration of a radionuclide in a large lake can be calculated as the sum of the concentration

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obtained by assuming uniform mixing and a partial mixing term, which is given by the concentration along the centerline of a plume in an unbounded body of water. The partial mixing term can be calculated as WoFSVD/d, where Wo is the annual radionuclide discharge to the lake, d is the depth of the lake, and FSVD is a dispersion factor that decreases with increasing distance along the shoreline from the release point and is plotted in Figure 3.10 of NCRP Report No. 123 (NCRP, 1996a). 5.7.5.4 Modeling of Transport in Sediment. Models described in the previous two sections assume that transport of radionuclides in sediment is unimportant. That assumption usually should be conservative in modeling exposures of individuals who receive the highest doses from discharges of radionuclides to surface water. Effects of attachment of radionuclides to sediments can be described using linear compartment models, in which dissolved radionuclides and radionuclides attached to sediments are assumed to be well mixed in each compartment (Jirka et al., 1983). For example, a lake can be divided into a water layer with an inflow and outflow of water and a settling velocity of sediments and a sediment layer that exchanges with the water layer by first-order kinetics. Models that account for sedimentation in lakes were developed by NRC (1978). More complex compartment models involving a series of mixing tanks can be used to describe the effects of sedimentation in rivers (Onishi et al., 1981). 5.7.6

Summary

Numerous phenomena can play a role in determining the rate of transport of radionuclides in groundwater. In performance assessment, only a few of those phenomena generally are considered to be important, including chemical interactions with soil (sorption), advection, dispersion (including diffusion), and radioactive decay and ingrowth. Phenomena that are considered important in an analysis of transport in groundwater often are represented in stylized ways that attempt to capture their nature without imposing prohibitively difficult data needs or requirements on methods of analysis. Sorption is most often treated as a reversible linear process, which leads to a retardation of radionuclides compared with the flow velocity. Advection is treated at a level of complexity commensurate with a flow analysis that is performed prior to a transport analysis. Complexities of dispersion often are taken into account using a dispersivity function, which is treated as a free parameter within a

226 / 5. PERFORMANCE ASSESSMENT MODELS certain range that cannot be specified precisely. Even decay chains of radionuclides with several members, for which the physical basis is well established, frequently must be treated in an approximate manner in transport analyses, owing to computational difficulties. For a given set of input parameters, the output of an analysis of transport in groundwater is a time history of radionuclide concentrations in groundwater at specified locations. Those concentrations then are used as input to an exposure pathway and dose analysis to produce estimates of annual individual dose for each set of parameters of interest. Transport of radionuclides in surface water may need to be considered in performance assessment, depending on site-specific conditions. Releases of radionuclides to surface water can occur by overland runoff of precipitation or transport in groundwater. At many sites, it should be acceptable to estimate concentrations of radionuclides in surface water by assuming that a discharge is uniformly mixed in the water body, which usually is a river or stream or a lake, and that attachment of radionuclides to sediments is unimportant. Solutions of an advective-diffusion equation at steady-state and methods of modeling transport of radionuclides attached to sediments are available in the literature if greater sophistication is desired. A more complex treatment of advection and diffusion might be warranted, for example, if exposure to water along the shoreline of a lake at shoreline distances close to a release are of concern. 5.8 Atmospheric Transport Analysis In performance assessments of near-surface waste disposal systems, the atmospheric pathway usually has been treated as a secondary means of radionuclide transport and exposure compared with releases to groundwater. There are three principal reasons for this assumption. First, buried waste must be unearthed or transported to the ground surface before atmospheric transport can occur. Second, because precursor events are required to unearth or transport waste to the surface, only a small fraction of the total inventory of buried waste would be subject to atmospheric transport, whereas the entire inventory would be subject to transport in groundwater. Finally, because of significant dilution during atmospheric transport, effects usually are considered to be localized and limited to exposures of inadvertent intruders or a population in the immediate vicinity of a disposal facility. The foregoing arguments notwithstanding, the importance of releases to the atmosphere in regard to exposures of off-site

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members of the public must be considered in performance assessment. Depending on conditions at a disposal site and the radionuclide inventories, releases to the atmosphere can be important compared with other release pathways. For example, at arid sites with a deep vadose zone, diffusion of gases through soils or human or biotic intrusion events may be the most important release mechanisms over long time periods. At any site, radioactive gases may be generated in disposed waste and released to the atmosphere. Finally, contaminated groundwater may discharge to the surface, resulting in seasonally dry sediment deposits or stream beds that could be suspended into the air, or contaminated water may be used to irrigate land, resulting in a distributed source of suspended radionuclides. This Section discusses recommended methods of estimating localized concentrations in air resulting from suspension of particles or gaseous releases and methods of estimating atmospheric transport of radionuclides for use in performance assessment. A summary of issues and concerns related to modeling of atmospheric transport also is included. 5.8.1

Models for Estimating Suspension of Particulates

Several models that can be used to estimate local concentrations in air due to suspension of particulate materials from the ground surface have been developed. These include a modified mass loading model, resuspension factor model, and resuspension rate model. 5.8.1.1 Modified Mass Loading Model. A simple model that is commonly used to estimate local concentrations of suspended particulates in air is a modified mass loading model. Because this model is simple, it is often used as a starting point in screening analyses, with a progression toward more complex models when justified. A mass loading model is particularly appropriate in analyzing concentrations of radionuclides in air and resulting inhalation doses in scenarios for inadvertent human intrusion (Section 6). However, this type of model does not provide a usable input to a model of atmospheric transport. A modified mass loading model is based on an assumption that the concentration of a radionuclide in air above ground is directly related to its concentration in surface soil that provides the source of suspended particulate material at that location. This model can be expressed as (NCRP, 1999):

228 / 5. PERFORMANCE ASSESSMENT MODELS C air = C soil × M air × E f , where: Cair Csoil Mair Ef

= = = =

(5.38)

radionuclide concentration in air (Bq m–3) radionuclide concentration in surface soil (Bq kg–1) concentration of particulate material in air (kg m–3) enhancement factor (dimensionless)

Other formulations of a mass loading model (Kennedy and Strenge, 1992) ignore the enhancement factor (Ef) and assume that it is implicitly included in the atmospheric mass loading factor (Mair). Mass loading and enhancement factors should be measured at each site, but this is often not possible. In many cases, generic values of these parameters obtained under similar conditions are used to estimate local concentrations of particulates in air. The enhancement factor (Ef) is defined as the ratio of a radionuclide concentration in particulate matter in air (Bq kg–1, dry weight) to the concentration in surface soil (Bq kg–1, dry weight). It takes into account that the fraction of contaminated material in surface soil that is suspended into air by wind or mechanical disturbances may not be the same as the fraction of all soil materials. Its value depends on soil type, the distribution of contamination with particle size, and the type of disturbance that causes suspension. In the absence of site-specific data, the following generic values can be used in performance assessment (NCRP, 1999; Shinn, 1991): Ef Ef

= =

0.7 for undisturbed surface soil 4 for recently disturbed soil32

The atmospheric mass loading factor (Mair) for various soils has been determined at different wind speeds and under different conditions of mechanical disturbance. The largest body of relevant data is reported in literature related to air pollution (e.g., DHEW, 1969; Hinton et al., 1986; Lillie, 1972; Stern, 1968). Additional information on specific subjects can be found in the Journal of the Air and Waste Management Association. Typical concentrations of suspended particulate matter in outdoor air can vary over several orders of magnitude. A national average of reported values for suspended soot and ash is ~1 to 32The

higher value applies to suspension of material at the ground surface within a few days after a disturbance.

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2 × 10–7 kg m–3 (Hinton et al., 1986). Other data indicate that average concentrations of suspended particulates are in the range of ~1 to 6 × 10–8 kg m–3 near urban locations and ~0.6 to 2.2 × 10–7 kg m–3 within urban areas, depending on the size of a city and extent of industrial activity (DHEW, 1969). Much higher atmospheric mass loadings can result from high winds, such as dust devils (whirlwinds), and concentrations of ~5 × 10–6 kg m–3 have been reported under such conditions (Sinclair, 1976). In work more closely related to needs of performance assessment, airborne dust at the Bikini Atoll was measured to determine potential inhalation exposures due to resuspension of plutonium-contaminated soils (Shinn et al., 1989). Background dust concentrations were ~2 ×10–8 kg m–3, with concentrations as high as 1.4 × 10–7 kg m–3 associated with tilling a bare field. Thus, a fairly vigorous mechanical disturbance resulted in somewhat higher local air concentrations than wind erosion alone. In other studies, upper and lower limits on airborne mass loadings of soil due to wind erosion as a function of particle size were estimated at the Hanford Site near Richland, Washington (Sehmel, 1975; 1977a; 1984). For particle sizes less than ~1 µm, the upper limit of atmospheric mass loading was ~7 × 10–7 kg m–3, and the upper limit was ~2 × 10–4 kg m–3 for particle sizes larger than 1 µm. Some studies have attempted to determine a long-term average dust loading for purposes of radiation dose assessment. For example, Soldat et al. (1973) determined that an average dust loading of 1 × 10–7 kg m–3 is appropriate in assessing potential environmental impacts of interim, retrievable surface storage of commercial high-level wastes. That value also was suggested by Anspaugh et al. (1975) for use in generic radiation dose assessments. NCRP Report No. 129 (NCRP, 1999) evaluated uncertainties associated with applying a modified mass loading model in a variety of land-use scenarios. For given concentrations of radionuclides in surface soil, consideration of the variability in annual-average mass loading and enrichment factors led to a conclusion that an uncertainty in estimated annual-average airborne concentrations of radionuclides, as represented by the ratio of the 95th percentile to the median, is about a factor of five. Use of annual-average parameter values is appropriate in performance assessments of low-level waste disposal facilities. 5.8.1.2 Resuspension Factor Model. An alternative to a mass loading model that is often used to estimate local air concentrations of radioactive materials that are suspended from surface soil by wind or mechanical disturbances is a resuspension factor model. This

230 / 5. PERFORMANCE ASSESSMENT MODELS model also relates concentrations of radionuclides in air directly to their concentrations in surface soil, and is expressed as (Kathren, 1984): C air = C surf × S f , where: Cair = Csurf = Sf =

(5.39)

radionuclide concentration in air (Bq m–3) radionuclide concentration on ground surface (Bq m–2) resuspension factor (m–1)

A resuspension factor model differs from a mass loading model in that a concentration in air is estimated on the basis of an activity per unit area on the ground surface, rather than an activity per unit mass (volume). A resuspension factor model should be used to estimate local concentrations of particulate material in air only when sufficient site-specific data are available and when it is reasonable to model soil contamination as a surface source rather than a source distributed throughout a volume of soil. As in the case of a mass loading model, a resuspension factor model does not provide a suitable input to a model of atmospheric transport. When contamination is distributed with depth in soil, rather than confined to the surface, an activity per unit area is estimated from an activity per unit volume at the surface using an “effective” depth, which is the depth of surface soil that can be suspended by wind. An effective depth normally is assumed to be in the range of 10–3 to 10–2 m, depending on the type of soil and its characteristics and the mechanism for resuspension. A resuspension factor model also can be modified to account for dilution by air and suspended dust that originates from other locations, but such modifications are somewhat arbitrary in the absence of site-specific data. Resuspension factors measured in numerous studies under a variety of conditions range from ~10–11 to 10–4 m–1 (Sehmel, 1984). This broad range of measured values makes it advisable to use site-specific resuspension factors, rather than generic values obtained from the literature. Several factors influence measured values, including the size of a source area, the uniformity of radionuclide distributions, wind speed, sampling height, soil particle sizes, soil moisture content, soil density, surface roughness, topography, vegetative cover, and temperature (Sehmel, 1984). A summary of reported resuspension factors is given in Table 5.4. As an alternative to using a single resuspension factor to estimate concentrations of resuspended particulates in air, Anspaugh et al. (1975) developed a time-dependent resuspension factor to

TABLE 5.4—Example resuspension factors reported in the literature. Condition and Source

Resuspension Factor (m–1)

Comments

Wind Stress 2 × 10–11 to 8 × 10–9 9 × 10–8 to 1 × 10–7 9 × 10–8 to 5 × 10–7

Bare soil, aqueous 91Y chloride 210 Po as oxide U3O8

Dunaway and White (1974)

1 × 10–9 to 1 × 10–4

Time-dependent model for plutonium in soil

Sehmel (1980)

9 × 10–11 to 3 × 10–4

Literature review

Shinn et al. (1986)

1.8 ×

10–13

to 6.1 ×

10–10

Garland and Pattenden (1990)

Plutonium aerosols at Nevada Test Site Nuclear weapon test debris: 13 y after deposition 22 y after deposition

<2 × 10–9 <5 × 10–10 3.6 × 10–9 to 4.9 × 10–8

Chernobyl 137Cs deposition; initial resuspension factor reduced by factor of 0.23 to 0.64 within the first year

Vegetation Healy and Fuquay (1959) Stewart (1967)

2.9 × 10–8 to 6 × 10–7 3×

10–8

to 5 ×

10–5

Fluorescent powder U3O8

Mechanical Disturbances Stewart (1967)

1.5 × 10–6 to 3 × 10–4

Sehmel (1973)

4.8 × 10–5 to 1.1 × 10–2 10–4

Sehmel (1977b)

1×

Sehmel (1980)

1 × 10–10 to 4 × 10–2

to 2 ×

ZnS, per disturbance ZnS, per disturbance Literature review

/ 231

10–5

Plutonium

5.8 ATMOSPHERIC TRANSPORT ANALYSIS

Stewart (1967)

232 / 5. PERFORMANCE ASSESSMENT MODELS account for the effects of environmental aging, sometimes referred to as weathering, on freshly deposited material. On the basis of observations of resuspension of plutonium from testing of nuclear weapons at the Nevada Test Site over a period of 17 y, primarily by the action of wind, the following resuspension factor was derived: S f ( t ) = 10 where: t Sf 10–4 10–9 λ

= = = = =

–4

e

–λ t

–9

+ 10 ,

(5.40)

time after an initial deposition in days (d) resuspension factor (m–1) resuspension factor at an initial time of deposition (m–1) resuspension factor after 17 y (m–1) 0.15 d–1/2

Although the time-dependent resuspension factor in Equation 5.40 is commonly used in environmental assessments, it may be difficult to justify for use in performance assessment. For example, suspension of particulates may be quite different at sites with more abundant rainfall and vegetation than at the Nevada Test Site and, depending on the exposure scenario, stresses other than wind that can result in substantially higher levels of resuspended material (e.g., mechanical tilling of surface soil) may need to be taken into account. In addition, radionuclides may be mixed over a greater depth of soil or attached to particles of different sizes compared with plutonium in surface soil at the Nevada Test Site. In a review of screening models for use in estimating dose due to contamination of surface soil (NCRP, 1999), an alternative to the time-dependent resuspension factor in Equation 5.40 was developed. On the basis of data obtained during the first two months after a deposition (Garland, 1982) and data obtained between 2 and 42 months after the Chernobyl accident (Garland et al., 1992), which showed a more rapid decline in the resuspension factor with time over a period of a few years than predicted by the model in Equation 5.40, and by assuming that the model in Equation 5.40 described the resuspension factor at long times after deposition, the resuspension factor as a function of time (t) in days after an initial deposition recommended by NCRP (1999) is given by: Sf = 10–6/t(d) m–1 for t ≤ 1,000 d

(5.41)

Sf = 10–9 m–1 for longer times, out to many years

Again, this model is intended to describe resuspension caused mainly by actions of wind.

5.8 ATMOSPHERIC TRANSPORT ANALYSIS

/ 233

Data obtained after the Chernobyl accident also were used by Garger et al. (1997) to investigate the long-term resuspension factor and its uncertainty. Those investigators concluded, first, that the uncertainty in an annual-average resuspension factor is about an order of magnitude when the resuspension factor is determined as a function of time and the predominant regional conditions of vegetative cover and climate and, second, that the decrease in the resuspension factor over a period of ~8 y depended on time as t–1.4. That time-dependence differs from assumptions in other models of resuspension factors described above. A variety of resuspension factor models were tested against Chernobyl data by Garger et al. (1999). It was found that empirical models can give predictions that agree with observations within an order of magnitude or better if models are adequately calibrated for site-specific conditions, but that they do not account for spatial heterogeneity or temporal variations. Available studies of resuspension factor models, such as those summarized above, indicate that there is considerable uncertainty (e.g., probably more than an order of magnitude) in applying models to sites that differ in vegetative cover and climate from sites where data on which the models are based were obtained. Such large uncertainties apply even when the purpose is to estimate annual-average resuspension factors. Consideration of radionuclides of concern also is potentially important. For example, data on resuspension obtained after the Chernobyl accident apply to 137Cs, and the long-term availability of other radionuclides for resuspension may differ substantially from the availability of 137Cs in surface soils near Chernobyl. 5.8.1.3 Resuspension Rate Model. A resuspension rate model is similar to a resuspension factor model, except it provides an estimate of the rate of release of radionuclides into air rather than an estimate of the concentration in air. A resuspension rate model can be written as: Q = C soil × A × d × R r , where: Q Csoil A d Rr

= = = = =

release rate of radionuclide into air (Bq s–1) radionuclide concentration in soil (Bq m–3) area of contaminated soil (m2) depth of resuspendable surface soil (meters) resuspension rate (s–1)

(5.42)

234 / 5. PERFORMANCE ASSESSMENT MODELS A resuspension rate model is particularly appropriate for use in scenarios that require a calculation of radionuclide transport in the atmosphere, because the model provides a release rate (source term) that is required as input to such calculations. The model also can be used to estimate air concentrations at a location of contamination by assuming a mixing height and average wind speed. In general, however, a modified mass loading model discussed in Section 5.8.1.1 is considered more appropriate for use in estimating localized concentrations in air, mainly because such a model can be based directly on measured mass loadings of particulate material in air and an assumption about the mixing height in air is not required. Further discussion of a resuspension rate model is provided by Healy (1980) and Healy et al. (1979). The resuspension rate in Equation 5.42 is based on observation and is often derived from a diffusion model and field measurements (Sehmel, 1984). In numerous studies, resuspension rates have been measured as a function of wind speed and particle size, and results vary from ~10–12 to 10–4 s–1 (Sehmel, 1984). A summary of reported resuspension rates is given in Table 5.5. TABLE 5.5—Example resuspension rates reported in the literature. Resuspension Rate (s–1)

Condition and Source

Comments

Wind Stress 3.3 × 10–8 to 4.4 × 10–8

UO2 powder

1.9 × 10–10 to 1.6 × 10–8

UO2 powder

Mishima and Schwendiman (1972) Vegetation Mishima and Schwendiman (1972)

1 × 10–10 to 1 × 10–8

Sehmel and Lloyd (1976)

3.3 × 10–8

Sehmel (1975) Dunaway and White (1974)

2.6 × 10–12 to 4.7 × 10–10

Calcium molybdate ZnS Plutonium at Nevada Test Site

Mechanical Disturbances Milham et al. (1975) Healy (1971)

1.0 × 10–9 to 3.0 × 10–9

Environmental plutonium

3 × 10–10 to 1 × 10–6

PuO2 on floors

5.8 ATMOSPHERIC TRANSPORT ANALYSIS

5.8.2

/ 235

Release of Gases by Diffusion

Radioactive gases can be generated in a low-level waste disposal facility by a number of mechanisms, including corrosion of metals, degradation of organic materials, and, in the case of radon, radioactive decay (Section 5.4.5.2). Radioactive gases or vapors of concern may include CO2, H2, CH4, H2O, H2S, Rn, 85Kr, and volatile forms of 129I and 79Se. As a gas is generated, it may dissolve in water and be transported to groundwater, or it may be transported upward through the emplaced waste, engineered barriers, and cover to the surface by diffusion processes. Once released to the atmosphere, radioactive gases may result in localized radiation doses, general environmental contamination, or off-site doses following atmospheric transport. For disposal in the unsaturated zone, diffusion is the most likely release mechanism for radioactive gases. The diffusion equation (Equation 5.12 in Section 5.4.5.1.2) normally is solved by assuming an initial uniform distribution of a radioactive gas in a disposal facility and a boundary condition of zero concentration at the surface (i.e., a gas is readily dispersed in the atmosphere once it reaches the surface). The resulting concentration as a function of location and time is used in Equation 5.13 to calculate the flux density [J (Bq m2 s–1)] at the ground surface, from which the release rate [Q (Bq s–1)] is obtained using the known area of the source region. A diffusion model for release of 222Rn can be calibrated by comparing model predictions with a known average background release rate or known concentration in outdoor air relative to an average concentration of its parent radionuclide 226Rn in surface soil. Indeed, in some dose assessments for inadvertent intruders, who are assumed to reside in a home on top of exposed waste or to be located on contaminated surface soil outdoors, data on doses from naturally occurring radon per unit concentration of radium in surface soil are used to estimate doses from exposure to radon, rather than a diffusion model (McDowell-Boyer et al., 2000; ORNL, 1997a). That approach circumvents possible shortcomings of a diffusion model, including its neglect of advective transport as discussed in the following section. If a diffusion model is used to estimate airborne releases of volatile radionuclides, judgment is required in determining a diffusion coefficient for each chemical element. Releases of volatile 3H from a low-level waste disposal facility usually are assumed to be in the form of tritiated water (Kincaid et al., 1995; Maheras et al., 1994; 1997; McDowell-Boyer et al., 2000; ORNL, 1997a). With that

236 / 5. PERFORMANCE ASSESSMENT MODELS assumption, a diffusion model for 3H also needs to take into account the absolute humidity (i.e., the ratio of the density of water vapor in air to the density of water in the liquid phase). That ratio usually is assumed to be ~9 × 10–6 (ORNL, 1997a). The dominant gaseous form of 14C often is assumed to be CO2. With that assumption, a diffusion model for 14C also needs to take into account the relationship between the concentration of CO2 in air and the concentration in liquid water bound in the waste, which can be calculated using Henry’s Law (Cooper and Alley, 2002). The diffusion coefficient of radon in an air-filled porous medium typically is assumed to be ~2 × 10–6 m2 s–1 (Rogers and Nielson, 1991a; Rogers et al., 1984). That value is appropriate for soils with a low moisture content, and it is conservative for soils with a relatively high moisture content. 5.8.3

Advective Transport

In addition to diffusion of gaseous radioactive materials from buried waste, evidence of advective flow of gases through soils has been reported. Most of the available information is concerned with movement of 222Rn through soils and into buildings (Nazaroff, 1988a; 1988b; 1992; Rogers and Nielson, 1991b). Advective flow of gases is driven by the weather including wind, temperature, rainfall, and changes in barometric pressure. Darcy’s Law is used to describe advective transport of a fluid through a porous medium with laminar flow. By assuming that resistance to flow is dominated by the viscosity of the fluid, Darcy’s Law may be written in differential form as (Nazaroff, 1992): k v = – ------ ∇P , µ where: v k µ P

= = = =

(5.43)

superficial velocity vector intrinsic permeability of soil dynamic viscosity of fluid dynamic or disturbance pressure

The superficial velocity vector (v) is the volumetric flow rate per unit cross-sectional area averaged over a region that is large relative to individual pores but small relative to the overall dimensions of the soil, and the dynamic or disturbance pressure (P) is the total pressure minus the hydrostatic component. Advective migration of radon through air in soil pores has been the subject of a number of studies (Clements and Wilkening, 1974;

5.8 ATMOSPHERIC TRANSPORT ANALYSIS

/ 237

Kraner et al., 1964; Rogers et al., 1983). For example, a study of radon exhalation from uncovered soil in New Mexico (Clements and Wilkening, 1974) found that barometric pressure changes of 1 to 2 kPa over a period of 1 to 2 d resulted in advective velocities on the order of 10–6 m s–1 at the surface of soil with a permeability of 10–12 m2. As a result, the flux of radon at the surface changed by 20 to 60 % compared with the flux associated with molecular diffusion alone. Similar effects have been reported by other investigators. However, effects are limited to periods of changing weather (i.e., changing atmospheric pressure). Field studies of uncovered soils suggest that, in the long term, molecular diffusion is more important than advection in causing release of radon to the atmosphere. However, the reverse is often true for releases into buildings, due to heating and ventilation of indoor air, unless a fractured system extends to the surface. Potential effects of advective transport should be considered in performance assessments when significant sources of radioactive gases may occur. Such effects could enhance releases to the atmosphere from a waste disposal site predicted by a diffusion model, or enhance indoor air concentrations in scenarios for inadvertent intrusion that consider houses constructed on, or in close proximity to, buried waste (Section 6). 5.8.4

Atmospheric Transport Models

Once radioactive materials are released into the atmosphere, they may be transported downwind. Atmospheric transport models are used to predict concentrations of radionuclides in air at specific locations in the vicinity of a given release. Numerous references contain information on atmospheric transport mechanisms, models and data (e.g., Brenk et al., 1983; Eisenbud, 1987; Kathren, 1984; NCRP, 1984a; 1996a; Randerson, 1984; Slade, 1968). Selection of an atmospheric transport model for use in performance assessments should be based on characteristics of the source, the assumed release and exposure scenario, and the availability of meteorological data. Relatively simple models are appropriate for most prospective assessments, because releases and receptors are both hypothetical. More complex models are needed if, for example, receptors are located near a source or there are significant terrain features (e.g., mountains or valleys) between the source and receptor locations. The simplest method of describing atmospheric transport, which is suitable for screening analyses to assess the potential importance of the air pathway to exposures beyond the boundary of

238 / 5. PERFORMANCE ASSESSMENT MODELS a disposal site, is to assume that releases occur from a point source and that the concentration in air at a receptor location is equal to the concentration in a discharged volume of air at the point of release, corrected for the fraction of the time that the wind blows toward the receptor. The method assumes that a receptor is close to the release and that complex terrain and other conditions are unimportant, and it is described by the following equation (NCRP, 1996a): fQ C air = ---------- , V where: Cair = f

=

Q V

= =

(5.44)

average concentration of radionuclide in air at receptor location (Bq m–3) fraction of time that wind blows toward receptor (dimensionless) release rate of radionuclide to atmosphere (Bq s–1) volumetric flow rate of air at release location (m3 s–1)

As an alternative to the simple and conservative model in Equation 5.44, the Gaussian plume model, which is the most widely used model of atmospheric transport, can be used. That model takes into account dispersion of a release in directions perpendicular to the wind direction. For a continuous release at a constant rate from a point source, the concentration as a function of location, denoted by the coordinates (x,y,z), can be written as (NCRP, 1984a):

Q C air ( x,y,z ) = ------------------------- e 2πσ y σ z u

where: Cair = Q

=

u = σy, σz = he

=

–y2 ------------2 2σ y

e

–z – he2 -------------------2 2σ z

+e

–z + he2 --------------------2 2σ z

,

(5.45)

concentration of radionuclide in air at steady-state (Bq m–3) rate of continuous release of radionuclide to atmosphere (Bq s–1) mean (horizontal) wind speed in x-direction (m s–1) horizontal (lateral) and vertical dispersion parameter (meters) effective source height (meters)

5.8 ATMOSPHERIC TRANSPORT ANALYSIS

/ 239

The formulation in Equation 5.45 accounts for “reflection” of a plume by the ground surface through use of a double exponential term in the height above ground (z). The horizontal and vertical dispersion parameters (σy and σz) are empirically-based functions of downwind distance (x), atmospheric stability conditions, source height, and surface roughness; suitable representations are given by Barr and Clements (1984) and Miller (1984). Simplifications of Equation 5.45 often are appropriate at low-level waste disposal sites. By assuming that the buoyancy of released material is unimportant [i.e., that the effective source height (he) is equal to the height of the top of the cover above the elevation of the local terrain] and that the difference in elevations at a nearby receptor location and the location of disposed waste is much smaller than the vertical dispersion parameter (σz), the sum of the exponential terms in z reduces to the value two. Then, by averaging radionuclide concentrations in air over the width of a 22.5 degree sector in a standard wind rose and taking into account the fraction of the time the wind blows in the direction of a receptor (f), Equation 5.45 reduces to (Miller, 1984):

2.032 fQ C air ( x ) = ------------------------- . xuσ z

(5.46)

The point-source model in Equation 5.45 or 5.46 can be applied to distributed sources as long as the distance from a source to a receptor is much greater than the longitudinal or lateral dimensions of the source. At low-level waste disposal sites, however, the distance to a nearby receptor may be comparable to, or even less than, the dimensions of the area over which waste is emplaced. In such cases, an area source model of atmospheric transport should be used. Atmospheric transport from an area source can be modeled by integrating Equation 5.45 over the longitudinal (upwind) and lateral (crosswind) dimensions of the source region and replacing the emission rate (Q) by an emission rate per unit area (QA) which normally would be assumed to be constant over the source region. The integral in the crosswind (y) direction is given by erfc(y/σy), where erfc is the complementary error function. The integral in the upwind (x) direction generally must be approximated using numerical methods. Further discussion of area source modeling is given in EPA (1995b).

240 / 5. PERFORMANCE ASSESSMENT MODELS 5.8.5

Summary

In performance assessments of low-level waste disposal facilities, the atmospheric release and transport pathway can be considered in relation to exposures of inadvertent intruders, where localized air concentrations are potentially important, standards for airborne emissions of radionuclides (EPA, 1989a), and dose assessments for off-site individuals that take into account all release and exposure pathways. In most analyses of off-site exposures, however, screening calculations can be used to show that the atmospheric pathway is unimportant. As a result, most analyses of atmospheric release and transport use generic data obtained from the literature. Suspension of particulate radionuclides from surface soil into air normally is expected to be of greatest importance in dose assessments for inadvertent intruders. In such assessments, use of a simple mass loading model is appropriate. That model is based on measurements of airborne concentrations of naturally occurring radionuclides, including uranium and thorium, relative to their concentrations in surface soil. Alternatively, a resuspension factor model could be used, although available data indicate that resuspension factors vary by many orders of magnitude depending on assumed human activities and site conditions. More complex models that require assumptions about local meteorological or soil conditions, such as a resuspension rate model, are not recommended for use in dose assessments for inadvertent intruders. On the other hand, if suspension of particulate radionuclides is important in a dose assessment for off-site individuals, a resuspension rate model may be suitable because it provides an appropriate input to an atmospheric transport model. However, a resuspension rate that would be appropriate at any site is likely to be quite uncertain. If atmospheric transport of radionuclides to off-site locations is considered, use of a Gaussian plume model at steady-state usually is appropriate. Transport from a point or area source can be modeled, depending on the distance to a receptor location relative to the surface dimensions of a disposal facility. 5.9 Biotic Transport In regard to disposal of low-level waste, the term “biotic transport” refers to actions of plants or animals that affect transport of waste materials from their location in a disposal facility to a location where radionuclides can enter into human exposure pathways, such as a terrestrial foodchain pathway (Kennedy et al., 1985). The potential importance of biotic transport and resulting human

5.9 BIOTIC TRANSPORT

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exposures at a specific site will depend on the site location, design of the disposal facility, and waste forms. Biotic transport normally is considered to be of secondary importance in performance assessments, mainly because of a lack of relevant site-specific data. Consequently, biotic processes normally are not modeled rigorously (Cook and Fowler, 1992; Kincaid et al., 1995; McDowell-Boyer et al., 2000; ORNL, 1997a). However, a complete analysis of specific disposal sites should at least include a qualitative consideration of potential impacts of biotic transport. There are numerous published studies of biotic transport and exposure mechanisms. However, most studies are related to historical low-level waste disposal practices and may not apply to highly engineered disposal systems. Current designs attempt to minimize the importance of intrusion by plants and animals by increasing the depth of disposal, using engineered cover systems, and using concrete or steel in vaults, waste containers, or waste forms. As a consequence of an inability to fully satisfy concerns about potential impacts of biotic transport, some attempts have been made to include a qualitative analysis in performance assessment (Maheras et al., 1994; 1997; ORNL, 1997a). This Section provides an overview of biotic transport considerations in performance assessments of low-level waste disposal systems. 5.9.1

Background

In-depth studies of movement of radionuclides through biotic processes have not been conducted at low-level waste disposal sites, although a number of cases of biotic transport have been observed (Fitzner et al., 1979; Landeen and Mitchell, 1981; Wallace et al., 1978; Winsor and Whicker, 1980). Laboratory and field observations at DOE sites demonstrated that biotic transport processes have occurred at historical disposal sites. What is not well established or assessed is the long-term importance of biotic transport processes, their potential importance at contemporary sites (i.e., effects potentially linked to specific biotic communities), their potential effects on engineered waste forms and disposal facilities, and the significance of biotic transport on exposure of humans. The most comprehensive evaluation of the potential importance of biotic transport on exposure of humans was conducted by McKenzie et al. (1982a; 1982b; 1983; 1984; 1985; 1986). Mechanisms of biotic transport and human exposure were identified and discussed, and generic assessments at arid and moist sites were conducted. From an initial qualitative assessment, McKenzie et al. (1982a) concluded that penetration of buried waste and active transport by burrowing mammals, invertebrates, and plant roots is

242 / 5. PERFORMANCE ASSESSMENT MODELS potentially the most important biotic transport mechanism in cases of disposal in trenches. Cataldo et al. (1987) noted that native vegetation has the potential to modify soil chemistry over the long term and, consequently, the mobility of radionuclides. Enhanced degradation of engineered barriers also could be a concern, but there are limited long-term data to quantify and evaluate potential effects. Intrusion and active transport at trench disposal sites can result in removal of small quantities of waste or contaminated soil through surface barriers to soils that overlie burial trenches. At historical disposal sites with minimal covers, accumulated material that is transported to the surface by independent biotic intrusion events combined over hundreds of years could lead to radiation doses at arid and moist disposal sites similar to those potentially received by human intruders. Modifications of waste forms or disposal facilities would likely change impacts on trench disposal by delaying, reducing and possibly eliminating biotic intrusion events. Thus, biotic intrusion should be less of a concern at engineered facilities than at historical disposal sites. However, biotic processes may still be important in determining the durability of engineered barriers and cover systems located near the ground surface and in enhancing mobilization of radionuclides as a result of plant rhizosphere processes. Owing to uncertainty associated with cumulative, long-term effects of biotic transport processes, there is not yet a consensus regarding scenarios of concern to performance assessment and appropriate methods of analysis. 5.9.2

Biotic Transport Processes

Three different mechanisms of biotic transport can occur at a near-surface disposal facility: (1) transport enhancement, (2) intrusion and active transport, and (3) secondary transport (McKenzie et al., 1982a). Transport enhancement occurs when plants or animals modify buried waste or engineered barriers and waste forms so that there is an increased potential for radionuclide transport. An example of transport enhancement is the formation of tunnels through barriers by burrowing animals or plant roots that can provide conduits for release of gases or enhanced infiltration of water. Another example is the formation of organic ligands in the root zone of plants, which may lead to incorporation of radionuclides in soluble organic complexes (Cataldo et al., 1987; Schilk et al., 1996). Microbial degradation of waste forms also could enhance generation of gases or create waste materials that are more soluble in infiltrating water. Intrusion and active transport occurs when biota

5.9 BIOTIC TRANSPORT

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penetrate a waste zone and cause a redistribution of waste material or contaminated soil. An example of intrusion and active transport is penetration of cover materials by deep-rooted plants and root uptake of radionuclides. In secondary transport, radionuclides are available to biota for additional displacement after they have been mobilized by other processes. An example is redistribution by plant leaves or fruit following intrusion into surface soils by tree roots. Depending on the disposal system and environmental setting, long-term effects of all three transport mechanisms could influence the performance of a disposal system and lead to significant pathways of human exposure. 5.9.2.1 Transport Enhancement. Radionuclide transport at lowlevel waste disposal sites can be enhanced over time by interactions of biota with man-made barriers, the waste environment, or waste itself (McKenzie et al., 1982a). Evidence of transport enhancement has been reported at closed DOE disposal sites (Arthur, 1982; Cornam, 1979; Fitzner et al., 1979; Hakonson and Bostick, 1976; Landeen and Mitchell, 1981; O’Farrell and Gilbert, 1975). A number of mechanisms of transport enhancement are possible. As noted previously, plant root exudates have the potential to enhance radionuclide transport. Burrowing mammals and invertebrates in soil or a cover system can form series of tunnels or chambers at varying depths. Decomposing plant roots also may result in a series of channels. Such tunnels or channels may promote escape of gases, increase infiltration of water to emplaced waste, and enhance erosion by wind or water. Studies have reported that pocket gophers, prairie dogs, and ants or termites can be part of natural biotic processes that occur in soil backfill and natural barriers placed above waste disposal trenches. Depending on the waste form, waste could enter decomposer food webs when buried materials include decomposable organic matter (e.g., paper products, wood, plant and animal tissues, or cloth). In near-surface soils, earthworms have been observed to significantly influence the rate of breakdown of organic material in soil, thus enhancing the availability of 137Cs to plants (Reichle et al., 1971). 5.9.2.2 Intrusion and Active Transport. Intrusion and active transport processes have been identified as the most likely to occur at low-level waste disposal sites, and are most commonly considered in assessments of radionuclide transport (McKenzie et al., 1982a). In this type of process, plant roots or burrowing animals access waste or a contaminated soil zone, and radioactive material is translocated to the ground surface, where it can enter other

244 / 5. PERFORMANCE ASSESSMENT MODELS transport and human exposure pathways. Numerous examples of processes that are involved in intrusion and active transport, where plants, small mammals, and invertebrates (insects) serve as environmental vectors to intrude into and transport buried wastes, are discussed in the literature. Burrowing animals transport bulk materials independently of their radionuclide composition, whereas plant roots can preferentially mobilize and transport biologically available radionuclides. Intrusion and active transport processes normally occur in the top 3 m of soil (Kennedy et al., 1985), although some large plants may have greater root depths. At western sites, for example, Russian thistle (tumbleweed) may grow in materials that comprise the cover of a disposal trench. That plant is an annual that inhabits disturbed desert sites, and it has a deep root system (up to 5 m). In autumn, the mature plant dries and detaches at the surface, at which time it is blown by wind until it encounters an obstruction. Thus, those plants can cause both vertical and horizontal movement of radioactive materials from a disposal site. An evaluation of the importance of intrusion and active transport processes at arid western and moist eastern disposal sites in the United States is given by Kennedy et al. (1985). 5.9.2.3 Secondary Transport. After radionuclides have been mobilized by such other means as groundwater seeps or overflow of water that has accumulated in burial trenches, actions of biota may cause further displacement in the surface environment. Displacement mechanisms could involve direct actions of mammals or invertebrates (e.g., ants or bees) and root uptake by plants. Secondary transport processes have been reported at closed low-level waste disposal sites, but they may be unimportant in cases of highly engineered waste disposal systems. 5.9.3

Pathways of Human Exposure

Intrusion and active transport and secondary transport processes can result in contamination of plant and animal species near a waste disposal site. Some of those species may directly enter the human foodchain. For example, deer, quail, rabbits, and other species that are hunted for food may ingest contaminated forage near a disposal site. Management of low-level waste disposal sites, including activities to limit vegetative growth and intrusion by mammals, may effectively eliminate foodchain pathways during a period of institutional control. However, the relative importance of those processes over long time periods after loss of institutional control, compared

5.10 EXPOSURE PATHWAYS AND RADIOLOGICAL IMPACTS

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with processes that are controlled by hydrology and geochemistry, will depend on the site, the design of a disposal facility and closure system, and waste characteristics. Chemical elements with relatively high bioaccumulation in plant and animal species have the potential for continual migration and trophic level transfer and cycling. Once radionuclides enter the environment, additional pathways may become important, including inhalation of suspended dust and ingestion of crops grown in contaminated soil. In developing exposure scenarios for use in performance assessment, the long-term potential for exposure pathways following biotic transport in the surface environment should be considered for such radionuclides as 14C, 90Sr, 99Tc, 129I, and 137Cs. 5.9.4

Summary

Biotic processes and a variety of associated effects of potential importance to performance assessments of low-level waste disposal facilities have been observed at historical disposal sites. Those processes can involve (1) enhancement of radionuclide transport in air and water through long-term degradation of engineered barriers, (2) active transport of materials through actions of plants or burrowing animals, and (3) additional displacement of radionuclides after other transport processes have occurred. However, it may be difficult to apply those observations at other disposal sites and to evaluate their implications with respect to enhancing potential exposures of humans. Biotic transport processes normally are considered to be of secondary importance compared with transport in groundwater or effects of inadvertent human intrusion, and there is no consensus on how to model those processes and their potential effects on human exposures quantitatively. The potential for, and importance of, biotic processes should be evaluated on a site-specific basis consistent with characteristics of an engineered disposal facility and the biotic systems that are expected to be encountered. 5.10 Exposure Pathways and Radiological Impacts 5.10.1 Introduction The last steps in performance assessments of near-surface disposal facilities for low-level waste involve evaluations of potential pathways of radiation exposure of humans and impacts (doses or health risks) resulting from such exposures (Figures 5.3 and 5.4). Exposure pathways of concern involve ingestion or inhalation of radionuclides and external exposure to radionuclides in air, water or soil.

246 / 5. PERFORMANCE ASSESSMENT MODELS Of all the aspects of performance assessment depicted in Figure 5.4, modeling of exposure pathways under conditions that occur in a variety of environments is the most mature and is supported by the greatest amount of relevant data as a result of many decades of studies of the behavior of radionuclides in the environment including, for example, studies of fallout from atmospheric testing of nuclear weapons, routine releases from operating nuclear facilities, and accidental releases such as occurred following the Chernobyl accident and at sites in the former Soviet Union where liquid high-level radioactive wastes were stored. The types of models and databases that have been developed as a result of those studies are widely documented including, for example, in reports by IAEA (1994c; 2001) and NCRP (1984a; 1999). Furthermore, as discussed in this Section, simple multiplicative-chain models of exposure pathways that relate radionuclide concentrations in the environment to dose or health risk by means of parameters that are constant in time are adequate for purposes of performance assessment, and there are no major issues in how exposure pathways should be modeled. Rather, issues that arise in modeling exposure pathways in performance assessment generally involve selection of appropriate pathways and parameter values at specific sites. A large number of exposure pathways may need to be considered in performance assessment. For example, the many pathways resulting from releases to air, surface soil, surface water, and groundwater that were assumed in an assessment of a low-level waste disposal facility at DOE’s Savannah River Site (Cook and Hunt, 1994) are depicted in Figure 5.19. However, experience has shown that only a few exposure pathways normally should be important at any site. These pathways usually can be identified using screening techniques. Given the performance objectives for low-level waste disposal discussed in Section 3.4.2, one of the desired endpoints of performance assessment normally is an estimate of the maximum annual dose to off-site members of the public. An estimate of dose to inadvertent intruders onto disposal sites also is a desired endpoint in many assessments. In estimating dose to an individual, the first step is to define assumed exposure scenarios, which are descriptions of conditions that are assumed to result in exposure (e.g., use of contaminated groundwater as a source of drinking water). Then, for each exposure scenario, assumed exposure pathways must be specified, and models developed and implemented for the purpose of estimating dose from each pathway. Finally, doses from all exposure pathways must be summed.

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Fig. 5.19. Potential pathways of exposure of humans to buried low-level waste considered in performance assessment at DOE’s Savannah River Site (Cook and Hunt, 1994).

248 / 5. PERFORMANCE ASSESSMENT MODELS In performance assessments of low-level waste disposal facilities, it generally is satisfactory to assume that exposure scenarios, and their associated exposure pathways, occur with a probability of unity. An important challenge in performance assessment is to develop scenarios that represent credible exposure situations at particular disposal sites. For purposes of performance assessment, the adequacy of assumed exposure scenarios at any site should be judged with respect to their relevance to demonstrating compliance with performance objectives. Exposure scenarios that are conservatively biased to some extent (e.g., an assumption that a member of the public obtains water for domestic use from a well near the boundary of a disposal site) often are appropriate, to help ensure that projected doses at future times would not be substantially less than maximum doses that might actually occur. However, worstcase scenarios need not be assumed, especially if they are unlikely to occur. As in all aspects of performance assessment, the challenge is to develop an appropriate balance between realism and conservatism in modeling exposure pathways at specific sites, as discussed in general terms in Section 2.4. Required inputs to models for estimating dose to individuals from assumed exposure pathways generally are radionuclide concentrations in environmental media (water, soil, air), including concentrations in a disposal facility in dose assessments for inadvertent intruders. For off-site individuals, required concentrations are the maximum values in any year after disposal and at any location beyond the boundary of a disposal facility. For inadvertent intruders, required concentrations are the maximum values in a disposal facility in any year after a presumed loss of institutional control over a disposal site when the assumed exposure scenarios would be credible. Maximum concentrations of radionuclides in any year are the appropriate quantities when performance objectives and other performance measures are expressed in terms of an annual dose. Given the half-lives of radionuclides of potential concern, consideration of variations in concentrations over time periods less than a year is not required. Radionuclide concentrations used as input to a dose assessment are obtained from assessments of the performance of a disposal facility that account for inventories in disposed waste, releases from the facility, and transport through the environment or access to disposed waste by inadvertent intruders. The following sections discuss suitable approaches to developing exposure scenarios and exposure pathway and dose assessment models for use in performance assessment, including recommendations on general approaches, exposure scenarios and associated

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exposure pathways, models of exposure pathways, and selection of parameter values. These discussions are concerned primarily with exposure scenarios and exposure pathway models for off-site members of the public. However, discussions on exposure pathway models usually apply to dose assessments for inadvertent intruders as well. Development of exposure scenarios for inadvertent intruders is treated as a separate problem in this Report (Section 6). Approaches to estimating the risk of stochastic health effects, primarily cancers, resulting from radiation exposure also are discussed. Estimates of risk are not required in demonstrating compliance with performance objectives that are expressed in terms of dose. However, risk is a useful endpoint of an assessment if, for example, potential impacts from radioactive waste disposal are to be compared with those from disposal of hazardous chemical wastes. 5.10.2 General Recommendations Exposure scenarios and exposure pathway and dose assessment models for use in performance assessment should be developed on the basis of the following considerations. 1. Exposure scenarios and exposure pathway models should be consistent with environmental conditions and living habits at a disposal site of concern. Thus, it is inappropriate to use standard assumptions about exposure scenarios, exposure pathways, pathway models, and model parameter values at all sites. A site-specific approach permits an elucidation of potentially important differences in exposure conditions among different sites and, thus, assists in site selection and development of site-specific waste acceptance criteria. 2. The previous recommendation notwithstanding, use of standard assumptions at most disposal sites is warranted for some aspects of a dose assessment including, for example, use of generic dose coefficients for internal and external exposure of reference individuals and standard intake rates of air, water and various foodstuffs by humans and livestock. 3. Assumptions about exposure pathways, pathway models, and model parameter values in dose assessments for off-site members of the public and inadvertent intruders at a particular site should be consistent, to the extent reasonable. 4. The responsibility for development and application of exposure scenarios should rest with performance analysts at

250 / 5. PERFORMANCE ASSESSMENT MODELS each disposal site. Analysts must provide adequate justifications for exposure scenarios, exposure pathways, pathway models, and model parameter values assumed in a dose assessment. These general recommendations are illustrated in more detail in discussions throughout the remainder of Section 5.10. 5.10.3 Exposure Scenarios for Off-Site Members of the Public This Section discusses recommendations on exposure scenarios that should be used in assessing dose to members of the public who are assumed to reside outside the boundary of a low-level waste disposal facility. These recommendations apply without regard for the types of processes or events that are assumed to result in releases of radionuclides beyond a facility boundary. The distinction between off-site members of the public and inadvertent intruders and the types of exposure scenarios for inadvertent intruders that would be appropriate for use in performance assessment are discussed in Section 6. 5.10.3.1 Definition of Environmental Conditions and Living Habits. Development of exposure scenarios for off-site members of the public is based to a significant extent on assumptions about environmental conditions and living habits of residents near a disposal site. Off-site exposures usually are projected to occur at times far into the future. Thus, an important issue is whether current environmental conditions and living habits may be assumed at any time, or whether possible changes should be taken into account. For example, climate changes or inevitable social and cultural changes could have a significant effect on credible exposure scenarios. In performance assessments for low-level waste disposal facilities, it generally is considered acceptable to base assumed exposure scenarios on current environmental conditions and living habits of residents near a disposal site. This approach has been recommended, for example, by NRC staff (NRC, 2000), and it is specified in regulations for disposal of high-level radioactive wastes at the Yucca Mountain Site in Nevada (EPA, 2001a). A rationale for this approach is that possible effects of changes in environmental conditions and living habits over time should be unimportant in developing an understanding of the ability of a disposal system to contain waste, and it is this understanding that is the most important consideration in selecting disposal sites and facility designs and in developing waste acceptance criteria. In the remaining

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discussions in Section 5.10.3, it is assumed that exposure scenarios would be developed on the basis of current environmental conditions and living habits of residents near a disposal site. 5.10.3.2 Exposure Scenarios for Different Release and Transport Pathways. At many disposal sites, release and transport of radionuclides to locations beyond the site boundary is expected to occur primarily by the groundwater pathway. Other pathways could be important, depending on site conditions and the disposal technology. At disposal sites where groundwater discharges to the surface within the site boundary, off-site transport could occur primarily by the surface water pathway. Transport to surface waters by erosion or runoff could be important when above-grade disposal is used. Releases to the atmosphere could be important for such volatile radionuclides as 3H, 14C, and radon. The air pathway also could be important even for radionuclides in particulate form (e.g., at sites where uncovering of waste by erosion of cover material could occur or at arid sites where radionuclides could be transported to the ground surface by evapotranspiration). General recommendations on exposure scenarios for off-site members of the public for different release pathways of concern that should be considered in performance assessment are summarized in Table 5.6. Given the national interest in protecting groundwater resources (EPA, 1991b), use of contaminated groundwater as a domestic water supply usually should be considered even if groundwater at or near a disposal site is not being used for that purpose at the time a facility is developed. However, use of groundwater or surface water for irrigation need not be considered if irrigation is not commonly practiced near a site. Although many scenarios for exposure of off-site members of the public may be of concern, the need to consider a particular exposure scenario does not necessarily imply that exposure pathway models must be developed and quantitative estimates of dose obtained in all cases. Depending on characteristics of waste, a disposal facility, and a disposal site, certain scenarios can be considered only qualitatively or semi-quantitatively (e.g., in comparison with other scenarios that are evaluated) but then excluded, with proper justification, from a quantitative dose assessment whenever they are unimportant. Screening methods also can be used to exclude exposure scenarios or exposure pathways from further consideration. Screening is important to efficient conduct of performance assessment when several scenarios and many exposure pathways could occur at a site. However, it is difficult to develop rigorous approaches to

252 / 5. PERFORMANCE ASSESSMENT MODELS TABLE 5.6—Exposure scenarios for different release pathways that should be considered in performance assessments of low-level waste disposal facilities. Release Pathwaya

Description of Exposure Scenariosb

Groundwater

Use of contaminated groundwater as: • domestic water supply; • water supply for agricultural purposes, including watering of livestock, irrigation of vegetable garden, and irrigation of pasture grass consumed by livestock.

Surface water

Scenarios involving use of contaminated groundwater, plus use of contaminated surface water: • as source of aquatic foodstuffs; • for recreational purposes (swimming and boating).

Atmosphere

Inhalation and external exposure to airborne radionuclides. Exposure to radionuclides deposited on ground surface, including: • external exposure to deposited activity; • ingestion of activity incorporated into foodstuffs (vegetables, meat and milk); • inhalation of activity resuspended into air.

aParticular release pathways, especially releases to surface water and atmosphere, should be considered only if they are potentially important at disposal site. bParticular exposure scenarios should be implemented and doses estimated only if scenarios are credible under current environmental conditions at disposal site. Screening methods may be used to indicate that certain exposure scenarios or associated exposure pathways are unimportant.

screening that would apply at all disposal sites. Rather, experience gained in previous analyses at similar sites can be used to select exposure scenarios and pathways for further analysis. Nonetheless, unimportant exposure scenarios and pathways may need to be analyzed in some cases to satisfy various groups with an interest in site-selection and licensing decisions. 5.10.4 Exposure Pathway Models 5.10.4.1 General Considerations. Selection of an exposure scenario leads to an assumption of particular exposure pathways, such as

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those indicated in scenario descriptions in Table 5.6. For each exposure pathway, a model must be developed to relate radionuclide concentrations in a contaminated environmental medium of concern, often referred to as the source compartment, to the desired endpoint of an assessment, which normally is an estimate of annual dose. Exposure pathway models and associated databases commonly used in performance assessment have been developed mainly from two sources of information. The first is studies of the environmental behavior of naturally occurring radionuclides and trace elements and radionuclides released to the environment by atmospheric testing of nuclear weapons or routine operations and accidents at nuclear facilities, and the second is controlled field or laboratory studies (e.g., IAEA, 1994c; 2001; NCRP, 1984a; 1999). Indeed, modeling of exposure pathways is the only component of performance assessment for which the required data can be based directly on environmental measurements, and there have been extensive international programs to develop, compare and validate models for transport of radionuclides in the biosphere and exposure pathways, including IAEA’s Biosphere Modelling and Assessment Methods (BIOMASS) and Validation of Environmental Model Predictions (VAMP) programs, the Biospheric Model Validation Study (BIOMOVS) program organized by the Swedish Radiation Protection Authority, and the BIOMOVS II program sponsored by the Swedish Radiation Protection Authority and organizations in Canada and Spain (Thiessen et al., 1997; SENES, 2005). However, use of that information in performance assessment involves an implicit assumption that chemical forms of radionuclides released to the environment from a low-level waste disposal facility will be similar to chemical forms of radionuclides or trace elements in studies used to develop exposure pathway models and databases. This assumption and its potential effects on exposure to radionuclides released from a low-level waste disposal facility have not been investigated. 5.10.4.1.1 Components of exposure pathway models. An exposure pathway model used in performance assessment usually provides an estimate of individual dose (most often annual dose) per unit concentration of a radionuclide in a contaminated environmental medium (source compartment). For intakes of radionuclides by ingestion or inhalation, such a model generally has three components, which are referred to as a transfer factor, a usage factor, and a dose coefficient. For external exposure, transfer factors are not relevant, and an exposure pathway model generally includes a

254 / 5. PERFORMANCE ASSESSMENT MODELS usage factor (residence time) and a dose coefficient. These model components are described in Table 5.7. Particularly for terrestrial foodchain pathways, a transfer factor may depend on several parameters. For example, in a pathway that involves consumption of milk obtained from dairy cattle that graze on pasture grass that is irrigated with contaminated water, transfer of radionuclides from contaminated water to an exposed individual depends on the extent of retention of radionuclides on plant surfaces and in the soil root zone, transfer of radionuclides from contaminated soil to pasture grass, and transfer of radionuclides from contaminated pasture grass consumed by livestock to milk or meat (IAEA, 1994c; 2001; NCRP, 1984a). If the desired endpoint of a performance assessment is an estimate of risk rather than dose, a risk coefficient would be included in an exposure pathway model. Approaches to estimating risk that could be used in performance assessment are discussed in Section 5.10.8. TABLE 5.7—Components of exposure pathway and dose assessment models. Model Component

Transfer factor

Description

Factor describing transfer of a radionuclide from a contaminated environmental medium (e.g., air, water or soil) to an exposed individual (e.g., through a terrestrial or aquatic foodchain). Not relevant for external exposure.

Usage factor

Intake of a contaminated material (e.g., foodstuffs, water, air or soil) by humans or livestock. Residence time for external exposure.

Dose coefficient

For ingestion or inhalation, committed effective dose equivalent or committed effective dose per unit activity intake of a radionuclide. For external exposure, effective dose-equivalent rate or effective dose rate per unit activity concentration of a radionuclide in an environmental medium (e.g., air, water, soil or ground surface), including consideration of shielding factors during indoor residence.

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5.10.4.1.2 Multiplicative-chain models. So-called multiplicativechain models normally are used to evaluate exposure pathways in performance assessment. In this type of model, the output (e.g., annual dose) is expressed as a simple product of independent parameters (i.e., relevant transfer factors, usage factors, and dose coefficients). A multiplicative-chain model of an exposure pathway is based on an assumption that concentrations of radionuclides are at equilibrium in all compartments between a source and a receptor [i.e., that concentrations in different compartments that describe an exposure pathway are constant multiples of concentrations in a source compartment (air, water or soil)]. For example, in a pathway that involves consumption of contaminated milk described in the previous section, radionuclide concentrations in pasture grass, cows, and milk are modeled as constant multiples of the concentration in water that is used to irrigate pasture. Radioactive decay during transfer from a source compartment to a receptor (e.g., during the time between production and consumption of milk) generally is unimportant in performance assessment, because all radionuclides that could be released from a disposal facility are sufficiently long-lived. An assumption that concentrations of radionuclides in different compartments that comprise an exposure pathway are constant multiples of concentrations in a source compartment is equivalent to assuming that the model is linear. This assumption is believed to be appropriate for trace levels of radionuclides in the environment, and it should provide conservative estimates of exposure should saturation or buffering of concentrations of radionuclides in any environmental compartment occur. As an example of a simple, multiplicative-chain model of an exposure pathway that is often used in performance assessment, the annual dose from consumption of vegetables that are assumed to be contaminated by root uptake from contaminated soil (the source compartment) can be represented as: H v = C v U v D ing ,

(5.47)

C v = B v C soil , where: Hv Cv Uv Ding

= = = =

annual dose concentration of radionuclides in vegetables annual consumption of vegetables by exposed individual dose coefficient for ingestion of radionuclide

256 / 5. PERFORMANCE ASSESSMENT MODELS Bv = Csoil =

plant-to-soil concentration ratio for radionuclide concentration of radionuclide in soil in root zone

Similar model equations can be developed for other exposure pathways of concern. Multiplicative-chain models of many exposure pathways have been discussed, for example, in reviews by IAEA (1994c; 2001) and NCRP (1984a). Given the long-term, prospective nature of performance assessment, the somewhat hypothetical nature of assumed exposure scenarios at future times, and the primary importance of scenarios that involve chronic exposure, simple multiplicative-chain models of exposure pathways generally are suitable for use in performance assessment. More complex, dynamic models that account for timeand seasonal-dependencies of exposure, such as those developed to evaluate consequences of atmospheric testing of nuclear weapons or accidental releases to the environment (Whicker and Kirchner, 1987) are not needed in performance assessment. 5.10.4.1.3 Specific-activity models. Dose assessments for 3H and C often are handled as special cases using so-called specificactivity models (NCRP, 1984a). A specific-activity model is appropriate when a radionuclide released to the environment is rapidly mixed with its naturally occurring stable-element counterpart. This type of model essentially assumes that the activity of a radionuclide per unit mass of its stable element in body tissues is at equilibrium with the activity per unit mass of its stable element in a source compartment or compartments in an individual’s exposure environment. A specific-activity model is attractive, when it is appropriate, because it obviates the need to consider transfer factors and, in some cases, usage factors for relevant exposure pathways. Use of a specific-activity model in estimating dose is illustrated by the following example for 14C. The atmosphere is the source compartment for nearly all carbon in body tissues, and by far the most important exposure pathways involve incorporation of atmospheric CO2 into plants by photosynthesis and subsequent intakes in terrestrial foods (Killough and Rohwer, 1978). Therefore, for a release of 14C to the atmosphere and assuming that released material is rapidly converted to 14CO2, the dose rate to an exposed individual can be estimated as the product of two factors: (1) the activity of 14C per unit mass of total carbon in the atmosphere at an assumed receptor location, as obtained from an atmospheric transport model for released 14C and the known density of CO2 in the atmosphere; and (2) the dose rate in body tissues per unit specific activity of 14C, as obtained from the 14

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known carbon content of body tissues and decay energy of 14C (Killough and Rohwer, 1978). Use of a specific-activity model to estimate dose in this case does not require knowledge of transfer and usage factors for various exposure pathways that contribute to intakes of 14C. The example of 14C released to the atmosphere described above provides a particularly compelling case for use of a specific-activity model, because the atmosphere is the source of almost all carbon in body tissues. Thus, if all of an individual’s foods from terrestrial sources are assumed to be obtained from the vicinity of a receptor location, the specific activity of carbon in body tissues should be nearly equal to the specific activity of carbon in the atmosphere at that location. If only a fraction of an individual’s foods from terrestrial sources is assumed to be obtained from the receptor location and the remainder from locations that are uncontaminated by a release, the dose obtained from a specific-activity model can be reduced by that fraction. A specific-activity model generally is appropriate for chronic releases of 14C to the atmosphere. However, this type of model should not be used to assess impacts of releases of 14C to groundwater or surface waters, because almost all carbon in body tissues comes from the atmosphere and the fraction of carbon in body tissues that comes from water is not known (Killough and Rohwer, 1978). A dose estimate based on an assumption that the specific activity of carbon in body tissues is the same as the specific activity in water thus would be unreasonably conservative. In performance assessment, conventional exposure pathway models of the kind described earlier in this Section should be used in assessing impacts of releases of 14C to water. A specific-activity model generally can be used to estimate dose from chronic releases of 3H to air or water on the basis of an assumption that a release is in the form of tritiated water. However, application of such a model to 3H is less straightforward than in the case of atmospheric releases of 14C (NCRP, 1984a). Although the dose rate per unit specific activity of 3H in body tissues is known (NCRP, 1979), a specific-activity model for releases of 3H to the atmosphere must take into account the atmospheric humidity of water vapor, which varies substantially with geographic location (Etnier, 1980). In addition, for any release pathway, significant intakes of water from several environmental compartments should be taken into account, because it generally would be quite conservative to assume that the specific activity of 3H in body tissues is equal to the specific activity in a compartment to which releases occur (e.g., the atmosphere, surface waters, or groundwater). In

258 / 5. PERFORMANCE ASSESSMENT MODELS general, it is not significantly more difficult to use conventional exposure pathway models for 3H.

5.10.4.2 Models of Foodchain Pathways. Models of terrestrial and aquatic foodchain pathways often take into account consumption of different types of foods obtained from that environment. However, given the prospective and somewhat hypothetical nature of assumed exposure scenarios, simplifying assumptions about the number of types of foods that would be consumed normally should be appropriate in performance assessment. Nonetheless, exceptions to recommendations discussed in the following two sections could be appropriate if residents near a disposal site have unusual dietary habits.

5.10.4.2.1 Terrestrial foodchain pathways. Ingestion of contaminated vegetables, milk and meat are exposure pathways that are often evaluated in performance assessment. Models of a vegetable pathway may include separate transfer and usage factors for such subcategories as leafy vegetables, nonleafy vegetables, fruits, and grains. Similarly, models of a milk pathway may consider goat’s milk as well as cow’s milk, and models of a meat pathway may consider pork, chicken and eggs as well as beef. In performance assessment, however, it usually is not necessary to distinguish between different types of vegetables, milk and meat, primarily because much of the available data on transfer and usage factors for these pathways will be generic rather than site-specific. It normally should be sufficient to develop a single set of transfer and usage factors for each of the vegetable, milk, and meat pathways.

5.10.4.2.2 Aquatic foodchain pathways. Aquatic foodchain pathways need to be considered in performance assessment only if releases of radionuclides to surface waters are potentially important. Thus, these pathways are unimportant at many sites. Some models of aquatic foodchain pathways include separate transfer and usage factors for fish, molluscs and crustaceans. In performance assessment, however, it normally should be sufficient to develop a single set of transfer factors and a single usage factor for a fish pathway. Those factors should be based on data for the type of aquatic food that is expected to be most important near a disposal site.

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5.10.5 Selection of Model Parameter Values Once exposure pathways are defined and model equations for the pathways are formulated, values of transfer factors, usage factors, and dose coefficients contained in the models must be selected. The following general recommendations apply to selection of those parameter values for use in performance assessment. 1. Standard dose coefficients for reference adults developed by ICRP or EPA should be appropriate for use in estimating dose from ingestion, inhalation and external exposure pathways. 2. Site-specific data on transfer and usage factors should be used whenever they are available and relevant to low-level waste disposal at a site of concern. If site-specific data are not available, as is often the case with transfer factors, generic values obtained from the literature must be used. However, the applicability of selected sets of generic data to a site-specific analysis should be considered and justified, because some data sets may not be representative of expected site conditions. 3. In selecting transfer and usage factors from site-specific or generic data, values that are believed to represent conditions that are reasonably likely to occur (e.g., central estimates) should be emphasized, rather than values that are believed to be conservative in most exposure situations, primarily because definitions of exposure scenarios are likely to be conservative. For example, it is generally assumed in performance assessment that individuals are exposed at locations of maximum projected concentrations of radionuclides in the environment. In addition, a conservative bias normally is used in modeling releases of radionuclides from a disposal facility and transport in the environment. Thus, even when transfer and usage factors in exposure pathway models are central estimates (e.g., mean or median values), it is expected that calculated doses to individuals will be conservatively biased. 4. The previous recommendation notwithstanding, intentionally conservative transfer and usage factors are appropriate for use in screening models. Conservative assumptions also may be appropriate in demonstrating compliance with applicable performance objectives, particularly if more realistic assumptions would be difficult or expensive to defend. As discussed more generally in Section 2.4, a choice between

260 / 5. PERFORMANCE ASSESSMENT MODELS conservative and more realistic assumptions is largely a matter of judgment. That choice could depend, for example, on whether use of conservative assumptions results in significant quantities of waste that would be unacceptable for disposal. In some cases, use of conservative parameter values may be mandated by regulatory authorities. An example is a common assumption that an exposed individual consumes 2 L d–1 of drinking water from a contaminated source (EPA, 1976; 1993a; 2000b; 2001a). 5. Generic data for some transfer factors, particularly those used to describe transfer of radionuclides through terrestrial foodchain pathways, may be obtained from several compilations. In selecting transfer factors for a large number of radionuclides of concern to low-level waste disposal, a minimal number of data sources should be used, and mixing of data from a large number of sources should be avoided unless it can be justified. Although use of generic data from any source involves some uncertainty, the recommended approach has the advantage that selected data at least should be internally consistent and reflect the biases of relatively few data analysts. Above all, selection of parameter values for use in exposure pathway models should be the responsibility of an analyst. Deviations from these general recommendations could be reasonable at specific sites, provided they are justified. 5.10.6 Sources of Generic Data on Model Parameter Values This Section discusses sources of generic data on dose coefficients, usage factors, and transfer factors that should be appropriate for use in most performance assessments. In the absence of site-specific data, especially data on usage factors and transfer factors, use of generic data from sources discussed in this Section should require a minimum of justification. However, generic data on usage factors and transfer factors could be inappropriate if population groups near a disposal site have unusual living and dietary habits. 5.10.6.1 Dose Coefficients. In principle, dose coefficients for internal and external exposure could be site-specific to reflect variations in ethnic makeup of potentially exposed populations at different sites. In practice, however, the only available data are generic dose

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coefficients for reference individuals, which are intended to represent all ethnic groups. 5.10.6.1.1 Internal exposure. Dose coefficients for ingestion and inhalation of radionuclides by members of the public are available from a number of sources, and those dose coefficients were calculated on the basis of a variety of models and assumptions. Therefore, selection of dose coefficients for internal exposure for use in a performance assessment should be based on an understanding of approaches used in calculating particular data sets. Acceptance of particular data sets by regulatory authorities also may be an important consideration. Dose coefficients developed by federal agencies and ICRP are described in Table 5.8 and discussed below. Federal Guidance on Dose Coefficients for Reference Adults. Dose coefficients for ingestion and inhalation of radionuclides that were developed for use by federal agencies are contained in EPA’s Federal Guidance Report No. 11 (Eckerman et al., 1988). These dose coefficients apply to reference adults, and they give 50 y committed doses per unit activity intake in the form of dose equivalents to specific organs or tissues, now called equivalent doses (ICRP, 1991; NCRP, 1993), and effective dose equivalents. Effective dose equivalent is a weighted sum of dose equivalents to specified organs or tissues that was developed in ICRP Publication 26 (ICRP, 1977). The weighting factor for each organ or tissue, as given in Table 5.9, is proportional to an assumed risk of fatal cancers or severe genetic effects per unit dose equivalent in that organ or tissue. Dose coefficients in Federal Guidance Report No. 11 were calculated using a consistent set of dosimetric and biokinetic models for all radionuclides, and dose coefficients in the form of committed effective dose equivalents are consistent with performance objectives for low-level waste disposal (Section 3.4.2). For some radionuclides, ingestion dose coefficients are given for two values of the gastrointestinal (GI)-tract absorption fraction, denoted by f1 (Eckerman et al., 1988). In such cases, the lower value applies to relatively insoluble chemical forms, and the higher value to more soluble compounds. For exposure pathways that involve ingestion of radionuclides in water or incorporated into foods, the dose coefficient corresponding to the higher f1 normally should be used, because radionuclides should be in a relatively soluble form. This assumption usually gives a higher dose. The same assumption can be used for such other pathways as direct ingestion of radionuclides in soil or ingestion of radionuclides deposited on surfaces of vegetation. However, if the chemical form of a radionuclide in these pathways is expected to be relatively insoluble, the

Source

Description of Dose Coefficients

EPA’s Federal Guidance Report No. 11 (Eckerman et al., 1988)

Organ dose equivalents and effective dose equivalents for ingestion and inhalation by reference adults based on dosimetric and biokinetic models in ICRP Publications 30 and 48 (ICRP, 1979a; 1979b; 1980; 1981a; 1981b; 1982a; 1982b; 1986).

DOE (1988c)b

Same as EPA’s Federal Guidance Report No. 11.

ICRP Publication 56 (ICRP, 1989)

For selected radionuclides, organ dose equivalents and effective dose equivalents for ingestion and inhalation by reference individuals of different ages; calculations for some radionuclides use newer physiologically-based biokinetic models.

ICRP Publications 67, 69 and 71 (ICRP, 1993b; 1995; 1996a)

For selected radionuclides, including those considered in ICRP Publication 56 (ICRP, 1989), organ equivalent doses and effective doses for ingestion or inhalation by reference individuals of different ages; calculations for inhalation use new lung model (ICRP, 1994a), and calculations for some radionuclides use newer physiologically-based biokinetic models.

ICRP Publication 72 (ICRP, 1996b)

Effective doses for ingestion and inhalation by reference individuals of different ages; calculations for inhalation use new lung model (ICRP, 1994a), and calculations for some radionuclides use newer physiologically-based biokinetic models.

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TABLE 5.8—Selected sources of dose coefficients for ingestion and inhalation of radionuclides by members of the public.a

For 222Rn and 220Rn and for indoor and outdoor exposure, effective dose equivalents to reference adults per unit activity concentration of radon in air per unit time and effective dose equivalents per unit exposure to emitted alpha-particle energy in air from short-lived radon decay products.

ICRP Publication 65 (ICRP, 1994b)

For 222Rn and for indoor exposure, effective doses to reference adults per unit activity concentration of radon in air and per unit exposure to emitted alpha-particle energy in air from short-lived radon decay products.

EPA’s Federal Guidance Report No. 13 (Eckerman et al., 1999; EPA, 2000d)

Organ equivalent doses and effective doses to members of the general public of different ages based on lung model and dosimetric and biokinetic models currently recommended by ICRP.c

aData

sets are discussed in Section 5.10.6.1.1. coefficients differ in some cases from those in Federal Guidance Report No. 11 (Eckerman et al., 1988), due to roundoff errors in generating values. cIn addition to committed doses, EPA (2000d) gives doses received versus time after intake. bDose

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264 / 5. PERFORMANCE ASSESSMENT MODELS TABLE 5.9—Tissue weighting factors (wT) used in calculating effective dose equivalent and effective dose.a Effective Dose Equivalent (HE)

Effective Dose (E) Tissue or Organ

w Tc

0.25

Gonads

0.20

Breast

0.15

Red marrow

0.12

Red marrow

0.12

Colon

0.12

Lung

0.12

Lung

0.12

Thyroid

0.03

Stomach

0.12

Bone surface

0.03

Bladder

0.05

Remainderd

0.30

Breast

0.05

Liver

0.05

Esophagus

0.05

Thyroid

0.05

Skin

0.01

Bone surface

0.01

Remaindere

0.05

Tissue or Organ

wT

Gonads

b

aEffective

dose equivalent and effective dose are weighted sums of equivalent doses to specified organs or tissues defined in ICRP Publications 26 and 60 (ICRP, 1977; 1991), respectively. bWeighting factor for gonads represents severe genetic effects in first two generations; all other weighting factors represent risk of fatal cancers. cWeighting factor for gonads represents severe genetic effects in all future generations; all other weighting factors represent total detriment taking into account risk of fatal cancers, weighted nonfatal cancer incidence, and years of life lost due to fatal health effects in organ or tissue relative to average years of life lost due to all fatal cancers. d Remainder tissues include five other organs, excluding skin, lens of the eye, and body extremities, receiving the highest doses, and each organ is assigned weighting factor of 0.06. e See ICRP Publication 60 (ICRP, 1991) for description of procedure for calculating contribution from remainder tissues.

dose coefficient corresponding to the lower f1 can be used. Thus, for some radionuclides, different ingestion dose coefficients reasonably could be used for different exposure pathways in the same assessment. In practice, however, information on chemical forms of radionuclides in low-level waste often is not available, in which case it would be reasonable to use dose coefficients corresponding to the higher values of f1 for all ingestion pathways.

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For some radionuclides, dose coefficients for inhalation are given for two or three lung clearance (i.e., solubility) classes (Eckerman et al., 1988). These are denoted by Class D for relatively soluble chemical forms that are cleared from the pulmonary region in a matter of days, Class W for moderately soluble forms that are cleared in a matter of weeks, and Class Y for relatively insoluble forms that are cleared in a matter of years. If information on chemical forms of inhaled radionuclides is not available, clearance classes generally should be chosen to maximize dose coefficients for the most important radioisotopes of particular elements. It usually is not appropriate to assume one clearance class for one isotope and another clearance class for another isotope of the same element in order to maximize the dose for each isotope, because all isotopes of a given element presumably have the same solubility unless there is evidence to the contrary based on knowledge of differing chemical forms in waste. When this situation occurs (e.g., in inhalation of Class-W and -Y compounds of certain isotopes of thorium), differences in dose coefficients for different clearance classes usually are not significant. Dose coefficients for ingestion or inhalation by reference adults also have been compiled by DOE (1988c). These data are based on the same dosimetric and biokinetic models as data in Federal Guidance Report No. 11 (Eckerman et al., 1988). However, there are differences in dose coefficients for some radionuclides, due primarily to roundoff errors in generating results in the DOE compilation. These differences generally are not significant. ICRP has developed dose coefficients for ingestion and inhalation of radionuclides that supersede those given in EPA’s Federal Guidance Report No. 11 (Eckerman et al., 1988); see Table 5.8. These dose coefficients are described in the following paragraphs. Use of dose coefficients currently recommended by ICRP has not been endorsed by federal agencies, although NRC has approved their use on a case-by-case basis (Vietti-Cook, 1999). In addition, as discussed below and in Section 5.10.8, dosimetric and biokinetic models used to obtain ICRP’s current database of dose coefficients for members of the public of different ages are incorporated in risk coefficients given in EPA’s Federal Guidance Report No. 13 (Eckerman et al., 1999) and in a database of dose coefficients that accompanies that report (EPA, 2000d). ICRP Recommendations on Effective Dose Coefficients for Reference Adults in General Population. In accordance with a recommendation in ICRP Publication 60 (ICRP, 1991), which has been endorsed by NCRP (1993), ICRP now calculates dose coefficients for ingestion and inhalation as committed effective doses per unit

266 / 5. PERFORMANCE ASSESSMENT MODELS activity intake, rather than effective dose equivalents. The effective dose also is a weighted sum of equivalent doses to different organs or tissues, but the number of organs or tissues considered explicitly is greater, and the weighting factor for each organ or tissue given in Table 5.9 includes relatively small contributions representing incidence of nonfatal cancers and years of life lost from fatal health effects in that organ or tissue relative to the average years of life lost from all fatal cancers, in addition to the risk of fatal cancers or severe genetic effects (ICRP, 1991). The committed effective dose per unit activity intake of a radionuclide usually differs somewhat from the committed effective dose equivalent. Such differences usually are due primarily to differences in biokinetic models for inhaled and ingested radionuclides. For reference adults in the general population, effective dose coefficients for ingestion and inhalation recommended by ICRP are compiled in Publication 72 (ICRP, 1996b). These dose coefficients are the same as, or differ very little from, recommended values for reference adult workers compiled in ICRP Publication 68 (ICRP, 1994c). Dose coefficients for specific organs or tissues used in calculating the effective dose coefficients for reference adults are compiled in ICRP Publications 56, 67, 69 and 71 (ICRP, 1989; 1993b; 1995; 1996a). Dose coefficients for reference adults in the general population in ICRP Publication 72 (ICRP, 1996b) differ from those in EPA’s Federal Guidance Report No. 11 (Eckerman et al., 1988) in three respects, in addition to the use of effective dose instead of effective dose equivalent. First, inhalation dose coefficients are based on a new model to describe the dynamic behavior of radionuclides in the respiratory tract (ICRP, 1994a). Second, calculations for many (but not all) radionuclides use newer physiologically-based biokinetic models (ICRP, 1989; 1993b; 1995; 1996a), rather than the previous empirical approach of modeling retention of radionuclides in the body by fitting retention or excretion data over time (ICRP, 1979a). As a result, in contrast to Federal Guidance Report No. 11, dose coefficients for reference adults in ICRP Publication 72 are not based on the same approach to biokinetic modeling for all radionuclides. Third, for some radionuclides, the value of the GI-tract absorption fraction (f1) has been revised. Age-Specific Dose Coefficients. The general population consists of people of all ages, not just adults. As described below, ICRP has developed dose coefficients for reference individuals of younger ages. However, age-specific dose coefficients should be used with caution, and only with knowledge of assumptions used in generating the data and the purpose of an analysis.

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Some age-specific dose coefficients for ingestion or inhalation, such as those developed by NRC (1977b), are based on outdated dosimetric and biokinetic models in ICRP Publication 2 (ICRP, 1959). Furthermore, dosimetric and biokinetic models for reference adults often were used in calculations for all age groups, and only the masses of particular organs were assumed to depend on age. Thus, age-specific dose coefficients other than those developed by ICRP and cited in Table 5.8 are considered inappropriate for use in any type of dose assessment. ICRP’s earliest age-specific dose coefficients for ingestion or inhalation of radionuclides (ICRP, 1989) have been superseded by data in the later publications, especially in regard to use of the effective dose (ICRP, 1991) and the new lung model (ICRP, 1994a). All dose coefficients account for the age dependence of organ masses and locations in the body. The age dependence of absorption, deposition and retention in the body is taken into account for many (but not all) radionuclides, and many calculations also are based on the newer physiologically-based biokinetic models. As a result, as in the case of the effective dose coefficients for reference adults discussed previously, age-specific dose coefficients currently recommended by ICRP are not based on the same approach to biokinetic modeling for all radionuclides. Although ICRP has developed age-specific dose coefficients for all radionuclides of concern to low-level waste disposal, those data are not considered appropriate for use in performance assessments for the purpose of demonstrating compliance with performance objectives expressed in terms of annual dose. Use of dose coefficients for reference adults, but not other age groups, for that purpose can be justified on the following technical grounds. First, scenarios for long-term, chronic exposure should be of primary importance in demonstrating compliance with performance objectives, and most such exposures would occur during adult years. Second, risks due to chronic lifetime exposure, taking into account the age dependence of dose and risk, should not differ greatly from risks due to exposure during adult years only (Kocher et al., 1988). Finally, for chronic lifetime exposure to different radionuclides, a limit on annual dose applied to any age at exposure corresponds much more poorly to a particular level of cancer risk than a limit on annual dose applied to adults only (Kocher and Eckerman, 1998; Kocher et al., 1988). More generally, the use of dose coefficients for adults can be justified on the grounds that, as emphasized in this Report, the purpose of performance assessment is to inform judgments about acceptable disposal practices, rather than to predict doses to any age group.

268 / 5. PERFORMANCE ASSESSMENT MODELS Dose Coefficients for Radon. Dose coefficients for inhalation of radon and its short-lived decay products in air normally are treated as special cases, in part because the dose per unit concentration of radon in air depends on environmental factors, including the extent of activity equilibrium between radon and its decay products and the fraction of the activity of decay products that is attached to respirable aerosols. Sources of dose coefficients for radon recommended by ICRP for use in radiation protection are summarized in Table 5.8. The dosimetry of inhaled radon and its decay products is discussed in reports by ICRP (1987a) and NCRP (1984b). For indoor or outdoor exposure to 222Rn or 220Rn, dose coefficients developed by ICRP (1987a) for assumed reference conditions that would be suitable for use in performance assessments are summarized in Table 5.10. These dose coefficients are in the form of doses per unit exposure (i.e., doses per unit activity concentration of radon in air per unit exposure time) and they can be multiplied by assumed indoor and outdoor exposure times in hours to give estimates of dose per unit concentration of radon in air. For indoor exposure to 222Rn, but not 220Rn, revised dose coefficients in the form of effective doses per unit activity concentration of radon in air and effective doses per unit exposure to the emitted alpha-particle energy in air produced by short-lived decay products of radon are given in ICRP Publication 65 (ICRP, 1994b). Those data should be used with caution in demonstrating compliance with performance objectives in the form of organ dose equivalents or effective dose equivalents, unless their use has been endorsed by regulatory authorities. Some exposure scenarios, especially scenarios for exposure of inadvertent intruders (Section 6.3), involve residence on contaminated surface soil. If contaminants include 226Ra or 232Th, doses from 222Rn or 220Rn and their short-lived decay products per unit concentration of the parent radionuclide in soil can be estimated using data on average annual doses to the public from exposure to naturally occurring radon (NCRP, 1987; UNSCEAR, 2000) and average background concentrations of the parent radionuclides in soil (NCRP, 1984c). Such a natural analog approach bypasses the need to model releases of radon from soil and resulting doses from inhalation of radon and its decay products (McDowell-Boyer et al., 2000; ORNL, 1997a). 5.10.6.1.2 External exposure. Compared with ingestion and inhalation, fewer options are available in selecting dose coefficients for external exposure to radionuclides in the environment. Therefore,

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TABLE 5.10—Dose coefficients for indoor and outdoor exposure of reference adults to radon and its decay products recommended by ICRP.a Dose per Unit Exposure (mSv per Bq-h m–3) Isotope/Location

Equivalent Dose Effective Dose Equivalentb

Bronchial Epithelium

Pulmonary Tissue

Indoors

1.5 × 10–4

2.0 × 10–5

1.0 × 10–5

Outdoors

2.0 × 10–4

2.7 × 10–5

1.4 × 10–5

Indoors

5.0 × 10–4

1.5 × 10–4

5.0 × 10–5

Outdoors

5.0 × 10–4

1.5 × 10–4

5.0 × 10–5

222

Rn

220Rn

aValues obtained from Section 3.3 and Table 4 of ICRP Publication 50 (ICRP, 1987a). bIn estimating effective dose equivalents, weighting factor of 0.06 is applied to mean doses to bronchial epithelium and pulmonary tissues, and effective dose equivalents for 220Rn include contributions from doses to bone surfaces, bone marrow, kidneys, and liver.

fewer judgments are required in selecting data sets. Compilations of external dose coefficients for a large number of radionuclides, including those of importance in low-level waste, are summarized in Table 5.11 and discussed below. Federal Guidance on Dose Coefficients for Reference Adults. Dose coefficients for external exposure to radionuclides in the environment currently recommended for use by federal agencies in assessing dose to the public are contained in EPA’s Federal Guidance Report No. 12 (Eckerman and Ryman, 1993). These dose coefficients apply to reference adults and give dose rates to specific organs and effective dose-equivalent rates per unit concentration of radionuclides for specified exposure situations. Dose coefficients for exposure to a contaminated ground surface are not expected to be relevant in most exposure scenarios for low-level waste disposal. Exceptions could occur if exposure pathways involving deposition of activity onto undisturbed ground are considered. All calculations assume that the source region is uniformly contaminated and either infinite or semi-infinite in extent.

270 / 5. PERFORMANCE ASSESSMENT MODELS TABLE 5.11—Selected sources of dose coefficients for external exposure to radionuclides in the environment.a Source

Description of Dose Coefficients

EPA’s Federal Guidance Report No. 12 (Eckerman and Ryman, 1993)

Organ equivalent-dose rates and effective dose-equivalent rates for reference adults for immersion in contaminated air, immersion in contaminated water, exposure to contaminated ground surface, and exposure to surface soil contaminated to depths of 1, 5 and 15 cm and to infinite depth.

EPA’s Federal Guidance Report No. 11 (Eckerman et al., 1988)b

For elemental 3H and selected noble-gas radionuclides, organ equivalent-dose rates and effective dose-equivalent rates for reference adults for immersion in contaminated air.

DOE guidance (DOE, 1988d)b

Organ equivalent-dose rates and effective dose-equivalent rates for reference adults for immersion in contaminated air, immersion in contaminated water, and exposure to contaminated ground surface.

aData

sets are discussed in Section 5.10.6.1.2. For additional references to external dose calculations, see NCRP (1999). b Dose coefficients differ by few tens of percent or less from those in Federal Guidance Report No. 12 (Eckerman and Ryman, 1993), which corrects small errors in previous calculations.

Dose coefficients for sources on the ground surface or distributed in a depth of surface soil in Federal Guidance Report No. 12 (Eckerman and Ryman, 1993) take into account the proper energy and angular dependence of the radiation field above ground for these exposure situations. These dose coefficients supersede values in a previous compilation (DOE, 1988d) and values that would be obtained from calculations for monoenergetic photon emitters in soil by Kocher and Sjoreen (1985). In addition, small errors of a few tens of percent or less in previous dose coefficients for immersion in contaminated air or water (DOE, 1988d; Eckerman et al., 1988) are corrected. For sources distributed in a volume of surface soil, dose coefficients in Federal Guidance Report No. 12 (Eckerman and Ryman, 1993) can be used with little error in estimating external dose from finite source regions with lateral dimensions of a few meters or more. For smaller source regions, those dose coefficients would provide conservative estimates of dose. If a more realistic estimate of dose from a small source region is desired, use of a numerical method implemented by a computer code is required.

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Effective Dose Coefficients. Radionuclide-specific dose coefficients for external exposure in the form of effective doses (ICRP, 1991) have not been published. However, for photon energies above ~0.1 MeV, which are of greatest importance in estimating external dose, effective doses from external exposure do not differ significantly from effective dose equivalents, because the whole body is irradiated nearly uniformly (ICRP, 1987b; 1997c). Age-Specific Dose Coefficients. Radionuclide-specific dose coefficients for external exposure described above apply to reference adults. Values for reference individuals of younger ages also have been published (NCRP, 1999). For higher-energy photons of greatest importance in estimating external dose, dose coefficients for younger age groups do not differ significantly from those for adults, because the radiations are not greatly attenuated in traversing the human body and the whole body is irradiated nearly uniformly (ICRP, 1997c). As with dose coefficients for inhalation and ingestion of radionuclides, dose coefficients for external exposure of reference adults generally should be used in performance assessment for the purpose of demonstrating compliance with performance objectives. Correction Factors to External Dose Coefficients. External dose coefficients for immersion in contaminated air and exposure to a contaminated ground surface or volume of surface soil discussed above assume that an exposed individual is standing on the ground outdoors. In many scenarios, however, external exposure is assumed to occur mainly during indoor residence. Shielding factors for different types of buildings have been summarized by NCRP (1984a; 1999), and values recommended for use in generic dose assessments are given by NRC (1977b). Taking into account the fraction of the time spent indoors, shielding during indoor residence should reduce estimated external doses by at least 30 to 40 %. For exposure to a contaminated ground surface, which usually should not be important in performance assessment, ground roughness and terrain irregularities could reduce external dose to an individual standing on the ground. Empirical corrections to account for these effects are discussed by Burson and Profio (1977). However, a reasonable approach to accounting for ground roughness and terrain irregularities is to assume that activity on the ground surface is distributed uniformly to a depth of 1 cm and then calculate external dose using dose coefficients for this volume source (Eckerman and Ryman, 1993).

272 / 5. PERFORMANCE ASSESSMENT MODELS 5.10.6.1.3 Summary of dose coefficients. Several options are available in selecting dose coefficients for ingestion and inhalation of radionuclides and external exposure to radionuclides in the environment for use in performance assessment. This is particularly the case in regard to internal exposure, where dose coefficients have been calculated in terms of different dosimetric quantities (i.e., effective dose equivalent or effective dose), for different age groups, and using different dosimetric and biokinetic models. Dose coefficients for reference adults generally should be used in demonstrating compliance with performance objectives for low-level waste disposal. In general, the most appropriate data sets are effective dose coefficients for ingestion and inhalation given in ICRP Publication 72 (ICRP, 1996b) and dose coefficients for external exposure in the form of effective dose equivalents given in Federal Guidance Report No. 12 (Eckerman and Ryman, 1993). For external exposure, the effective dose should not differ substantially from the effective dose equivalent used in federal guidance (ICRP, 1987b; 1997c). For some scenarios involving external exposure, correction factors to account for building shielding or ground roughness and terrain irregularities could be applied. Given that DOE’s performance objectives for low-level waste disposal are in the form of effective dose equivalents and NRC allows use of that dosimetric quantity in lieu of doses to the whole body or any organ, as specified in 10 CFR Part 61 (NRC, 1982b) (Section 3.4.2), regulatory authorities may require that outdated dose coefficients for ingestion and inhalation in Federal Guidance Report No. 11 (Eckerman et al., 1988) be used in demonstrating compliance. However, permission to use current dose coefficients may be granted by NRC on a case-by-case basis (Vietti-Cook, 1999). Differences between effective dose equivalents and effective doses for adults should have little impact on decisions about acceptable waste disposals at any site. 5.10.6.2 Usage Factors. Generic data on usage factors, including intake rates of foodstuffs, water, air and soil by humans and livestock and residence times for external exposure pathways, often are appropriate for use in performance assessment. However, available site-specific data should be used whenever they are applicable to assumed exposure pathways. An example is an evaluation of usage factors for terrestrial and aquatic foodchain pathways and external exposure pathways for the population near the Savannah River Site (Hamby, 1992). Site-specific data are especially important if nearby residents have unusual dietary and other living habits, as may occur, for example, in populations of Native Americans.

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Sources of generic data on usage factors that should be appropriate for use in most performance assessments are listed in Table 5.12. Particular usage factors are discussed in the following paragraphs. To be consistent with the recommended use of ingestion dose coefficients for reference adults (Section 5.10.6.1.1), use of consumption rates of foodstuffs and water by adults is appropriate in demonstrating compliance with performance objectives. As discussed in Section 5.10.5, consumption rates believed to represent reasonably likely conditions, rather than conservative values, normally should be used in performance assessment. In selecting consumption rates of terrestrial foodstuffs, attention must be given to whether the data are reported on a dry- or fresh-weight basis. Consumption rates should be compatible with TABLE 5.12—Selected sources of generic usage factors for use in exposure pathway models. Usage Factors

Source of Generic Data

Consumption of foodstuffs and water by humansa

IAEA (1994c; 2001) Kennedy and Strenge (1992) NRC (1977b) Rupp (1980) Yang and Nelson (1986) Wood et al. (1997a; 1997b)

Consumption of soil by humansa

NCRP (1999) Wood et al. (1997a)

Consumption of water or pasture grass by livestock

IAEA (1994c; 2001) Kennedy and Strenge (1992) NRC (1977b) Shor and Fields (1980)

Breathing rates by humansa

ICRP (1975) NRC (1977b) Wood et al. (1997a)

Exposure times for external exposure pathways;b fraction of time spent indoors

NCRP (1984a) NRC (1977b; 1981a) Oztunali and Roles (1986) Wood et al. (1997c)

a Usage factors for adults should be emphasized in demonstrating compliance with performance objectives for low-level waste disposal. b Values in assumed scenario often can be based on judgment of analyst.

274 / 5. PERFORMANCE ASSESSMENT MODELS transfer factors for terrestrial foodchain pathways discussed in the following section. Attention also must be given to whether data on consumption rates of pasture grass by livestock are reported on a dry- or fresh-weight basis. Data on consumption rates of soil were reviewed by NCRP (1999); see also Wood et al. (1997a). An average consumption rate of soil by adults appears to be ~50 mg d–1; consumption by children may be considerably higher (Wood et al., 1997a). The soil-ingestion pathway may be particularly important for radionuclides with very low bioaccumulation in plants and animals (e.g., plutonium). Breathing rates by humans depend on the level of activity (ICRP, 1975; NRC, 1977b; Wood et al., 1997a). A standard breathing rate by an average adult that should be suitable for use in most performance assessments is 0.9 m3 h–1, or ~8,000 m3 y–1. 5.10.6.3 Transfer Factors for Foodchain Pathways. This Section considers sources of generic data on transfer factors for radionuclides in terrestrial and aquatic foodchain pathways that should be suitable for use in performance assessment. Factors describing transfer of radionuclides from contaminated surface soil to the atmosphere are discussed in Section 5.8. Site-specific data on transfer factors in foodchain pathways should be used whenever they are available and relevant to a performance assessment. If generic data are used, they should be based on actual measurements rather than indirect evidence, such as an assumption that a transfer factor for a particular element can be based on data for a chemical homolog. Sources of generic data on transfer factors and related parameters for terrestrial and aquatic foodchain pathways that should be appropriate for use in most performance assessments are listed in Table 5.13. These and other sources of data are discussed in the following sections. 5.10.6.3.1 Terrestrial foodchain pathways. Terrestrial foodchain pathways often considered in performance assessment include consumption of vegetables grown in contaminated garden soil and consumption of milk and meat obtained from dairy and beef cattle that consume contaminated water or pasture grass (Table 5.6). Generic Transfer Factors. For the pathway involving consumption of vegetables grown in contaminated soil, the most important model parameter is the plant-to-soil concentration ratio, usually denoted by Bv. Similarly for the milk and meat pathways, the most important parameters are the intake-to-milk and intake-to-meat transfer coefficients in dairy and beef cattle, usually denoted by Fm

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and Ff, respectively. Of the various sources of generic data for these transfer factors listed in Table 5.13, compilations by Baes et al. (1984) and NCRP (1996a; 1999) are particularly useful in prospective dose assessments involving a large number of radionuclides, because default values are provided for nearly all chemical elements. Other sources of generic data on plant-to-soil concentration ratios can be useful in performance assessment. The compilation TABLE 5.13—Selected sources of generic transfer factors and related parameters for terrestrial and aquatic foodchain pathways. Exposure Pathway

Terrestrial foodchains

Aquatic foodchains

Parameter

Source of Generic Data

Plant-to-soil concentration ratios in vegetables (Bv); Intake-to-milk or -meat transfer coefficients in dairy and beef cattle (Fm, Ff)

Baes et al. (1984) IAEA (1994c; 2001) NCRP (1996a; 1999) Ng (1982) Ng et al. (1977; 1982a; 1982b) Peterson (1983) Sheppard et al. (1991)

Fresh-weight to dry-weight ratio of vegetation

Baes et al. (1984) Peterson (1983)

Retention of radionuclides in soil root zone

Baes and Sharp (1983)

Distribution coefficients (Kd) of radionuclides in soil

Baes et al. (1984) EPA (1999b; 1999c) IAEA (1994c) Sheppard and Thibault (1990)

Parameters describing transfer of radionuclides from air to vegetation

IAEA (1994c; 2001) NCRP (1984a)

Fish flesh-to-water concentration factor

Kennedy and Strenge (1992) NRC (1977b) Peterson (1983) Poston and Klopfer (1988) Strenge et al. (1986)

276 / 5. PERFORMANCE ASSESSMENT MODELS by Peterson (1983) particularly addresses the difference between root uptake and foliar deposition as the source of activity in vegetation for chemical elements with very low root uptake (e.g., plutonium). Ongoing compilations by the International Union of Radioecologists (IUR, 1989) provide an extensive database for a limited number of radionuclides. The study by Sheppard et al. (1991) presents the first direct measurements of root uptake of particular chemical forms of 14C in different types of soil. Transfer factors for the vegetable, milk and meat pathways also are given by NRC (1977b). However, these data should be used with caution. Some transfer factors were derived inappropriately (e.g., plant-to-soil concentration ratios were based on separate data sets for stable element concentrations in plants and soils obtained from different locations), and other transfer factors have been superseded by later compilations, such as those in Table 5.13. Selection of Plant-to-Soil Concentration Ratios. In selecting plant-to-soil concentration ratios, as well as consumption rates of contaminated vegetables by humans and consumption rates of pasture grass by dairy and beef cattle, attention must be given to whether the data are reported on a dry- or fresh-weight basis. Either basis is acceptable, but concentration ratios and consumption rates used in a dose assessment must be consistent in this regard. Factors to convert fresh weight of vegetation to dry weight are discussed by Baes et al. (1984) and Peterson (1983). This factor is not the same for leafy and nonleafy vegetables. For example, Baes et al. recommend generic fresh- to dry-weight conversion factors of 0.07 for leafy vegetables and 0.43 for nonleafy vegetables. Thus, the value of 0.25 recommended by NRC (1977b) appears to be only a rough approximation for all types of vegetation. In selecting plant-to-soil concentration ratios, attention also must be given to whether the data apply to leafy or nonleafy vegetables. In performance assessment, it is reasonable to select a single set of plant-to-soil concentration ratios that would apply to all vegetables consumed by humans (Section 5.10.4.2.1). Information that can be taken into account in selecting a single data set for all vegetables includes differences in intakes of leafy and nonleafy vegetables by humans, and differences in plant-to-soil concentration ratios for the two types of vegetation. Similar considerations apply in selecting plant-to-soil concentration ratios for pasture consumed by livestock. Retention of Radionuclides in Surface Soil. If an exposure pathway involving irrigation of food crops or pasture grass with contaminated water is included in a performance assessment, a model of

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long-term retention of deposited radionuclides in the soil root zone should be included. Without such a model, buildup of radionuclides in surface soil and, thus, doses from consumption of contaminated foodstuffs could be greatly overestimated. If irrigation over many years at a constant rate is assumed, a retention model provides estimates of equilibrium concentrations of radionuclides in surface soil relative to their concentrations in irrigation water. The soil leaching model of Baes and Sharp (1983) can be used to estimate this concentration ratio. An important parameter in the model of radionuclide retention in surface soil is the distribution coefficient (Kd), which is the ratio of the concentration of a radionuclide in the solid phase in soil (Bq kg–1) to the concentration in soil water (Bq L–1) at equilibrium. Higher values of Kd result in longer retention times in surface soil and, thus, higher concentrations of radionuclides in surface soil relative to those in irrigation water. Measured distribution coefficients of many chemical elements in soil are highly variable. Therefore, site-specific data should be used whenever they are available and relevant. In the absence of appropriate site-specific data, generic data obtained from sources such as those listed in Table 5.13 must be used. The compilation by Sheppard and Thibault (1990) is particularly useful in selecting data for use at specific sites, because default data for different soil types (sand, loam, clay and organic) are provided. Compilations by Baes et al. (1984), EPA (1999b; 1999c), and Sheppard and Thibault (1990) provide information on probability distributions of Kds for many chemical elements that would be suitable for use in parameter uncertainty analyses. Since the distribution coefficient that might be used in a model of leaching in surface soil is likely to be highly uncertain for many radionuclides, generic values used in performance assessment should be selected from a minimal number of sources, so that the assumptions for different radionuclides will at least be internally consistent. In applying a soil leaching model in a dose assessment for terrestrial foodchain pathways, it is important to take into account that the distribution coefficient (Kd) usually is negatively correlated with the plant-to-soil concentration ratio (Bv). That is, high values of Kd usually are associated with low values of Bv, and vice versa, because radionuclides are readily available for root uptake only if they are relatively soluble in soil water (Baes et al., 1984; Sheppard and Thibault, 1990). Therefore, if selected values of Kd and Bv are both at the upper end of their ranges of measured values, dose estimates for terrestrial foodchain pathways following irrigation with contaminated water could be unreasonably conservative.

278 / 5. PERFORMANCE ASSESSMENT MODELS Additional Parameters in Models of Atmospheric Deposition onto the Ground. If transfer of radionuclides from the atmosphere to the ground surface is important in a performance assessment, additional parameters must be considered in models of terrestrial foodchain pathways. These parameters include: (1) the deposition rate or net flux of radionuclides from air to the ground surface, which usually is described in terms of a deposition velocity; (2) the fraction of deposited material that is intercepted and immediately retained on the surfaces of vegetation, which usually is called the interception fraction; (3) the fraction of activity deposited on plant surfaces that is translocated to edible parts of the plant, which usually is called the translocation factor; (4) the time required for activity deposited on plant surfaces to be lost from the plant, which usually is represented by a weathering half-time; and (5) the standing biomass of vegetation per unit area. Representative values of these parameters that should be suitable for use in performance assessment are discussed by IAEA (1994c; 2001) and NCRP (1984a). Terrestrial foodchain pathways resulting from atmospheric releases of 3H and 14C normally are treated as special cases using specific-activity models (Section 5.10.4.1.3). 5.10.6.3.2 Aquatic foodchain pathways. Models of aquatic foodchain pathways used in performance assessment normally do not need to distinguish between consumption of fish, molluscs and crustaceans. Rather, it usually is satisfactory to use a single set of transfer factors for the fish pathway. Transfer factors should be chosen to represent the most important type of aquatic foodstuff that could be consumed near the disposal site (Section 5.10.4.2.2). The only transfer factor of concern in models of the fish pathway is the ratio of the concentration of a radionuclide in fish to the concentration in water (i.e., the concentration factor). Sources of generic data on concentration factors in freshwater fish are listed in Table 5.13. An extensive data set also is given by Napier et al. (1988), but sources of these concentration factors are not provided and those data should be used with caution whenever they differ from those found in other compilations. In selecting concentration factors in fish for use in dose assessments, attention should be given to whether the values apply to freshwater or saltwater fish. The compilation by Poston and Klopfer (1988), for example, provides data in both types of fish. In addition, attention should be given to whether selected values apply to fish flesh or the whole fish, because the concentration factors of some elements (e.g., strontium) in the skeleton can differ substantially from those in flesh. Selected concentration factors

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should be representative of fish-eating habits of populations for whom doses are calculated. For example, some Native Americans often consume whole fish. 5.10.6.3.3 Summary of transfer factors. Transfer factors and related parameters in terrestrial and aquatic foodchain pathways may be selected from several sources, and widely varying values can be found for some radionuclides. Thus, application of generic data to site-specific assessments may be highly uncertain, and considerable judgment may be required in selecting values for use in a particular assessment. It is partly because of these uncertainties that it is desirable to use a minimal number of different sources in developing data sets for use in performance assessment. Transfer factors believed to represent reasonably likely values are appropriate for use in performance assessment (Section 5.10.5), particularly when values are supported by relevant site-specific data. However, given the large uncertainties in many transfer factors, values that are likely to be conservative could be selected. Whether choosing conservative transfer factors is reasonable depends largely on the consequences. When use of conservative transfer factors is important to decisions about acceptable waste disposals, attention would need to be given to developing and justifying more realistic values. Thus, in selecting transfer factors, as in all other aspects of performance assessment, judgment on the part of the analyst is required. 5.10.7 Uncertainties in Dose Assessment Models Uncertainties in dose assessment models included in a performance assessment depend on uncertainties in transfer factors, usage factors, and dose coefficients. This Section discusses available information on uncertainties in these parameters and their significance; other NCRP reports also include assessments of uncertainty in these parameters (NCRP, 1999; 2004). 5.10.7.1 Uncertainties in Transfer and Usage Factors. Transfer and usage factors included in dose assessment models often are generic values based on data obtained at a variety of locations and under a variety of conditions. This is particularly the case in regard to transfer factors in the vegetable, milk, meat and fish ingestion pathways. In well-studied cases, such as the air-grass-cow-milk pathway for radioiodine and root uptake of radiostrontium from soil, uncertainties in transfer factors appear to be less than an order of

280 / 5. PERFORMANCE ASSESSMENT MODELS magnitude (NCRP, 1984a; 1999; Ng et al., 1982a). In other cases, such as root uptake of actinide elements from soil, uncertainties in transfer factors may be an order of magnitude or more (Ng et al., 1982a). Transfer of radionuclides from surface soil into the atmosphere also appears to be highly uncertain (Section 5.8). The applicability of generic transfer factors to specific sites has not often been investigated. Thus, it usually is not known if ranges or probability distributions of parameters obtained from generic data sets would provide reasonable representations of uncertainty in values at a particular site. Usage factors often are based on generic data sets, although data applicable to particular geographical regions are increasingly becoming available (Hamby, 1992; Yang and Nelson, 1986). Available region-specific data indicate that uncertainties in most usage factors for ingestion and inhalation pathways are less than an order of magnitude. Although generic data on usage factors for external exposure pathways (i.e., residence times) are available (Wood et al., 1997a), values used in performance assessment often must be based largely on judgment of the analyst. In such cases, it may not be meaningful to assign uncertainties to assumed values, especially when they are biased conservatively. 5.10.7.2 Significance of Uncertainties in Transfer and Usage Factors. As indicated in the previous section, there may be considerable uncertainty in using generic data sets of transfer and usage factors in performance assessments at specific sites. However, for reasons described below, these uncertainties may not be significant with respect to an assessment of the overall performance of a disposal system. First, in assessing the overall uncertainty in the performance of disposal systems, uncertainties in calculated release rates of radionuclides to the environment (the source term) and transport of radionuclides in the environment should be significantly greater than uncertainties in transfer and usage factors in an exposure pathway model. There are at least some generic environmental data that can be used to calibrate exposure pathway models (e.g., see references in Table 5.12 and 5.13), but there are few data that can be used to calibrate models of the source term and environmental transport. Uncertainties in estimated inventories of radionuclides in disposed waste also can be an important contributor to the overall uncertainty in a performance assessment (ORNL, 1997a). Second, to the extent that generic transfer and usage factors are used in dose assessments, uncertainties in exposure pathway

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models would be the same at all disposal sites. Therefore, an accounting of these uncertainties may not provide useful information in evaluating whether one disposal site should be better than another. This would not be the case only if the most important radionuclides or exposure pathways in a performance assessment were not the same at different sites. Similar considerations would apply in comparing different disposal technologies at the same site. Finally, some performance assessments have indicated that if releases occur primarily to water, exposure pathways other than direct ingestion of radionuclides in drinking water often should be relatively unimportant, regardless of uncertainties in transfer factors or usage factors in other exposure pathways involving use of contaminated water (McDowell-Boyer et al., 2000; ORNL, 1997a). This conclusion is based, in part, on an assumption that the performance objective for the drinking water pathway would be substantially more restrictive than the performance objective for all exposure pathways combined (Section 3.4.2). Although analyses supporting this conclusion were based on generic transfer and usage factors, the conclusion would be incorrect only if these data would greatly underestimate intakes by pathways other than drinking water. This outcome appears unlikely at most sites, given that terrestrial foodchains are poor accumulators of most radionuclides under commonly occurring environmental conditions (Baes et al., 1984; NCRP, 1984a; 1999; Whicker, 1983; Whicker and Schultz, 1982). Thus, to the extent that acceptable disposals of radionuclides are controlled by allowable releases to groundwater or surface waters, uncertainties in models of exposure pathways other than drinking water would be unimportant in demonstrating compliance with performance objectives, and the only significant uncertainty in exposure pathway models would be the uncertainty in the consumption rate of water. The uncertainty in this parameter is only a factor of about two to three (Rupp, 1980; Yang and Nelson, 1986) and, thus, is unimportant. These considerations would not apply at very arid sites where releases to water may not occur. Use of specific-activity models in estimating dose from atmospheric releases of 3H and 14C or releases of 3H to water also involves some uncertainty, due primarily to uncertainties in estimating intakes of water or carbon dioxide from different environmental compartments, including compartments that are not contaminated by a release. However, uncertainties in estimated doses should be consistent with or less than those arising from use of standard exposure pathways models for other radionuclides and, thus, should be insignificant in an assessment of the overall uncertainty in the performance of a disposal facility.

282 / 5. PERFORMANCE ASSESSMENT MODELS 5.10.7.3 Sources of Uncertainty in Dose Coefficients. Uncertainties in dose coefficients for ingestion and inhalation arise primarily from uncertainties in the chemical form of radionuclides in the environment and the variability among individuals in organ masses and locations and in absorption, deposition and retention of radionuclides in the body. Uncertainties in dose coefficients for external exposure arise primarily from the variability among individuals in body size and uncertainties in calculating photon fields in air at a receptor location. Uncertainties in dose coefficients for internal exposure that account for the variability among individuals in organ masses, absorption of ingested or inhaled radionuclides into blood, deposition of absorbed radionuclides in different body organs, and retention of absorbed radionuclides in the body have been evaluated in detail for a few radionuclides, including 131I, 137Cs, and 239Pu (Apostoaei and Miller, 2004; Apostoaei et al., 1999; Dunning and Schwarz, 1981; Grogan et al., 2000; 2001; Matthies et al., 1981). However, this type of analysis has not been performed for some important radionuclides in low-level waste. A more general evaluation of uncertainties in dose coefficients for inhalation and ingestion of several radionuclides is given by NCRP (1998) (see also Bouville et al., 1994), and general discussions of the magnitude of uncertainties in dose coefficients for some radionuclides are given by Eckerman et al. (1999). The uncertainty in selected parameters used in calculating dose coefficients for internal exposure has been considered for some radionuclides. An example is the uncertainty in the GI-tract absorption fraction of plutonium and other transuranic elements (ICRP, 1986). Published analyses indicate that dose coefficients for internal exposure could be uncertain by about a factor of two to three in the best cases, but by as large as an order of magnitude or more in the worst cases. A general discussion of uncertainties in dose coefficients for external exposure is given by Eckerman et al. (1999). However, studies by ICRP (1987b; 1997c) indicate that these uncertainties should be unimportant in performance assessment, because uniform whole-body irradiation by higher-energy photons is the primary concern and the energy absorbed per unit body mass should not vary greatly among individuals in a population. 5.10.7.4 Significance of Uncertainties in Dose Coefficients. Uncertainties in dose coefficients could be substantial, particularly for internal exposure. However, uncertainties in dose coefficients for a particular radionuclide should be about the same at all disposal sites. Thus, for a given radionuclide composition of waste, an

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accounting of these uncertainties would not be useful in distinguishing between the disposal capabilities of different sites or facility designs. Furthermore, estimates of dose obtained in performance assessment apply to hypothetical exposure situations at far future times and to reference individuals, particularly reference adults, with prescribed characteristics (ICRP, 1975), and dose coefficients developed by ICRP can be viewed as one of the defining characteristics of reference individuals. Therefore, in evaluating compliance with performance objectives, it is reasonable not to ascribe uncertainty to dose coefficients. This is the usual approach, for example, in demonstrating compliance with dose limits for workers. Uncertainty analyses then could focus on models describing release of radionuclides and transport in the environment, which are the important factors in determining the waste-isolation capabilities of any disposal system. 5.10.8 Approaches to Estimating Risk An estimate of dose normally is the desired endpoint of a performance assessment, given that performance objectives for low-level waste disposal are expressed in terms of dose (Section 3.4.2). However, risks of stochastic health effects also could be calculated, if so desired. The simplest approach to estimating risk is to apply a risk coefficient, expressed as the stochastic risk per unit effective dose, to an estimate of effective dose from all radionuclides and exposure pathways combined. For purposes of risk management and risk assessment in general terms, ICRP (1991) and NCRP (1993) currently recommend a nominal risk coefficient for fatal cancers in the whole population of 0.05 Sv–1; a nominal risk coefficient for cancer incidence is ~50 % higher (EPA, 1994). These coefficients take into account the dependence of radiation risks on age at exposure and the age dependence of risks of death from all other causes. When applied to the dose from inhalation or ingestion, there is an implicit assumption that the committed dose due to an intake at a given age is received at the time of intake. Nominal risk coefficients also can be applied to estimates of effective dose equivalent. ICRP (1991) and NCRP (1993) also have developed a recommendation on the total stochastic detriment per unit effective dose for exposure of the whole population. The total detriment is the sum of three terms: (1) the risk of fatal cancers of 0.05 Sv–1 discussed above; (2) a weighted detriment for severe genetic effects of 0.01 Sv–1, which represents the genetic risk in all future generations; and (3) a weighted detriment for nonfatal cancers of 0.013 Sv–1, which is not

284 / 5. PERFORMANCE ASSESSMENT MODELS the risk of a nonfatal cancer. The total detriment thus is 0.073 Sv–1. The concept of total detriment is used mainly in developing the tissue weighting factors (wT) in the effective dose (Table 5.9), but it is not intended to be used in estimating risks from radiation exposure. EPA has developed an alternative method of estimating risk from exposure of the whole population to radionuclides in the environment (Eckerman et al., 1999). Instead of estimating risk on the basis of estimates of effective dose and a nominal risk coefficient, EPA calculated radionuclide- and pathway-specific risk coefficients expressed as the risk per unit activity intake by specified pathways or the risk per unit activity concentration in specified environmental media for external exposure pathways. EPA’s estimated risk coefficients assume that concentrations of radionuclides in the environment are constant over an individual’s lifetime, whereas the nominal risk coefficient used by ICRP and NCRP assumes that the dose rate is constant. EPA has provided radionuclide-specific risk coefficients for inhalation, ingestion of tap water, ingestion of food, and external exposure due to air submersion, exposure to a ground plane, and exposure to a soil volume (Eckerman et al., 1999). For inhalation and ingestion pathways, assumed intakes of contaminated environmental media (air, tap water or food) depend on age. Risk coefficients are given for both fatal cancers and cancer incidence. Methods used by EPA to calculate radionuclide- and pathwayspecific risk coefficients differ from the method used by ICRP (1991) to calculate a nominal risk per unit effective dose for any radionuclide and pathway in several respects (Eckerman et al., 1999; EPA, 1994; NAS/NRC, 1999b). In essence, EPA estimates risks due to chronic lifetime exposure as a convolution (integral) over age at exposure from birth to death and over time after exposure of: (1) absorbed dose rates in organs and tissues as a function of time after exposure, with estimates for intakes of radionuclides obtained using age-specific biokinetic and dosimetric models currently recommended by ICRP (Section 5.10.6.1.1); (2) estimates of risk at any future age per unit absorbed dose received at a given age, which differ in some respects from the nominal tissue weighting factors and radiation weighting factor for alpha particles used by ICRP (1991) in estimating effective dose; and (3) the probability of death from all competing causes as a function of age, as obtained from U.S. life tables. The risk at any future age per unit dose received at a given age is estimated using a relative-risk model for most organs and tissues, except an absolute-risk model is used for bone, skin and thyroid. The relative-risk model incorporates age-specific background cancer risks from all causes in the U.S. population.

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Differences between approaches used by EPA and ICRP to estimate risk are significant mainly in cases of intakes of long-lived, alpha-emitting radionuclides that have long retention times in the body. For example, for inhalation of 232Th, the fatal cancer risk estimated by EPA is about a factor of four less than the value obtained using ICRP’s effective dose coefficient for adults (ICRP, 1996b) and a nominal risk coefficient of 0.05 Sv–1. For ingestion of 232Th, the difference is about a factor of five, with EPA’s estimate again lower. For inhalation or ingestion of radionuclides with relatively short retention times in the body and for external exposure, EPA’s risk estimates are ~12 % higher than ICRP’s, due mainly to the longer average lifespan in the U.S. population compared with the population assumed by ICRP. Thus, differences in risk coefficients for fatal cancers estimated by ICRP and EPA could have a significant effect on estimates of risk from disposal of low-level waste only if most of the dose is due to inhalation or ingestion of long-lived, alpha-emitting radionuclides.

5.10.9 Summary Section 5.10 has discussed the development and implementation of exposure scenarios and exposure pathway models for use in performance assessments of low-level waste disposal facilities. These discussions have focused on dose assessment modeling for members of the public located outside the boundary of a disposal facility. Recommendations have been developed on environmental conditions and living habits of exposed individuals, the types of exposure scenarios that could be applied to different release pathways, use of simple, multiplicative-chain models for all exposure pathways, use of generic and site-specific data, and sources of generic data on dose coefficients, usage factors, and transfer factors that normally should be acceptable. Uncertainties in exposure pathway and dose assessment models also were discussed. Uncertainties in dose coefficients, transfer coefficients, and usage factors may be substantial (e.g., an order of magnitude or more in some cases), and these uncertainties, especially uncertainties in transfer coefficients and usage factors, generally should be considered in performance assessment. However, it is unlikely that uncertainties in exposure pathway models would be significant compared with uncertainties in predictions of releases of radionuclides from disposal facilities and transport in the environment. Furthermore, uncertainties in exposure pathway models should be about the same at any disposal site and, thus,

286 / 5. PERFORMANCE ASSESSMENT MODELS should not be important in distinguishing between the capabilities of disposal facilities at different sites or different types of disposal facilities or facility designs at the same site. Many of the recommendations on exposure pathway and dose assessment modeling are based essentially on the concept that performance assessments need only provide reasonable assurance that disposals will comply with applicable performance objectives (Section 3.5.3). Inherent in this concept is the view that performance assessment can be based on assumptions judged to be credible at each disposal site. Deliberately pessimistic assumptions about exposure pathways, representing conditions unlikely to be experienced, need not be assumed, particularly if such assumptions would result in severe restrictions on acceptable waste disposals. Furthermore, it is reasonable to standardize certain assumptions, particularly assumptions about current environmental conditions and living habits of exposed individuals, when exposure conditions at far future times are unknowable. Use of appropriate standardized assumptions in evaluating exposure pathways and dose focuses attention on the site-specific performance of engineered barriers and the natural environment, which are the parts of a disposal system that are important in providing waste containment and isolation. Development of exposure scenarios and exposure pathway models generally should be site-specific to the extent reasonable. The following statements provide examples of potentially important site-specific considerations: an assumption that an individual would use contaminated water to irrigate foodcrops or pasture grass probably is reasonable at relatively arid sites but may be unreasonable at humid sites; the importance of releases of radionuclides in particulate form from contaminated surface soil into the atmosphere may differ greatly at humid and arid sites; and, root uptake of radionuclides from soil by plants may depend on the soil type and concentrations of chemical homologs in soil (Whicker, 1983). Although site-specific considerations are important in developing and implementing exposure scenarios, standardization of some parts of dose assessment models for use at any site is appropriate. An important example is the use of standard dose coefficients for internal and external exposure of reference adults. In general, with proper justification, an appropriate mix of site-specific and generic assumptions about exposure scenarios, exposure pathways, pathway models, and model parameter values is reasonable. Any seeming inconsistencies in dose assessment models among different disposal sites should not be construed as indicating a lack of

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credibility in the assessments. Rather, if choices are properly justified, such differences are an inherent aspect of the process of developing site-specific performance assessment models. Recommendations on development and application of exposure scenarios and dose assessment models discussed in this Report place a considerable burden on the analyst to develop appropriate models and justify assumptions. Important issues that must be considered include the degree of conservatism to be built into models, including the impact of conservative assumptions on the acceptability of waste disposals at a given site, and the applicability of generic data or data obtained at the disposal site. However, these judgments are of the same kind as those that arise in all other aspects of performance assessment. NCRP believes that placing responsibility for these judgments on the analyst, rather than prescribing assumptions and calculational procedures to be used in dose assessment modeling at all sites, is the best approach to achieving reasonable and defensible results.

6. Inadvertent Human Intrusion In this Report, models for assessing inadvertent human intrusion at near-surface disposal facilities for low-level waste and potential radiation doses to intruders are treated separately from discussions in Section 5 on models for assessing the normal, undisturbed performance of disposal facilities and potential radiation doses to off-site members of the public. This distinction is made, in part, because assessments of inadvertent intrusion may not be required in all performance assessments. Furthermore, scenarios for inadvertent intrusion are fundamentally different from scenarios for undisturbed performance of disposal systems. Section 6 mainly discusses two topics concerned with inadvertent human intrusion. The first is the role of inadvertent intrusion in determining acceptable disposals of radioactive wastes, and the second is the types of scenarios for inadvertent intrusion that should be appropriate for use in performance assessments of near-surface disposal facilities for low-level waste. The types of exposure pathway models discussed in Section 5.10 generally should be applicable to exposures of inadvertent intruders.

6.1 Introduction

Near-surface disposal facilities for low-level waste generally are required to provide protection of future inadvertent intruders (Section 3.4.2). This requirement is implemented, in part, by establishing waste acceptance criteria in the form of limits on concentrations of radionuclides that are developed on the basis of assessments of dose in assumed scenarios for inadvertent intrusion. Concentration limits can be based on generic assessments of inadvertent intrusion that are applied at all disposal sites, as in NRC’s waste classification system in 10 CFR Part 61 (NRC, 1982b), or they can be established on a site-specific basis, as in DOE’s approach to determining acceptable near-surface disposals of low288

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level waste. Discussions on the development and application of scenarios for inadvertent intrusion at near-surface disposal facilities for low-level waste in Section 6 are based primarily on the following considerations. First, in determining acceptable near-surface disposals of low-level waste, requirements for protection of inadvertent intruders generally are considered separately from requirements for protection of off-site individuals or populations. This distinction is considered appropriate primarily because exposures of inadvertent intruders according to the types of scenarios normally assumed in performance assessment reasonably could occur only if some form of institutional control over disposal sites or societal memory of past disposals is not maintained, whereas exposures of off-site members of the public could occur at any time after disposal, without regard for institutional controls or societal memory. Second, exposure scenarios for inadvertent intruders often assumed in performance assessment should be regarded as largely hypothetical constructs that provide reasonable upper bounds on radiological impacts that could be experienced by individuals who might actually intrude onto disposal sites at some time in the future. The role of assumed intrusion scenarios is to provide information to support design and control measures to preclude intrusion or to limit impacts should inadvertent intrusion occur. Third, given the intention to maintain some form of institutional control or societal memory over disposal sites for the foreseeable future, exposures of inadvertent intruders are unlikely to occur according to the types of chronic exposure scenarios commonly assumed in performance assessment (i.e., inadvertent intrusion can be regarded as an accidental occurrence). Nonetheless, use of such scenarios to determine acceptable disposals provides an important form of defense-in-depth in protecting the public should institutional controls fail to function as intended and societal memory of the disposal facility not be maintained. Fourth, given that criteria for protection of inadvertent intruders at near-surface disposal facilities for low-level waste generally are expressed only qualitatively or in terms of dose (Section 3.4.2), probabilities of occurrence of intrusion scenarios usually are not considered in performance assessment. Rather, reasonable bounding scenarios are assumed to occur with a probability of unity, and radiological consequences of assumed scenarios are evaluated. Recommendations of ICRP (1998) on use of risk, taking into account probabilities of occurrence of assumed scenarios, are discussed in Section 3.3.3.

290 / 6. INADVERTENT HUMAN INTRUSION Fifth, many different scenarios for inadvertent intrusion could be postulated. However, different scenarios should be considered to represent mutually exclusive exposure situations, and calculated doses for different scenarios generally should not be added. This consideration is based primarily on an assumption that any individual who might intrude onto a disposal site is unlikely to be a participant in more than one assumed scenario. Finally, standardized approaches to evaluating inadvertent intrusion at near-surface disposal facilities for low-level waste, such as those used in developing NRC’s waste classification system in 10 CFR Part 61 (NRC, 1981a; 1982a; Oztunali and Roles, 1986; Oztunali et al., 1986), have been developed, and these approaches often are appropriate for use in site-specific performance assessments. However, generic approaches should be applied with caution at particular disposal sites, and only with proper consideration of the applicability of assumptions to site-specific conditions. The remainder of Section 6 is organized as follows. The historical role of inadvertent human intrusion in radioactive waste disposal and current regulatory requirements for protection of inadvertent intruders at near-surface disposal facilities for lowlevel waste are reviewed in Section 6.2. Largely generic scenarios for inadvertent intrusion at near-surface disposal facilities that have been widely used in performance assessment are described in Section 6.3, and approaches to selecting site-specific scenarios are discussed in Section 6.4. Required inputs to dose assessments for inadvertent intruders are described in Section 6.5, and outputs of those assessments and their use developing site-specific waste acceptance criteria in the form of limits on concentrations of radionuclides are discussed in Section 6.6. A brief discussion of how effects of inadvertent human intrusion on off-site releases of radionuclides are normally taken into account in performance assessments is given in Section 6.7. Recommended approaches to considering inadvertent intrusion in performance assessment are summarized in Section 6.8. Particular models that could be used to obtain estimates of dose or risk to inadvertent intruders, on the basis of assumed exposure scenarios, are not discussed in Section 6. Rather, discussions emphasize conceptual approaches that could be used to obtain such estimates and the types of subjective judgments involved. Implementation of particular exposure scenarios and exposure pathway models generally should be the same for inadvertent intruders as for off-site members of the public, and suitable approaches are discussed in Section 5.10.

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6.2 Role of Inadvertent Human Intrusion in Radioactive Waste Disposal 6.2.1

Historical Perspective

The development of strategies for disposal of radioactive wastes focuses on the issue of determining what kinds of disposal systems generally should be acceptable for different types of waste. Many technical, economic and social factors must be considered in addressing this issue. Beginning about three decades ago, the concept of a hypothetical inadvertent intruder onto waste disposal sites following loss of active institutional control, as distinct from an off-site member of the public who might be exposed at any time after disposal, has played a central role in distinguishing between waste containing relatively low concentrations of radionuclides that should be generally acceptable for near-surface disposal and waste containing higher concentrations of radionuclides, particularly those with relatively long half-lives, that should require disposal far below the ground surface (e.g., in a geologic repository). In the United States, the concept that waste disposal systems should provide protection of inadvertent intruders was first used in establishing policies for disposal of solid waste containing alpha-emitting radionuclides. In 1970, AEC established a policy that solid waste with concentrations of alpha-emitting radionuclides >0.4 kBq g–1 was unacceptable for shallow-land disposal, but required burial or storage in a retrievable manner (Hollingsworth, 1970). The concentration limit for shallow-land disposal of waste containing alpha-emitting radionuclides was based on higher concentrations of naturally occurring radium in the Earth’s crust. That is, shallow-land disposal of waste containing alpha-emitting radionuclides in concentrations <0.4 kBq g–1 was judged acceptable when doses to an individual who might access waste should not be significantly greater than the unavoidable dose due to naturally occurring radium and its decay products in surface soil, excluding the dose due to radon. In the late 1970s, an explicit distinction between inadvertent intruders and off-site members of the public was used for the first time in a study of disposal systems for low-level waste (Adam and Rogers, 1978). That study was an important step in NRC’s development of licensing criteria for near-surface disposal of lowlevel waste in 10 CFR Part 61 (NRC, 1982b). As discussed in Section 3.4.2.1 and the following section, NRC’s licensing criteria include a requirement to protect future inadvertent intruders.

292 / 6. INADVERTENT HUMAN INTRUSION Thus, the concept was developed that radioactive waste would be generally acceptable for near-surface disposal if concentrations of radionuclides were sufficiently low that potential exposures of inadvertent intruders would be acceptable, but waste would require deeper disposal if concentrations of radionuclides were sufficiently high that potential exposures of inadvertent intruders at near-surface disposal facilities would be unacceptable. Inherent in an assumption that deeper disposal would be generally acceptable for higher-activity wastes was the view that normal human activities would not likely result in inadvertent intrusion into waste located far below the ground surface. In addition, the concept of inadvertent intrusion, as it was developed initially, involved scenarios for direct intrusion into solid waste in a disposal facility. That is, the concept was used only to determine allowable residual concentrations of radionuclides in a disposal facility following loss of active institutional control over a disposal site. Only protection of inadvertent intruders generally has been of concern in distinguishing between waste that is acceptable for near-surface disposal and waste that requires disposal far below the ground surface. Protection of individuals who might deliberately intrude into disposal facilities, with knowledge of the presence of waste and the potential hazards involved, has never been a concern in determining acceptable disposals of radioactive waste (NRC, 1982b). As described in Section 3.4.2 and discussed further in the following section, the concept that near-surface disposal facilities should protect future inadvertent intruders is now an integral part of regulatory requirements for disposal of low-level waste in the United States.

6.2.2

Regulatory Requirements for Near-Surface Disposal

The following sections discuss aspects of NRC and DOE requirements related to protection of inadvertent intruders at nearsurface disposal facilities for low-level waste.

6.2.2.1 NRC Requirements. NRC’s licensing requirements for near-surface disposal of low-level waste in 10 CFR Part 61 (NRC, 1982b) include a qualitative performance objective for protection of inadvertent intruders. This requirement is implemented, in part, by NRC’s waste classification system, which includes generally

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applicable limits on concentrations of radionuclides in Class-A, -B and -C wastes (Table 3.3). These concentration limits were developed on the basis of (1) the types of scenarios for inadvertent intrusion discussed in Section 6.3, (2) assumed dose criteria that were consistent with radiation protection standards for the public at that time (AEC, 1960), (3) an assumption that active institutional control over disposal sites would preclude inadvertent intrusion for 100 y after disposal, (4) assumptions about the ability of engineered barriers and disposal well below the ground surface to preclude inadvertent intrusion into waste or limit the probability of intrusion following an assumed loss of active institutional control, and (5) data on expected concentrations of radionuclides in non-DOE low-level waste (NRC, 1982a). NRC’s approach of protecting inadvertent intruders by means of a generally applicable waste classification system was intended, in part, to obviate the need for site-specific analyses of exposure scenarios for inadvertent intruders in licensing disposal facilities. NRC essentially assumed that generic assessments of exposure scenarios would provide adequate protection of inadvertent intruders at any disposal site. NRC’s regulatory approach in 10 CFR Part 61 clearly presumes that it is reasonable to distinguish between inadvertent intruders and off-site members of the public in determining acceptable nearsurface disposals of low-level waste. This distinction is evidenced by the assumption, described in Section 3.4.2.1, of higher dose criteria for inadvertent intruders than for off-site members of the public (NRC, 1982a). NRC also considered that exposures of inadvertent intruders at licensed near-surface disposal facilities in accordance with scenarios assumed in developing the waste classification system were accidental, unlikely occurrences. On the other hand, the dose criteria used in developing the waste classification system are sufficiently low that risks to inadvertent intruders, taking into account the probability that an assumed scenario would occur as well as the dose in such a scenario, still should be acceptable. Although NRC intends that dose assessments for inadvertent intruders would not be needed in licensing specific disposal sites, such assessments could be required by Agreement States that license disposal of low-level waste. This could be the case, for example, if radionuclide compositions of wastes sent to a particular facility differed substantially from those assumed by NRC in developing the waste classification system, particularly if wastes contained substantial amounts of radionuclides (e.g., uranium) not included in the waste classification system, or if a disposal technology significantly different from that considered in developing the waste classification system were used.

294 / 6. INADVERTENT HUMAN INTRUSION 6.2.2.2 DOE Requirements. Requirements for disposal of low-level waste at DOE sites in Order 435.1 (DOE, 1999a; 1999b), as well as the previous Order 5820.2A (DOE, 1988a) that applied at some disposal facilities, do not include limits on concentrations of radionuclides that would be generally acceptable for near-surface disposal, as in NRC’s waste classification system discussed above. However, DOE essentially requires that doses to inadvertent intruders should be limited in accordance with radiation protection standards for the public (DOE, 1990; NRC, 1991a), and that a sitespecific assessment of doses to inadvertent intruders shall be performed for the purpose of developing waste acceptance criteria in the form of limits on concentrations of radionuclides. The concentration limit for any radionuclide could be site-specific, depending on differences in assumptions that might be used in developing and implementing site-specific scenarios for inadvertent intrusion. DOE requirements for disposal of low-level waste were developed after NRC’s waste classification system had been established. DOE’s decision not to adopt NRC’s generally applicable limits on concentrations of radionuclides for near-surface disposal was based primarily on two considerations. First, NRC’s waste classification system was based to a significant extent on expected radionuclide compositions in non-DOE low-level wastes. However, many important DOE wastes do not resemble non-DOE wastes and such wastes would not be readily accommodated by NRC’s waste classification system; an important example is uranium-bearing wastes. Second, DOE intends to dispose of its low-level waste at known sites having differing characteristics and using a variety of technologies. Therefore, it is unreasonable to impose the same disposal limits at all DOE sites when credible scenarios for inadvertent intrusion could depend significantly on the particular site and disposal technology and any generally applicable limits presumably would be based on scenarios developed for relatively unfavorable disposal sites and technologies. However, the approach to regulating disposal of low-level waste taken by DOE is similar to NRC’s in that a clear distinction is made between protection of off-site members of the public and protection of inadvertent intruders. In addition, DOE and NRC both recognize that inadvertent intrusion in accordance with the widely assumed scenarios discussed in Section 6.3 would be an accidental, rather than expected, occurrence at licensed near-surface disposal facilities, because some form of institutional control over disposal sites or societal memory of past disposals is expected to be maintained for the foreseeable future.33

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6.3 Widely Used Scenarios for Inadvertent Intrusion Scenarios for inadvertent intrusion into near-surface disposal facilities that have been widely used in performance assessment are based primarily on assumptions used in developing NRC’s waste classification system in 10 CFR Part 61 (NRC, 1981a; 1982a; Oztunali and Roles, 1986; Oztunali et al., 1986). These scenarios involve direct intrusion, or attempts at direct intrusion, into waste disposal facilities by individuals who are assumed to come onto a disposal site after loss of institutional control. An implicit assumption in all scenarios is that an intruder has no knowledge of the existence of a disposal facility. Widely used scenarios for inadvertent intrusion can be divided into those involving a single, acute exposure during the time that an initial contact with waste or a disposal facility occurs and those involving long-term, chronic exposure during permanent residence on a disposal site after waste or a disposal facility has been accessed. Implicit in scenarios for chronic exposure is an assumption that active and passive institutional controls are no longer effective. This Section discusses widely used scenarios for acute and chronic exposure of inadvertent intruders and other scenarios that could be considered. The intent is to discuss basic assumptions used in developing intrusion scenarios. Methods of implementing scenarios to develop quantitative estimates of dose or risk to inadvertent intruders are discussed in Section 5.10. 6.3.1

Scenarios for Acute Exposure

Three scenarios involving a single, acute exposure of inadvertent intruders, often referred to as the construction, discovery and drilling scenarios, have been widely used. These scenarios are summarized in Table 6.1 and described in the following sections. 6.3.1.1 Construction Scenario. In the construction scenario, an inadvertent intruder is assumed to excavate directly into solid waste in a disposal facility while digging a foundation for a home. During excavation activities, an intruder normally is assumed to 33

An assumption that some form of institutional control will be maintained is indicated, for example, by discussions in Chapter 4 of NRC’s draft environmental impact statement for 10 CFR Part 61 (NRC, 1981a) and a provision in DOE requirements that assessments of potential impacts on inadvertent intruders should assume intrusion “for a temporary period” (DOE, 1999b).

296 / 6. INADVERTENT HUMAN INTRUSION TABLE 6.1—Widely used scenarios for single, acute exposure of inadvertent intruders often assumed in performance assessments of low-level waste disposal facilities.a Scenario

Description

Construction

While excavating foundation for a home on a disposal site, inadvertent intruder accesses solid waste and receives external, inhalation and ingestion exposures.

Discovery

While excavating foundation for a home on a disposal site, inadvertent intruder encounters intact engineered barrier that precludes access to solid waste and receives external exposure. Scenario generally applies only to disposal facilities constructed with engineered barriers that are impenetrable by normal methods of excavation.

Drilling

While drilling a well on a disposal site, inadvertent intruder brings solid waste to ground surface and receives external, inhalation and ingestion exposures.

aScenarios are based largely on assumptions used in developing NRC’s generally applicable waste classification system in 10 CFR Part 61 (NRC, 1981a; 1982a; Oztunali and Roles, 1986; Oztunali et al., 1986).

receive exposure by three pathways: (1) external exposure to photon-emitting radionuclides in waste, (2) inhalation of radionuclides released into the air, and (3) ingestion of radionuclides transferred to the hands. The exposure time for the construction scenario normally is assumed to be only a few days. 6.3.1.2 Discovery Scenario. In the discovery scenario, an inadvertent intruder is assumed to excavate at a disposal site, as in the construction scenario described above. However, direct access to waste is assumed to be precluded by the presence of intact engineered barriers that are not readily penetrated by the types of excavating equipment normally used at a site. Thus, the discovery scenario is relevant when a disposal facility is constructed with high-integrity engineered barriers (e.g., reinforced concrete) or when highintegrity waste packages are used, but the scenario would not apply, for example, to disposal of unpackaged waste in shallow trenches capped only with soil or other easily penetrable materials.

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The discovery scenario is assumed to occur during a short period of time (a few days or less) that an intruder attempts to excavate into a disposal facility constructed with impenetrable engineered barriers, and it is relevant only for as long as the barriers are assumed to remain intact (perhaps a few hundred years). Since waste in a facility is assumed not to be directly accessed during excavation activities, the only exposure pathway for this scenario is external exposure to photon-emitting radionuclides in waste while working next to engineered barriers. 6.3.1.3 Drilling Scenario. In the drilling scenario, an inadvertent intruder is assumed to drill through solid waste in a disposal facility while constructing a well for a domestic water supply. Contaminated drilling waste is assumed to be brought to the surface and, during the course of drilling activities, the intruder is assumed to receive external, inhalation and ingestion exposures to radionuclides in a pile of drilling wastes. The exposure time for the drilling scenario normally is assumed to be only a few days. 6.3.2

Scenarios for Chronic Exposure

Three scenarios involving long-term, chronic exposure of inadvertent intruders, often referred to as the agriculture (or homesteader); resident, nonagriculture; and postdrilling scenarios, have been widely used. These scenarios are summarized in Table 6.2 and described in the following sections. 6.3.2.1 Agriculture Scenario. The agriculture (or homesteader) scenario is assumed to occur after an inadvertent intruder has established a permanent and largely self-sufficient homestead on a disposal site. In many applications of this scenario, some waste exhumed from a facility during excavation of the foundation for a home as a result of the construction scenario described in Section 6.3.1.1 is assumed to be mixed with native soil in a vegetable garden, and the intruder’s home is assumed to be located directly on top of exposed waste in the disposal facility. In addition, disposed waste is assumed to be indistinguishable from native soil at the time excavation occurs. On the basis of these assumptions, exposure pathways frequently assumed in the agriculture scenario include: (1) ingestion of vegetables grown in contaminated garden soil; (2) direct ingestion of contaminated soil, either from the garden in conjunction with vegetable intakes or from the vicinity of the home during normal yard activities; (3) external exposure to contaminated soil while working in the garden or residing in the home

298 / 6. INADVERTENT HUMAN INTRUSION TABLE 6.2—Widely used scenarios for chronic exposure of inadvertent intruders often assumed in performance assessments of low-level waste disposal facilities.a Scenario

Agriculture (homesteader)

Description

Following construction of a home on top of disposed waste, with some exhumed waste mixed with soil in a vegetable garden, inadvertent intruder consumes contaminated vegetables and contaminated soil from garden or vicinity of home and receives external and inhalation exposures while working in garden and residing in home. In some applications, inadvertent intruder obtains contaminated water from on-site well used for drinking water, irrigation of vegetable garden, consumption by dairy and beef cattle supplying milk and meat, and irrigation of pasture grass consumed by cattle, and receives external, ingestion and inhalation exposures to radionuclides originating in the water.

Resident, nonagriculture

Following construction of a home directly on top of intact engineered barrier in disposal facility, inadvertent intruder receives external exposure while residing in home. In some applications, inadvertent intruder receives exposures from use of contaminated well water as in agriculture scenario. Scenario generally applies only to disposal facilities constructed with engineered barriers that are impenetrable by normal methods of excavation.

Postdrilling

Following drilling of a well through disposed waste, with drilling waste mixed with soil in a vegetable garden, inadvertent intruder consumes contaminated vegetables and contaminated soil from garden and receives external and inhalation exposures while working in garden. In some applications, inadvertent intruder receives exposures from use of contaminated well water as in agriculture scenario.

a Scenarios are based largely on assumptions used in developing NRC’s generally applicable waste classification system in 10 CFR Part 61 (NRC, 1981a; 1982a; Oztunali and Roles, 1986; Oztunali et al., 1986).

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on top of the disposal facility; and (4) inhalation of radionuclides released into the air from contaminated soil while working in the garden or residing in the home. The agriculture scenario described above assumes that direct intrusion into residual solid waste in a disposal facility is the primary concern in evaluating potential exposures of inadvertent intruders. In some applications of this scenario, additional exposure pathways are considered on the basis of an assumption that an intruder’s domestic water supply is obtained from a contaminated well on the disposal site. For example, exposure to contaminated groundwater was considered by NRC in developing concentration limits for 14C and 129I in the waste classification system in 10 CFR Part 61 (NRC, 1982a). Particular exposure pathways often considered in such a scenario include: (1) direct ingestion of contaminated water, (2) ingestion of vegetables irrigated with contaminated water, (3) ingestion of milk and meat obtained from dairy and beef cattle that drink contaminated water or consume pasture grass irrigated with contaminated water, (4) direct ingestion of soil contaminated by irrigation, (5) external exposure to the contaminated soil, and (6) inhalation of radionuclides released into the air from the contaminated soil. 6.3.2.2 Resident, Nonagriculture Scenario. As is the case of the discovery scenario discussed in Section 6.3.1.2, the resident, nonagriculture scenario is relevant only when a disposal system is constructed with high-integrity engineered barriers. An example would be a below-grade facility with a reinforced concrete roof. In this scenario, an inadvertent intruder is assumed to reside in a home located directly on top of an impenetrable engineered barrier above the waste. The only exposure pathway of concern is external exposure to photon-emitting radionuclides in the waste while residing in the home. The resident, nonagriculture scenario also may consider exposures of an inadvertent intruder to contaminated groundwater obtained from a well on the disposal site. The exposure pathways in this case would be the same as those described above for the agriculture scenario. 6.3.2.3 Postdrilling Scenario. The postdrilling scenario describes exposures of a permanent resident on a disposal site to contaminated drilling waste exhumed during construction of a well as a result of the drilling scenario described in Section 6.3.1.3. No other means of accessing residual solid waste in a disposal facility is assumed to occur. Since the volume of waste exhumed by drilling

300 / 6. INADVERTENT HUMAN INTRUSION should be considerably less than the volume of waste exhumed in the construction scenario described previously, the postdrilling scenario often assumes that all drilling waste is mixed with native soil in a vegetable garden. Then, as in the agriculture scenario, an intruder is assumed to receive exposures from several pathways including: (1) ingestion of vegetables grown in contaminated garden soil, (2) direct ingestion of contaminated soil from the garden in conjunction with vegetable intakes, (3) external exposure to contaminated soil while working in the garden, and (4) inhalation of radionuclides released into the air from contaminated soil while working in the garden. The postdrilling scenario also may consider exposures of an inadvertent intruder to contaminated groundwater obtained from a well on the disposal site. The exposure pathways in this case again would be the same as those described previously for the agriculture scenario. 6.3.2.4 Groundwater Pathway for Chronic Intrusion Scenarios. As indicated in the previous three sections, some applications of the widely used scenarios for chronic exposure of inadvertent intruders include exposure to contaminated groundwater on a disposal site, as well as exposure to residual solid waste in a disposal facility. There have been two points of view about whether exposure to contaminated groundwater should be included in those scenarios. The view that exposure to contaminated groundwater should be included in scenarios for chronic exposure of inadvertent intruders is based on the general requirement that near-surface disposal systems must provide protection of inadvertent intruders and an interpretation that this requirement applies to inadvertent intrusion onto disposal sites, not just to intrusion into disposal facilities themselves. That is, dose assessments for inadvertent intruders should consider all credible exposure pathways at a disposal site for the assumed exposure scenarios. The opposite view is based on an interpretation that requirements for protection of inadvertent intruders basically are concerned with limiting potential exposures to radionuclides remaining in a disposal facility at any time when intrusion onto a disposal site is considered to be a credible occurrence, and that limiting releases from a disposal facility into groundwater is a separate issue of concern mainly in regard to protecting off-site members of the public. With this assumption, exposure of inadvertent intruders to contaminated groundwater on a disposal site, although not considered explicitly, is assumed to be limited to acceptable levels by other performance objectives that apply to

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releases beyond the site boundary. This view is consistent with the general idea that determining acceptable near-surface disposals of radioactive waste essentially involves achieving an appropriate balance between allowable releases beyond a site boundary to protect the general public and the environment and allowable residual levels in a disposal facility after loss of institutional control to protect inadvertent intruders. If exposure to contaminated groundwater on a disposal site is considered in scenarios for inadvertent intrusion, the required assessment is more complicated than if such exposure is not considered, because time histories of radionuclide concentrations in groundwater will, in general, be quite different from time histories of concentrations remaining in a disposal facility. Therefore, maximum doses from exposure to contaminated groundwater on a disposal site should occur at different times than maximum doses from direct intrusion into solid waste in a disposal facility. As a result, in evaluating the maximum dose from all exposure pathways combined, the two doses would not be additive. 6.3.3

Comparison of Standard Scenarios for Inadvertent Intrusion

For most disposal systems, only one or two of the widely used scenarios for acute or chronic exposure of inadvertent intruders described in Sections 6.3.1 and 6.3.2 should be the most restrictive in determining limits on concentrations of radionuclides that would be acceptable for disposal. Generic and site-specific analyses have indicated that if institutional controls over a disposal site are assumed not to be effective beyond some time after disposal (e.g., 100 y), scenarios for acute exposure usually are less important than scenarios for chronic exposure (Aaberg and Kennedy, 1990; Cook et al., 2002; Kennedy and Peloquin, 1988; McDowell-Boyer et al., 2000; NRC, 1981a; 1982a; ORNL, 1997a). Furthermore, among the widely used scenarios for chronic exposure, the resident, nonagriculture scenario usually is less important than the agriculture or postdrilling scenario. However, exceptions can occur. For example, if requirements for protection of inadvertent intruders were applied for only a finite period of time and engineered barriers were assumed to preclude excavation or drilling into waste during that time, a resident, nonagriculture scenario could be the most important for photon-emitting radionuclides (Cook et al., 2002; McDowell-Boyer et al., 2000). Thus, dose assessments for inadvertent intruders often would need to consider only the chronic agriculture and postdrilling

302 / 6. INADVERTENT HUMAN INTRUSION scenarios. The relative importance of these two scenarios depends primarily on assumptions about the ability of engineered barriers to deter intrusion into waste by excavation or drilling. Suppose, for example, that both excavation and drilling into a disposal facility reasonably could occur at the same time, either because no engineered barriers are used or because neither excavation nor drilling would be credible until such barriers have lost their integrity. In this case, an agriculture scenario should result in higher estimates of dose for any radionuclide than a postdrilling scenario, because greater volumes of waste normally are assumed to be exhumed in an agriculture scenario, thus resulting in higher doses from exposure to exhumed material when it is mixed with native soil in a vegetable garden. In addition, exposure during indoor residence on top of exposed waste in a disposal facility usually is assumed to occur in an agriculture scenario but not in a postdrilling scenario. On the other hand, suppose that a postdrilling scenario could occur prior to an agriculture scenario, because drilling through a disposal facility would be a credible occurrence even in the presence of intact engineered barriers but excavation into waste would not. In such a situation, a postdrilling scenario could result in higher doses for shorter-lived radionuclides (e.g., 90Sr) that would decay appreciably over an assumed lifetime of engineered barriers. The relative importance of the agriculture and postdrilling scenarios thus depends on assumptions about the time after disposal at which these scenarios could occur and the composition of waste. A postdrilling scenario usually could be the more important only if drilling into waste could occur well before excavation and significant quantities of relatively short-lived radionuclides would be present in waste at the time of drilling. 6.3.4

Other Scenarios for Inadvertent Intrusion

Scenarios for inadvertent intrusion described in Sections 6.3.1 and 6.3.2 involve three implicit assumptions: (1) all active and passive institutional controls over disposal facilities will be lost at some future time (nominally 100 y); (2) disposal facilities are located on otherwise undisturbed land that would revert to its natural state following loss of institutional control; and (3) disposal sites would be suitable for residential use. However, other scenarios involving different future land uses could be assumed. For example, at sites subject to remediation under CERCLA (Superfund), scenarios for industrial, commercial or recreational land use often are assumed on the basis of current uses of such sites (Means, 1989). Similar scenarios have been considered in dose assessments

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to support rulemakings on cleanup of radioactively contaminated sites and facilities (EPA, 1993b; Kennedy and Strenge, 1992; NRC, 1997).34 In general, alternative scenarios for inadvertent intrusion involving, for example, industrial, commercial or recreational land use probably do not need to be considered in performance assessments of low-level waste disposal facilities. Because these types of scenarios would exclude many exposure pathways that are important in the widely used scenarios for residential use discussed previously (e.g., pathways involving consumption of contaminated foodstuffs obtained from a disposal site) and because permanent residence on an industrial, commercial or recreational site should not occur, scenarios involving residential use generally should result in higher estimates of dose than alternative scenarios and, thus, should be more restrictive in determining acceptable disposals. In addition, at sites that are not remote from populated areas at the present time, it may be difficult to justify an assumption that residential use of disposal sites would not be a credible occurrence at times far in the future if institutional controls are lost. Scenarios for inadvertent intrusion normally used in performance assessment are based on an assumption that permanent residence on disposal sites would be a credible occurrence at any time after loss of active institutional control. However, various forms of passive institutional control described in Section 3.5.1.2 could be effective in precluding permanent residence for some time after active institutional controls are removed, although they may not preclude limited access to disposal sites over short time periods (e.g., by hunters, campers or accidental trespassers). If passive institutional controls were assumed to be effective in precluding permanent residence on a disposal site, it would be reasonable to consider scenarios involving only short-term exposure. Such scenarios generally should result in lower estimates of dose than scenarios involving permanent residence. An assumption that passive institutional controls would allow only short-term exposures of inadvertent intruders has not been applied in developing waste acceptance criteria at specific sites. The foregoing discussions indicate that the widely used scenarios for inadvertent intrusion discussed in Sections 6.3.1 and 6.3.2 may not be appropriate for unusual types of facilities or when other 34EPA (1994). U.S. Environmental Protection Agency. “40 CFR Part 196 – Environmental Protection Agency radiation site cleanup regulation,” unpublished draft Federal Register notice (June 2) (U.S. Environmental Protection Agency, Washington).

304 / 6. INADVERTENT HUMAN INTRUSION assumptions, such as the presence of passive institutional controls that would be effective in preventing scenarios for long-term, chronic exposure, are appropriate. In such cases, reasonable scenarios for inadvertent intrusion can be developed using principles of scenario selection discussed in the following section. 6.4 Selection of Site-Specific Scenarios for Inadvertent Intrusion This Section discusses the issue of selecting and implementing exposure scenarios for inadvertent intrusion into near-surface disposal facilities that would be appropriate for use in performance assessment. Again, intrusion scenarios are used to develop waste acceptance criteria in the form of limits on concentrations of radionuclides. This discussion is concerned primarily with selection of scenarios for application to specific sites and facility designs, rather than their application to development of generally applicable requirements [e.g., as in NRC’s 10 CFR Part 61 (NRC, 1982a; 1982b)]. 6.4.1

Application of Widely Used Scenarios to Site-Specific Assessments

In most cases, widely used scenarios for inadvertent intrusion into near-surface disposal facilities described in Sections 6.3.1 and 6.3.2 should be applicable, at least in a general way, to site-specific performance assessments, because those scenarios are intended to apply to many disposal sites and facility designs (NRC, 1981a). Furthermore, those scenarios should provide higher estimates of dose to inadvertent intruders than many other scenarios that might be considered (i.e., the widely used scenarios should be somewhat conservative). However, in site-specific assessments of inadvertent intrusion, such as those required by DOE (1988a; 1999b) or that may be required by Agreement States, the widely used scenarios for inadvertent intrusion should not be applied uniformly at all disposal sites. That is, assumptions about the time when scenarios occur, exposure pathways for each scenario, and model parameter values for each exposure pathway should not necessarily be the same at all sites. Although uniform application of exposure scenarios is appropriate in developing generally applicable disposal requirements, this approach would result in the same concentration limits of radionuclides at any site and site- and design-specific factors that are potentially important in determining potential doses to inadvertent

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intruders would not be taken into account. Furthermore, a uniform approach to assessments of inadvertent intrusion at all sites presumably would be based on assumptions that are appropriate for relatively unfavorable sites and designs with minimal engineered features (NRC, 1981a; 1982a), and such an approach could result in considerable underestimates of disposal capabilities of more favorable sites and more robust engineered disposal systems. Such underestimates could have important consequences if, as a result of using unreasonably pessimistic assumptions, significant amounts of waste would be unacceptable for disposal. 6.4.2

Judgmental Factors in Selecting Exposure Scenarios

Selection and implementation of scenarios for inadvertent intrusion in either generic or site-specific applications essentially involves largely subjective judgments about several factors, which are summarized in Table 6.3. In addition to an assumption about the earliest time after disposal when inadvertent intrusion would be a credible occurrence, which usually is defined by an assumed time period for active institutional control over a disposal site, judgments must be made about the following factors: • the ability of engineered barriers and methods of waste emplacement to deter access to waste by inadvertent intruders beyond an assumed period of active institutional control; • particular scenarios for inadvertent intrusion that could occur (i.e., how an intruder could access, or attempt to access, waste in a disposal facility or elsewhere on a disposal site); and • particular actions that an inadvertent intruder might take after waste is accessed or approached, based primarily on consideration of environmental conditions at a disposal site. Additional assumptions about the effectiveness of any passive institutional controls that might be maintained after loss of active institutional control (e.g., warning signs, land-use records) could affect all three of these factors. In making the kinds of judgments described above, assumed scenarios for inadvertent intrusion and assumptions about exposure pathways and model parameter values should be consistent with: (1) the design of a disposal facility, particularly the expected long-term integrity of engineered barriers against the normal kinds of excavation and drilling activities that could occur at a site; (2) planned waste emplacements, such as use of deeper disposal or

306 / 6. INADVERTENT HUMAN INTRUSION TABLE 6.3—Judgmental factors involved in developing and implementing exposure scenarios for inadvertent intruders. Factor

Description

Time of intrusion

Time after disposal when inadvertent intrusion onto disposal site could occur, usually defined by assumed time period for active institutional control.a

Effect of engineered barriers and waste emplacements

Ability of engineered barriers and methods of waste emplacement to deter access to waste in disposal facility beyond assumed period of active institutional control.

Actions by inadvertent intruder in accessing waste

How inadvertent intruder could access, or attempt to access, waste in disposal facility or at other locations on disposal site (e.g., in groundwater).

Actions by inadvertent intruder after waste is accessed

How inadvertent intruder could be exposed after waste in disposal facility or at other locations on disposal site is accessed.

Passive institutional controls

After loss of active institutional control, ability of passive institutional controls to deter access to waste or disposal site, or to affect actions by inadvertent intruder in accessing waste or after waste is accessed.

aMinimum

period of active institutional control of 100 y after facility closure normally is assumed (DOE, 1988a; 1999b; NRC, 1982b).

disposal beneath other wastes as a deterrent to inadvertent intrusion (NRC, 1981a); and (3) local environmental conditions that would affect reasonable actions by an inadvertent intruder while attempting to access waste or reside permanently on a site. In essence, dose assessments for inadvertent intruders should be based on assumptions about scenarios, exposure pathways, and model parameter values that are credible at a particular site and for a particular facility design. Although assumptions may be conservatively biased, as is normally the case in the widely used scenarios discussed in Section 6.3, use of worst-case or highly unlikely assumptions about inadvertent intrusion should be avoided. The kinds of judgments that must be applied in selecting credible exposure scenarios for inadvertent intruders are illustrated by the following examples.

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In performance assessments of low-level waste disposal facilities at DOE’s Oak Ridge and Savannah River sites (Cook et al., 2002; McDowell-Boyer et al., 2000; ORNL, 1997a), an attempt by an inadvertent intruder to drill through a facility was judged to be a credible occurrence. However, even though disposal facilities at both sites were assumed to be constructed with similar types of engineered barriers, a drilling scenario was evaluated differently at the two sites by taking into account important differences in local environmental conditions. The disposal facility at Oak Ridge is located on a hard-rock formation (ORNL, 1997a). Given this condition, an assumption was made that the types of equipment normally required in drilling through rock at that site could easily penetrate concrete barriers used in constructing the facility. Thus, it was assumed that intact engineered barriers would not deter drilling into waste at any time after loss of active institutional control, and a postdrilling scenario was assumed to be a credible occurrence at 100 y after disposal. In contrast, the disposal facilities at Savannah River are located on formations consisting mostly of sand and clay (Cook et al., 2002; McDowell-Boyer et al., 2000), and an assumption was made that the types of drilling equipment normally used at Savannah River could not readily penetrate an intact concrete barrier. Thus, it was assumed that a drilling scenario would not be a credible occurrence until engineered barriers would be expected to lose their integrity, at which time excavation into a facility while digging a foundation for a home also would be credible. Therefore, a postdrilling scenario was judged not to be relevant in determining acceptable disposals at Savannah River, because an agriculture scenario always results in higher estimates of dose per unit concentration of radionuclides in waste at the time intrusion occurs. As another example, the likelihood that an inadvertent intruder would excavate into waste while digging a foundation for a home, which is the initiating assumption in the construction and agriculture scenarios described in Sections 6.3.1.1 and 6.3.2.1, depends on the depth of burial of waste. An excavation for a foundation typically extends no more than 2 to 3 m below the ground surface (NRC, 1981a). Thus, at sites where erosion is not expected to be significant, burial of waste at depths considerably below 2 to 3 m could be assumed to preclude intrusion into waste by excavation for a very long time (Cook et al., 2002; McDowell-Boyer et al., 2000), even in the absence of intact engineered barriers to deter excavation. Under such conditions, it would not be reasonable to base decisions about acceptable disposals on an assessment of an agriculture scenario.

308 / 6. INADVERTENT HUMAN INTRUSION At some locations and for some facility designs, it is possible that there would be no credible scenarios for inadvertent intrusion into disposed waste. Suppose, for example, that a disposal facility consists of small-diameter boreholes in which waste is emplaced at a depth sufficient to preclude intrusion during normal excavation activities, and suppose further that drilling at the site is unlikely due, for example, to (1) an absence of mineral resources or useable groundwater supplies or (2) the considerable difficulty or expense in drilling for water at great depths. Under such conditions, it could be reasonable to assume that inadvertent intrusion into waste is unlikely, and widely used scenarios for inadvertent intrusion may not provide a reasonable basis for establishing site-specific limits on acceptable disposals of radionuclides. Disposal limits then would need to be based on other considerations, including that, as a matter of policy, some higher-activity wastes might be declared unacceptable for near-surface disposal at any site and using any technology. 6.4.3

Summary of Principles of Scenario Selection

Discussions in Sections 6.4.1 and 6.4.2 emphasize that selection of appropriate scenarios for inadvertent intrusion at any disposal site and for any facility design is largely a matter of judgment. Although the widely used intrusion scenarios discussed in Section 6.3 may be reasonable for many sites and facility designs, scenarios assumed in site-specific performance assessments should be tailored to particular site conditions and facility designs as they would influence actions that an inadvertent intruder reasonably might undertake. Furthermore, assumptions about exposure pathways and model parameter values should be credible for particular site conditions and facility designs. 6.5 Inputs to Dose Analyses for Inadvertent Intruders In scenarios involving direct intrusion, or attempts at direct intrusion, into a disposal facility, required inputs to dose assessments for inadvertent intruders are estimated concentrations of radionuclides in a facility at the time scenarios are assumed to occur. This Section discusses different factors that may be taken into account in estimating those concentrations on the basis of estimated concentrations in waste prior to disposal. These factors include the time after disposal when intrusion occurs, radioactive decay prior to intrusion, waste dilution following disposal, and radionuclide transport from a facility prior to intrusion.

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Time of Occurrence of Intrusion

The time between disposal of waste and an assumed occurrence of scenarios for inadvertent intrusion must be considered in any assessment, because concentrations of many radionuclides in waste may change significantly over that time (e.g., due to radioactive decay). The time period prior to an assumed occurrence of inadvertent intrusion includes the period after facility closure when active institutional control over a disposal site would be maintained. In current regulations for near-surface disposal of low-level waste, the period of active institutional control for purposes of assessments of inadvertent intrusion is assumed to be 100 y (DOE, 1988a; 1999b; NRC, 1982b).35 Assessments also should consider the time after loss of active institutional control when inadvertent intrusion would be precluded by intact engineered barriers in a disposal system, particular waste emplacements, or use of passive institutional controls. 6.5.2

Radioactive Decay

Radioactive decay prior to the time intrusion occurs should be taken into account in dose assessments for inadvertent intruders. 35Although a period of active institutional control of 100 y is assumed at near-surface disposal facilities for low-level waste for purposes of assessing potential impacts of waste disposal on inadvertent intruders, a much longer control period will be required at most sites if current NRC and EPA regulations that apply to remediation of contaminated sites remain in effect. NRC’s License Termination Rule (10 CFR Part 20, Subpart E; NRC, 1997) specifies that licensed sites and facilities could be released for unrestricted use following license termination if the annual effective dose equivalent to a member of the public who might reside on the site does not exceed 0.25 mSv. NRC’s dose criterion for unrestricted release of sites is much lower than dose criteria used in developing generally applicable limits on concentrations of radionuclides that would be acceptable for near-surface disposal at facilities licensed under 10 CFR Part 61 (NRC, 1982a) (Section 3.4.2.1). Similarly, EPA regulations in the National Contingency Plan (40 CFR Part 300; EPA, 1990) developed under CERCLA (Superfund) specify that the lifetime cancer risk to a member of the public at contaminated sites should not exceed ~10–4, and that criterion corresponds to annual doses much lower than those used by NRC and DOE in establishing limits on concentrations of radionuclides at near-surface disposal facilities (see Section 3.4.2.1 and 3.4.2.2 and NCRP, 2004). Thus, to the extent that low-level waste disposal facilities contain significant quantities of long-lived radionuclides, current regulations for release of contaminated sites essentially require perpetual institutional control over such facilities.

310 / 6. INADVERTENT HUMAN INTRUSION Given an assumption of a 100 y time period for active institutional control, as described above, reductions in concentrations in a disposal facility due to radioactive decay are particularly important for such shorter-lived radionuclides as 90Sr and 137Cs. Radioactive decay also can result in significant increases in radionuclide concentrations at future times and potential doses to inadvertent intruders. For example, if releases from a disposal facility over time are not taken into account (Section 6.5.4), potential doses from disposal of uranium in its natural isotopic abundance increase for more than 106 y, due to buildup of 226Ra and its short-lived decay products (Kocher, 1995; McDowell-Boyer et al., 2000; ORNL, 1997a). Similar increases in calculated doses over time occur for other long-lived radionuclides with radiologically significant long-lived decay products (e.g., 233U and 237Np). 6.5.3

Waste Dilution Following Disposal

An additional factor that normally should be considered in a dose assessment for inadvertent intruders is the presence of uncontaminated materials in a disposal facility (e.g., engineered barriers, materials used as backfill). The presence of uncontaminated materials in a facility reduces average concentrations of radionuclides to which an intruder would be exposed compared with concentrations in waste prior to disposal. The construction and agriculture scenarios or the discovery and resident, nonagriculture scenarios described in Sections 6.3.1 and 6.3.2 involve excavations into a disposal facility over areas of tens of m2 (NRC, 1981a). Thus, multiple waste packages or even multiple disposal units normally would be accessed during a single excavation. If it is assumed that an intruder would excavate into a disposal facility at random locations, which is a reasonable assumption in the absence of knowledge of the presence of a disposal facility, an appropriate volume to be assumed in converting radionuclide inventories in waste prior to disposal to average concentrations in a disposal facility to which an intruder would be exposed is the entire volume of a facility, including any uncontaminated regions in individual waste packages, between waste packages, and between disposal units (McDowell-Boyer et al., 2000; ORNL, 1997a; Wood et al., 1994). Thus, in exposure scenarios involving excavation or attempts at excavation into a disposal facility, average concentrations in waste prior to disposal could be reduced by a factor given by the fraction of the volume of a disposal facility that is occupied by waste after facility closure. If such a reduction is not taken into account,

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potential doses to inadvertent intruders following excavation or attempts at excavation into a facility clearly would be overestimated, because the presence of uncontaminated material in a disposal facility serves to reduce doses per unit concentration of radionuclides in disposed waste in such scenarios.36 The fraction of a disposal facility that is occupied by waste is termed a “waste dilution factor.” This factor depends on the design of a disposal facility, and the depth at which waste is buried also should be considered. Use of a waste dilution factor in developing waste acceptance criteria in the form of limits on concentrations of radionuclides is illustrated in Section 6.6.2. Even in drilling and postdrilling scenarios described in Sections 6.3.1.3 and 6.3.2.3, in which only a few waste packages normally would be accessed, concentrations of radionuclides in waste prior to disposal could be reduced by the fraction of the volume of individual disposal units that is occupied by waste, based again on the reasonable assumption that drilling would occur at random locations. The thickness of a cover on a disposal facility and the depth of uncontaminated soil between the facility and an underlying aquifer also could be taken into account, because concentrations of radionuclides in drilling wastes would be diluted by the presence of such uncontaminated materials. In these scenarios, however, it would not be appropriate to take into account uncontaminated regions between disposal units. Waste dilution factors should be consistent with the design of a disposal facility, locations of waste emplacement, and assumptions about intrusion scenarios. Depending on the facility design, reasonable approaches to considering the presence of uncontaminated material in a facility can reduce estimated doses to an inadvertent intruder by a factor of two or more when excavation into a facility is assumed (McDowell-Boyer et al., 2000; ORNL, 1997a). Waste dilution factors can be much higher when a drilling scenario is assumed and the depth of an aquifer is much greater than the thickness of emplaced waste in a disposal facility. 36A

reasonable alternative would be to assume that the probability of intrusion at locations of waste is less than unity and adjust estimated doses, assuming that intrusion occurs, accordingly. Such an approach was used by NRC in establishing limits on concentrations of radionuclides in Class-C waste (Section 3.4.2.1). NRC assumed that the probability of intrusion into small volumes of Class-C waste in a much larger disposal facility is 10 %, and concentration limits in Class-C waste were increased by a factor of 10 to account for the low probability of intrusion into such waste (NRC, 1982a).

312 / 6. INADVERTENT HUMAN INTRUSION 6.5.4

Consideration of Radionuclide Transport

Dose assessments for inadvertent intruders often assume that concentrations of radionuclides in disposed waste are not reduced by mobilization and transport in infiltrating water prior to the time intrusion into the waste could occur. That assumption provides conservative estimates of concentrations of radionuclides in a facility at future times and is appropriate, for example, in developing generally applicable disposal requirements, such as NRC’s waste classification system in 10 CFR Part 61 (NRC, 1982b). Neglect of mobilization and transport in water also may be reasonable in site-specific assessments of inadvertent intrusion when radionuclides are expected to be immobile or intrusion is assumed to occur within a few hundred years after disposal. However, that assumption may be quite conservative for mobile radionuclides and in cases, such as disposal of uranium discussed in Section 6.5.2, where a maximum calculated dose would occur at very long times after disposal, due to buildup of radiologically significant decay products. In order not to obtain overly conservative estimates of dose to inadvertent intruders in site-specific performance assessments and, thus, overly restrictive limits on radionuclide concentrations in waste, reductions in concentrations due to mobilization and transport from a disposal facility prior to the time intrusion is assumed to occur could be considered. However, as described below, care should be taken in estimating these reductions. If mobilization and transport of radionuclides from a disposal facility is taken into account in a dose assessment for inadvertent intruders, assumptions should tend to overestimate concentrations of radionuclides remaining in a disposal facility, in order not to underestimate potential doses. Since performance assessments often intend to model releases from disposal facilities conservatively (i.e., to overpredict actual releases), in order not to underpredict doses beyond the site boundary, lower estimates of releases probably should be used in estimating concentrations of radionuclides remaining in a disposal facility. Thus, the same source-term model normally should not be used in assessing off-site releases and exposures of inadvertent intruders. An approach to modeling the source term for use in intruder dose analyses is described in a performance assessment for a site on DOE’s Oak Ridge Reservation (ORNL, 1997a). An approach of using conservative assumptions to describe releases of radionuclides and retention of radionuclides in a disposal facility, as discussed above, is internally inconsistent (i.e., a

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correct mass balance between material released and material retained is not maintained in an analysis). However, this is also the case when depletion of radionuclides in a disposal facility due to mobilization and transport is not taken into account in a dose assessment for inadvertent intruders. The intent of using conservative assumptions in both cases is that neither releases to the environment nor doses to inadvertent intruders would be underestimated. 6.6 Outputs of Dose Analyses for Inadvertent Intruders This Section discusses outputs of dose assessments for inadvertent intruders and their use in developing waste acceptance criteria in the form of limits on concentrations of radionuclides in waste at the time of disposal. 6.6.1

Scenario Dose Conversion Factors

Outputs of models to assess dose to inadvertent intruders normally are given as doses per unit concentration of radionuclides, either annual doses per unit concentration for scenarios involving chronic exposure or total doses per unit concentration for scenarios involving a single, acute exposure. These outputs are referred to as scenario dose conversion factors. The following discussion considers scenario dose conversion factors for individual radionuclides and their use in developing concentration limits in waste at the time of disposal. Given that waste generally contains more than one radionuclide, a sum-of-fractions rule applies in determining acceptable concentrations. For mixtures of radionuclides, this rule states that the sum of the ratios of the concentration of each radionuclide to its concentration limit must not exceed unity. Radionuclide concentrations to which scenario dose conversion factors should be normalized are average concentrations in a disposal facility to which an intruder would be exposed, because intrusion is assumed to occur at random locations (Section 6.5.3). Then, if a performance criterion for inadvertent intruders expressed in terms of dose [e.g., effective dose (E)], denoted by Eintr, is specified, the limit on average concentration of a radionuclide in the disposal facility ( C F ) at the time when intrusion is assumed to occur is given by: E intr C F = ------------------ , SDCF

(6.1)

314 / 6. INADVERTENT HUMAN INTRUSION where: SDCF =

6.6.2

scenario dose conversion factor for particular radionuclide

Waste Acceptance Criteria Based on Intruder Dose Assessment

Equation 6.1 relates outputs of a dose assessment for inadvertent intruders (i.e., radionuclide-specific SDCFs) to limits on average concentrations of radionuclides in a disposal facility at the time intrusion is assumed to occur. For the purpose of establishing operating limits at a disposal facility, the desired quantities are the corresponding limits on average concentrations of radionuclides in waste prior to disposal. If fD(t) denotes a correction factor for radioactive decay between the time of disposal and the time intrusion is assumed to occur and G denotes a waste dilution factor described in Section 6.5.3, which takes into account the presence of uncontaminated material in a disposal facility, the limit on average concentration of a radionuclide in waste prior to disposal, denoted by C W , can be represented by: E intr CF C W = -------------------------- = ----------------------------------------------- . G × fD ( t ) SDCF × G × f D ( t )

(6.2)

For radionuclides without radiologically significant decay products, application of Equation 6.2 is straightforward, and the correction factor for radioactive decay is given by: fD ( t ) = e where: T1/2

=

– 0.693t -----------------T 1/2

,

(6.3)

half-life of radionuclide

However, when buildup of radiologically significant long-lived decay products occurs between the time of disposal and the assumed time of intrusion (e.g., in decay of uranium), the limit on average concentration of the parent radionuclide in waste at the time of disposal is obtained by replacing the term SDCF × fD(t) in Equation 6.2 by a sum of products of the scenario dose conversion factors and decay factors for all radiologically significant members of the decay chain, including the parent, given by:

∑ [ SDCFi × fD,i ( t ) ] , i = radionuclide index,

(6.4)

6.7 EFFECTS OF INADVERTENT INTRUSION

where: fD,i =

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activity of ith radionuclide at time t relative to initial activity of parent radionuclide (i = 1)

The decay factor for the ith radionuclide can be obtained from the Bateman equations for radioactive decay (Evans, 1955)

6.7 Effects of Inadvertent Intrusion on Off-Site Releases of Radionuclides A potentially important issue in performance assessment is whether the disruptive effects of excavation or drilling activities that are assumed to occur in scenarios for inadvertent human intrusion should be considered in estimating releases of radionuclides from a disposal facility to the environment over time (i.e., in assessing the normal performance of a disposal system). The actual occurrence of inadvertent human intrusion could compromise the integrity of engineered and natural barriers to release and transport of radionuclides and, thus, could result in significant increases in concentrations of radionuclides in the environment and doses to off-site individuals. In developing 10 CFR Part 61, NRC (1981a) considered possible effects of inadvertent human intrusion on the normal performance of a disposal facility by assuming that some caps on disposal trenches would be disturbed. However, disruptive effects of inadvertent intrusion on release and transport of radionuclides were not modeled separately from effects of natural processes, such as erosion or intrusion by plants and burrowing animals. An approach of considering only natural processes and events in estimating release and transport of radionuclides effectively decouples assessments of inadvertent intrusion from assessments of releases to the environment. This approach can be justified, in part, on the grounds that in assessing releases for the purpose of demonstrating compliance with performance objectives for members of the public, minimal credit usually will be taken for the ability of engineered barriers to reduce releases from a disposal facility and transport in the environment at times beyond a few hundred years, primarily because it is difficult to justify assumptions about barrier performance over longer times. Pessimistic assumptions about long-term performance of engineered barriers represent, at least crudely, possible effects of inadvertent intrusion on barrier performance over time.

316 / 6. INADVERTENT HUMAN INTRUSION The issue of whether effects of inadvertent human intrusion on release and transport of radionuclides need to be taken into account in assessing the performance of near-surface disposal facilities also was considered by DOE (Wood et al., 1994). A recommendation was developed that effects of human intrusion on barriers to release of radionuclides from disposal facilities need not be considered in assessing off-site releases. The rationale for this recommendation included the argument presented above. In addition, DOE noted that the concept of inadvertent intrusion is largely a hypothetical construct developed originally for the purpose of defining general categories of radioactive waste and later as the basis for establishing design and waste acceptance criteria at near-surface disposal facilities. Since inadvertent intrusion is a largely hypothetical occurrence, countless scenarios could be envisioned but these scenarios may never occur, particularly during the time that engineered barriers are effective. Thus, inadvertent intrusion is conceptually different from the kinds of natural processes that are expected to influence the performance of disposal facilities, the effects of which can reasonably be bounded. The approach taken by NRC and DOE of not explicitly considering effects of inadvertent human intrusion on the normal performance of near-surface disposal facilities for low-level waste differs from an approach specified in NRC and EPA regulations that apply to the Yucca Mountain Facility for disposal of spent nuclear fuel and high-level waste (EPA, 2001a; NRC, 2001), and a general approach recommended by ICRP (1998). As noted in Section 3.3.4, those regulations and recommendations call for assessments of effects of stylized scenarios for inadvertent intrusion on the normal performance of a disposal facility. At the Yucca Mountain Facility in particular, NRC and EPA regulations require an assessment of doses to off-site members of the public that would result from drilling through a waste package, and the specified performance objective for such releases is the same as the performance objective that applies to undisturbed performance (i.e., absent human intrusion). Consideration of drilling intrusion also was required in demonstrating compliance with EPA regulations that apply to disposal of DOE’s defense transuranic waste at the WIPP Facility in New Mexico (EPA, 1993a; 1996). In that case, however, projected consequences of drilling on releases to the accessible environment took into account estimated probabilities of deep and shallow drilling at the site and an assumption of drilling at random locations. NRC and DOE have not yet reconsidered their position that effects of inadvertent human intrusion on the normal performance of near-surface disposal facilities for low-level waste need not be

6.8 SUMMARY

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considered in demonstrating compliance with performance objectives in light of requirements in more recent EPA and NRC regulations for disposal of high-level and transuranic wastes in geologic repositories and ICRP recommendations. 6.8 Summary Section 6 has discussed the development and implementation of scenarios for inadvertent human intrusion at near-surface disposal facilities for low-level waste. Site-specific analyses of inadvertent intrusion are required in performance assessments at DOE disposal sites (DOE, 1988a; 1999b), and they may be required at sites for disposal of non-DOE low-level waste. These discussions have emphasized conceptual approaches that could be used in developing and implementing exposure scenarios and the many subjective judgments involved. The concept of inadvertent human intrusion as applied to nearsurface disposal facilities has seemingly contradictory aspects. On the one hand, commonly assumed scenarios involving permanent residence on a disposal site should be unlikely occurrences at disposal sites for the foreseeable future, because some form of active or passive institutional control or societal memory of disposal activities presumably will be maintained for a long time. On the other hand, waste acceptance criteria in the form of limits on concentrations of radionuclides usually are based on an assumption that scenarios for inadvertent intrusion involving permanent residence on a disposal site will occur immediately after loss of active institutional control, and that such control will be relinquished within a time after facility closure that is short relative to half-lives of many important radionuclides in low-level waste. This seeming contradiction can be reconciled by recognizing that assessments of inadvertent intrusion are not intended to provide estimates of dose to individuals who might actually intrude onto disposal sites at some future time. Rather, the purpose is to ensure protection of inadvertent intruders by developing facility design and waste acceptance criteria on the basis of pessimistic assumptions about future human activities at disposal sites. An intention to maintain institutional control over disposal sites for the indefinite future is an important factor in limiting the types of intrusion scenarios that might actually occur. In effect, assumptions about inadvertent intrusion invoked in performance assessment provide a form of defense-in-depth in protecting members of the public should active and passive institutional control over disposal sites and societal memory of past disposals no longer be maintained at some time in the future.

318 / 6. INADVERTENT HUMAN INTRUSION Given these considerations, recommended approaches to considering inadvertent intrusion in performance assessments of nearsurface disposal facilities for low-level waste may be summarized as follows: • Widely used scenarios for inadvertent intrusion that were developed by NRC and DOE, particularly the agriculture and postdrilling scenarios involving permanent occupancy of a disposal site by a resident homesteader, should, in a general way, be appropriate at most sites and for most facility designs. Furthermore, for any site and facility design, those scenarios should provide reasonably conservative estimates of dose to individuals who might actually intrude onto a disposal site. It usually is not reasonable to develop worst-case scenarios that would give significantly higher estimates of dose but would have only a small probability of occurrence at any site, unless such assumptions would not have a significant effect on the ability of a site to accept waste for disposal. • The previous recommendation notwithstanding, it usually is not appropriate in conducting site-specific performance assessments to apply widely used exposure scenarios for inadvertent intruders in the same way at all sites and for all facility designs. Rather, assumed exposure scenarios, exposure pathways, and model parameter values should be reasonably consistent with site conditions, facility design, and local environmental conditions as they would affect credible activities by an inadvertent intruder. It also is reasonable to assume that future inadvertent intruders would be at least as technically knowledgeable as the current generation and would take appropriate precautionary actions if recognizable engineered barriers or wastes were encountered. Assumptions about inadvertent intrusion should focus on conservative but credible occurrences for the particular site and facility design. Furthermore, substantial modifications of widely used exposure scenarios could be appropriate for unusual disposal sites or facility designs. • In implementing scenarios for inadvertent intrusion, it should be assumed that disposal facilities would be accessed at random locations because, by definition, an inadvertent intruder has no prior knowledge of disposals at a site. Thus, dose assessments for inadvertent intruders can be based on average concentrations of radionuclides in a disposal facility, rather than maximum concentrations at any location.

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Estimates of such average concentrations can take into account the dilution of radionuclides in disposed waste due to the presence of uncontaminated materials in a disposal facility. Assumed waste dilution factors should be consistent with the design of a facility, any selective emplacements of waste, and assumptions regarding intrusion into the waste (e.g., by excavation or drilling). • In site-specific dose assessments for inadvertent intruders, reductions in radionuclide concentrations in disposal facilities over time due to mobilization and transport of radionuclides in infiltrating water could be considered. However, assumptions used in modeling radionuclide removal from a disposal facility should be expected to underestimate actual releases, in order not to underestimate potential doses to inadvertent intruders. Thus, the model of the source term used to estimate off-site releases of radionuclides in a performance assessment would not be appropriate for use in a dose assessment for inadvertent intruders if that model is intended to overestimate actual releases. • At the present time, NRC and DOE regulations and guidance do not specify that effects of inadvertent human intrusion on the ability of natural and engineered barriers to inhibit release and transport of radionuclides from a nearsurface disposal facility need to be considered explicitly in performance assessment. However, if suitably conservative models are used to describe degradation of engineered and natural barriers over time, possible effects of inadvertent human intrusion may be represented implicitly.

7. Uncertainty, Sensitivity and Importance Analysis 7.1 Introduction Uncertainty and sensitivity are important attributes of any model. Uncertainty analysis considers the lack of sureness or confidence in predictions of a model due, for example, to uncertainties in model parameters or the structure of the model itself, and sensitivity analysis considers the responsiveness of model predictions to selected perturbations of the model’s parameters. This Section discusses the role of uncertainty and sensitivity analysis in performance assessments of low-level waste disposal facilities and approaches to uncertainty and sensitivity analysis that should be suitable for use in performance assessment. Methods of uncertainty and sensitivity analysis have been studied extensively, and a large body of scientific literature is devoted to each. However, that literature and the experience developed from it must be used with caution in the context of performance assessment, due to the unusual nature of this type of assessment and the significant amount of judgment that must be exercised. Results of uncertainty and sensitivity analyses of performance assessments need to be interpreted carefully, lest an erroneous and unwarranted implication of precision in calculations be imputed. To address these concerns, this Report uses the term “importance analysis” to describe the types of uncertainty and sensitivity analysis and their interpretations that are appropriate for use in regulatory decision making. Section 7 is organized as follows. Section 7.2 provides a definition of importance analysis and a description of the rationale for introducing this term. Section 7.3 discusses the purpose of importance analysis and the context in which importance analysis is carried out. Section 7.4 discusses the nature of uncertainties in performance assessment and how they are fundamentally different from uncertainties encountered in many other fields. Section 7.5 discusses different mathematical approaches to treating uncertainty in performance assessment models, merits of each approach, and how results of uncertainty analysis are used in importance analysis. These discussions emphasize the need to use judgment in 320

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an uncertainty analysis, and they reflect the needs of performance assessment as represented in previous sections of this Report. Section 7.6 discusses the role of sensitivity analysis in importance analysis. Section 7.7 discusses the use of importance analysis in performance assessment. Finally, Section 7.8 summarizes recommendations on importance analysis. Some discussions of how importance analysis is used in the context of performance assessment also are contained in Section 4, and discussions on the concept of “reasonable assurance” in Section 3.5.3 are relevant to importance analysis.

7.2 Description of Importance Analysis Uncertainty is an unavoidable aspect of decisions about the behavior of complex waste disposal systems over long periods of time. There is no question that uncertainties exist, and that a decision maker must account for them. However, there are important questions about the nature of these uncertainties, and about how to treat them in a way that leads to reasonable decisions regarding regulatory compliance. The principal challenge of the performance assessment process is to translate results of uncertain calculations to the needs of a decision maker, who must ultimately use available information to make a “yes” or “no” decision about the acceptability of a waste disposal system. The process of performance assessment and its use in regulatory decision making must acknowledge the nature of uncertainties in performance assessment and the essential role of judgment in any assessment. When considering uncertainty in performance assessment, the unusual nature of this type of analysis must be recognized. As emphasized throughout this Report, the primary purpose of performance assessment is to provide essential input to a decision about compliance of a disposal facility with applicable regulatory requirements. Consequently, analysts need not be unduly concerned with representing uncertainty in the outcome (e.g., a prediction of the dose that might be received by some individual in the future), since the “correct” answer is both unknown and unknowable. Rather, analysts should be concerned primarily with identifying important assumptions and parameter values which, when changed within credible bounds, can affect a decision about regulatory compliance. This distinction between numerical uncertainty and an identification of factors of importance to a regulatory decision is vital to understanding the need for importance analysis and suitable approaches to such analyses.

322 / 7. UNCERTAINTY, SENSITIVITY AND IMPORTANCE ANALYSIS In an attempt to clarify these issues, the term “importance analysis” as used in this Report is defined as an integration and interpretation of results obtained from the performance assessment process for the purpose of identifying assumptions and parameter values which, when changed within credible bounds, can affect a decision about regulatory compliance; see also NCRP (1996c).37 The issue of concern here is more than a matter of semantics. Uncertainty and sensitivity analysis, in the context often understood in other fields, is not necessary in performance assessment. More traditional analyses are intended to represent the variability of possible results around an expected outcome and effects of uncertainty in different input parameters on the variability in projected outcome (i.e., traditional analyses are directed toward a representation of uncertainty in the actual behavior of a system). Furthermore, traditional analyses are meaningful only when a large amount of empirical data on performance is available. However, such is not the case in regard to long-term projections of the performance of waste disposal systems, which generally are based on limited data. The distinction between importance analysis and traditional uncertainty and sensitivity analyses reflects the reality that, at the present time, our understanding of processes affecting the long-term performance of waste disposal systems and the availability of relevant data are insufficient to support rigorous modeling of all aspects of a disposal system. In the presence of limited information, judgment regarding the degree of conservatism in assumptions and parameter values used in performance assessment will be a critical factor in defending a decision about regulatory compliance. 7.3 Purpose of Importance Analysis Performance assessments of waste disposal facilities are conducted to provide a technical basis for decisions about regulatory 37The

term “importance analysis” as used in this Report differs somewhat from the concept of “importance measures” that has been used in probabilistic risk and reliability analyses of engineered systems and applied in performance assessments of geologic repositories for disposal of high-level wastes (e.g., Eisenberg and Sagar, 2000). The latter type of analysis focuses more on identifying aspects of a disposal system that are important in determining projected outcomes and less on identifying assumptions about system performance that are important to a regulatory decision.

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compliance. A decision maker needs to know two things about the results of an assessment: (1) the magnitude of projected outcomes (e.g., maximum annual dose), which are to be compared with applicable performance objectives; and (2) the credibility of projected outcomes with respect to supporting a regulatory decision. Given the limited amount of data on long-term performance of waste disposal systems, projected outcomes are highly uncertain by any standard statistical measure. Consequently, the key issue is to identify conditions of uncertainty in assumptions or parameters that can affect a regulatory decision. If a decision can be shown to be insensitive to judgments about uncertain conceptual models, model formulations, and parameter values, a decision can be made and defended with confidence. Thus, while a regulatory decision may be based on a single projection of outcomes, the justification for a decision must be based on an importance analysis. An identification of conditions that can affect a decision is the primary concern of importance analysis, rather than a quantification of uncertainties in model calculations. Consequently, an importance analysis must be structured in a way that maximizes confidence in a regulatory decision in the presence of limited data. The goal of importance analysis must be to consider the range of credible conceptual models, mathematical models, and combinations of parameter values and determine if credible combinations could lead to a decision of noncompliance. 7.4 Nature of Uncertainties in Performance Assessment 7.4.1

Characteristics of Uncertainties

A careful distinction must be made between the fundamental characteristics of uncertainty and mathematical approaches to representing uncertainty. Uncertainties in performance assessment are a fact of life, and they must be addressed in the process of establishing the credibility of an assessment. A number of mathematical techniques can be used to represent uncertainty, and their relative merits are a subject of debate, but those techniques must not be confused with the fundamental issue that uncertainties exist and must be represented appropriately for reasonable decision making to take place. As emphasized in Section 2, models used in performance assessment should be commensurate with available data, and the same is true of uncertainty analysis. Just as in an example of using a three-dimensional model when there are insufficient data to characterize a hydrologic system, use of rigorous techniques

324 / 7. UNCERTAINTY, SENSITIVITY AND IMPORTANCE ANALYSIS of uncertainty analysis when there are only a few data points does not necessarily yield a better understanding of uncertainties in an assessment. An important characteristic of uncertainty in performance assessment is that it is not a function of a disposal site but, rather, is a function of our knowledge about the site. The site itself is not uncertain; it is determined by the existing geologic structure and its evolution in time. In addition, when a disposal facility is imposed upon a site, there will be a single, definitive behavior of the disposal system in the future, which will result in a single, definitive outcome (e.g., maximum annual dose to any individual). Uncertainty lies in our inability to perceive the current state of the site and to know the future behavior of the system. 7.4.1.1 Type-A and Type-B Uncertainties. Uncertainties of importance to performance assessment may be due to stochastic variation or lack of knowledge (IAEA, 1989b; 1999; Kaplan and Garrick, 1981). Stochastic variation refers to the variability attributed to a property of a system based on repeated measurements. Examples include a time series of rainfall measurements or repeated measurements of the sorption capacity of a soil or engineered material. Uncertainty due to stochastic variation is referred to as “Type A” (IAEA, 1989b) or “aleatory.” In contrast, uncertainty due to lack of knowledge is founded on an incomplete characterization, understanding or measurement of a system. Examples include the projection of disposal system behavior into an uncertain and immeasurable future, incomplete or inaccurate information on radionuclide inventories, and incomplete understanding of processes governing release and transport of radionuclides. Intrinsic spatial variability also is included in this category, since its uncertainty results from an incomplete characterization of a material or process below some spatial scale of concern. Uncertainty due to lack of knowledge is referred to as “Type B” (IAEA, 1989b) or “epistemic.” Type-B uncertainties generally are the most important in performance assessments of waste disposal facilities, because measurements made to support assessments tend to be limited. Furthermore, measurements tend to be based on interpretations of the behavior of a system and to be focused on what can be measured rather than what should be measured ideally. For example, even though rainfall data may appear to be stochastically distributed (Type A), measurements apply only to current or historical conditions. Once the future evolution of a site is considered, the stochastic nature of the data becomes less relevant than a Type-B

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uncertainty associated with projections into the future. Similarly, repeated measurements of sorption in a particular soil provide information about the current Type-A uncertainty, but there is an even larger component of Type-B uncertainty associated with use of the data in a performance assessment. This uncertainty is embodied, for example, in concerns about the following factors: (1) whether the soil is representative of site conditions and, if so, over what time and spatial scale; (2) the influence of spatial variability on use of measured values; (3) whether the geochemistry of the soil will evolve over time and, if so, how; (4) whether laboratory measurements apply to field conditions; and (5) whether the sorption model is appropriate for an analysis. 7.4.1.2 Classification of Model Uncertainties. Uncertainties in performance assessment have been further classified in terms of model uncertainty, including uncertainty in conceptual models and mathematical models chosen to represent them, uncertainty about the future of a site, and parameter uncertainty (Davis et al., 1990). This classification is not fundamental, because an uncertainty that falls in a particular class necessarily contains elements of the other two. For example, an estimate of uncertainty in a measured parameter that may be used in performance assessment (e.g., resulting from spatial variability) requires the use of a particular model for interpretation and implementation of the data and particular assumptions about the applicability of the data in an uncertain future. Thus, even though a parameter may be measured (with some uncertainty), there will always be additional uncertainty about its applicability to performance assessment. Nevertheless, this classification of uncertainty can be useful in interpreting results of performance assessments. For example, the question of whether an advective-dispersion equation is appropriate in modeling radionuclide transport at a site is usefully classified as a model uncertainty, whereas the specific value of the dispersivity contained in the model is more usefully regarded as a parameter uncertainty. At a fundamental level, however, uncertainty in use of a model equation and uncertainty in a parameter value to be used in the model are inseparable. A classification of uncertainties in performance assessment into uncertainties about models, uncertainties about the future, and uncertainties in model parameters is useful, even though such a classification is not fundamental. Given this viewpoint, uncertainty in each aspect of a performance assessment can be classified and an appropriate treatment of those uncertainties applied. The following sections discuss the three classes of uncertainty in more detail.

326 / 7. UNCERTAINTY, SENSITIVITY AND IMPORTANCE ANALYSIS 7.4.2

Uncertainty in Models

The process of developing a site-specific model begins with a perception (i.e., set of assumptions) about current conditions, which is developed on the basis of site-specific data (Davis and Olague, 1991). A perception of reality can never be complete, and it often is not unique, particularly in modeling groundwater systems that cannot be observed directly. An incomplete perception, both now and in the future, is one of the primary sources of uncertainty in modeling. Next, simplifying assumptions are made to develop a conceptual model, which is a qualitative description of the processes, geometry and boundary conditions associated with a system. Finally, those qualitative ideas are translated into a quantitative mathematical model, which is a set of equations that represent the conceptual model. The mathematical model can be solved, with site-specific input parameters, to obtain the output of interest (e.g., annual dose). Model uncertainty encompasses uncertainty in the conceptualization of a disposal system, uncertainty in its mathematical representation, and uncertainty in the solution of the mathematical representation (Bonano and Cranwell, 1988). Uncertainty in a conceptual model arises from a number of sources. There may be uncertainty associated with an initial perception of current conditions arising, for example, from misinterpretations of site-specific data or inadequacies in data reduction techniques, or available data may be consistent with more than one interpretation. Uncertainty will be introduced by simplifying assumptions that are necessary to make the problem tractable mathematically. For example, natural systems are dynamic but are usually modeled as steady-state. In addition, models are most often developed by a single analyst or a small group of analysts using professional judgment to translate available data into a model. The model is therefore limited by the abilities and imagination of developers and reviewers, in addition to limitations in data. Also, despite the usefulness of experience with sites other than the one being considered, a poor conceptual model can be developed for an apparently similar site on the basis of inappropriate preconceived ideas (Bonano and Cranwell, 1988). In many cases, uncertainty in a conceptual model is the dominant type of uncertainty in a performance assessment. If an inadequate conceptual model is assumed, uncertainties associated with a mathematical model and model input parameters become irrelevant. In some cases, even processes that control system behavior are not well understood. For example, a proper conceptual model

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for unsaturated flow in fractured porous media has not been identified. There can be simultaneous flow in the porous matrix and fractures, but conditions that cause one process or the other to dominate flow are not understood sufficiently to permit a formulation of equations to describe exchange of material between fractures and the matrix, except in a heuristic manner that has not been substantiated (or refuted) by experiment (Updegraff et al., 1991). In addition to uncertainties associated with an underlying conceptual model, uncertainty in mathematical modeling arises from methods required to obtain a solution of equations of interest (Davis and Olague, 1991) or from an inability to represent a conceptual model in a suitable mathematical form (Bonano and Cranwell, 1988). If an analytical solution technique is used, uncertainty can be introduced if the solution is incorrect or if truncation of an infinite series, such as an error function, is used as part of a solution. Furthermore, as simplifications are introduced, a mismatch of boundary conditions commonly occurs, either between segments of an analysis or between physically reasonable conditions and conditions for which a solution can be obtained. Numerical solution methods will almost always be implemented by means of a computer program, and use of computer programs can introduce uncertainty due to errors in numerical approximations of equations and errors in coding. An additional source of uncertainty in applying computer codes that should not be overlooked is user error. 7.4.3

Uncertainty in Future Site Conditions

Uncertainty about the future of a disposal site is a direct result of our inherent lack of knowledge about how a site will evolve in time. Climatic and geologic conditions, as well as human living habits, that will occur in the future are not known. Consequently, a method of accommodating alternative future conditions must be considered. Treatment of uncertainty in future conditions is primarily a philosophical and ethical question, rather than a technical issue, and should be resolved by regulators. Since many aspects of the future cannot be predicted, regulators should define those future conditions to be included in a performance assessment and how they should be taken into account. For example, standards for disposal of spent nuclear fuel and high-level waste at the Yucca Mountain Site specify exposure conditions for future individuals residing near the site to be assumed in a performance assessment that are likely to be conservative with respect to actual human exposures at future times (EPA, 2001a). That approach eliminates

328 / 7. UNCERTAINTY, SENSITIVITY AND IMPORTANCE ANALYSIS the need to consider alternative exposure conditions, and it allows a performance assessment to focus on the important question of the waste isolation capabilities of the disposal system. 7.4.4

Uncertainty in Model Parameters

Parameter uncertainty reflects an incomplete knowledge of constitutive coefficients in a mathematical model. This type of uncertainty is identified, in part, with uncertainty in actual values and in statistical and spatial distributions of data used to infer values of model parameters. Sources of parameter uncertainty include measurement error, insufficient data, and inconsistency between the spatial scale of measurement and the scale of a model. Different sources of parameter uncertainty often are discussed as though they are independent of model uncertainty. However, model uncertainty and parameter uncertainty often are not independent. For example, uncertainty in use of an equilibrium sorption model that is parameterized in terms of a distribution coefficient (Kd) frequently is treated as a parameter uncertainty by focusing on values of Kd or their probability distributions. However, use of this model requires numerous other assumptions that are implicit in the model, including assumptions of equilibrium, reversibility of sorption, linearity of the sorption isotherm, and lack of spatial variability. Furthermore, selection of an equilibrium sorption model tends to reinforce the way the model parameter is measured and interpreted. Sorption measurements tend to be expressed in terms of Kd not because that is necessarily the best way to express them, but because that is how they will be used. This example shows some of the ways that model uncertainties can be linked with parameter uncertainties. 7.4.4.1 Measurement Errors. Measurement errors generally are considered to be either random or systematic. In the context of uncertainties in performance assessment, random errors in parameter measurements are relatively unimportant, and are not discussed further here. However, systematic errors are a greater concern and are unavoidable in many important measurements that are needed to adequately characterize a site. They can result, for example, from use of laboratory data or data obtained at another site. Many hydrologic measurements must be made using disturbed samples, and in situ measurements of most parameters are not possible. For example, the only available approach to fully evaluating moisture characteristic curves in the unsaturated zone

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(Section 5.5.3) involves laboratory measurements on core samples. A core sample will never have the same characteristics as undisturbed soil in the field, and there will always be systematic, and largely unknown, errors associated with laboratory measurements. The issue then becomes whether those errors are significant. Other common hydrologic parameters (e.g., porosity, degree of anisotropy, dispersivity) also are measured primarily in the laboratory and suffer from the same potential for systematic error. This source of error generally has been ignored in the literature, largely because it is not easily evaluated. Independent methods of deriving the same information about a site for comparison with data obtained in laboratory tests usually are not available. When nonintrusive, field-scale geophysical measurements can be made, data often are difficult to interpret. 7.4.4.2 Insufficient Data. Site characterization requires that sufficient data be collected to define, to the extent necessary for the purpose of performance assessment, the natural hydrogeologic environment at a site and the long-term behavior of natural and engineered systems under current and future climatic conditions. This task is difficult even for engineered systems, such as concrete vaults, which are relatively homogeneous and evolve in time in a relatively well understood manner. However, there generally is a large inherent spatial and temporal variability associated with natural systems that must be accounted for in data collection, data interpretation, and modeling. One of the crucial tasks in site characterization is to develop confidence that spatial variability does not include a preferential path of release and transport of radionuclides that would result in noncompliance of a disposal system with applicable performance objectives. Insufficient data can prevent the development of adequate performance assessment models. However, the amount of data that is sufficient to build an “adequate” model generally cannot be identified, or even defined unambiguously at a specific site. Professional judgment is necessary in evaluating the sufficiency of data. Problems of unavailability of data or difficulties in determining the sufficiency of available data can be circumvented, to some extent, by using conservative parameter values in lieu of detailed site characterization data. All regulatory decisions will need to be made under conditions of uncertainty, and data sufficiency is an important aspect of confidence in a decision. Consequently, conservative parameter values can be used to reduce uncertainty in a regulatory decision even if they do not promote a fundamental understanding of a disposal system.

330 / 7. UNCERTAINTY, SENSITIVITY AND IMPORTANCE ANALYSIS 7.4.4.3 Dependence of Measurements on Scale. One of the most important sources of parameter uncertainty is an inconsistency between the spatial scale of measurements of a parameter and the desired scale of model simulations. This inconsistency particularly arises when a field-scale model parameter cannot be measured directly but must be inferred on the basis of laboratory or relatively small field-scale tests and application of an assumed scaling rule. The dependence of measurements on spatial scale often is an important concern in modeling flow and transport of radionuclides in the environment, but it also can affect the ability to model engineered barriers in disposal systems. For example, of considerable importance to performance assessment is the dependence of dispersion in radionuclide transport on the spatial scale and the uncertainty that results from attempts to identify an appropriate dispersivity for use in an assessment. Dispersion is one of the primary factors that can be used to reduce projected concentrations of radionuclides in groundwater, but it is difficult to quantify owing to its considerable dependence on scale. Laboratory-scale dispersivities tend to be lower than field-scale values by orders of magnitude, but there is no adequate way to quantify field-scale dispersivities without conducting field tests. Furthermore, dispersivities at field scale are not measured but are inferred from a combination of measurements, modeling and interpretation, and a value determined from any such combination depends greatly on the technique used. As another example, the highly nonlinear nature of vadose-zone flow models coupled with the extreme and complex spatial variability that characterizes most vadose-zone flow problems makes an identification of suitable methods of averaging local measurements to obtain field-scale parameters a formidable obstacle. 7.5 Mathematical Methods of Treating Uncertainty This Section describes mathematical methods that have been developed to represent uncertainties in inputs to performance assessments and to propagate those uncertainties through an analysis to obtain an estimate of their effects on model output. These methods must be viewed from the perspective of performance assessment, particularly the need for an importance analysis rather than a conventional uncertainty analysis. The role of importance analysis is to identify those assumptions and parameter values that cause performance objectives to be exceeded in an analysis (Section 7.3). The challenge for performance assessment is to use available methods of considering uncertainty to accomplish

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this task when only limited data are available and many important uncertainties are not quantifiable except in highly subjective ways (IAEA, 1999). 7.5.1

Introduction to Mathematical Methods

Methods of uncertainty analysis used in performance assessment can be categorized as deterministic, probabilistic, or a combination of the two. Deterministic methods are those in which a single scenario, model and set of input parameters are used to calculate a single value of the model output. Uncertainty is treated by conducting numerous deterministic analyses using different assumptions, including different scenarios, and choosing one of them to be the basis for a decision about the acceptability of a disposal facility, after considering any conditions that may cause performance objectives to be exceeded. In probabilistic methods, probabilities are assigned to scenarios and probability density functions are assigned to parameters. The idea of assigning probabilities to models also has been considered. Probabilistic analyses consist of a series of deterministic analyses, and relationships between different deterministic results are expressed using probability theory. Early advocates of probabilistic approaches to performance assessment viewed probability distributions of input parameters as being based on repeated measurements. That view represents the frequentist theory of probability. More recently, it has been recognized that all probability distributions used in performance assessment involve subjective elements, due to the limited availability of data. Consequently, Bayesian and other judgmental probability theories have increasingly been emphasized, although those approaches have important flaws. Klir and Folger (1988) discuss uncertainty in a broad context. Uncertainties are divided into numerous categories and, starting from this context, uncertainties that arise in performance assessment can be shown to be of particular types that have particular approaches to describing and propagating uncertainty. Of particular importance to performance assessment are so-called fuzzy measures and possibility measures (Dubois and Prade, 1988), since those approaches were developed to represent subjective and informational (Type-B) uncertainties. Recent work suggests that those approaches may eliminate some problems associated with Bayesian probabilities (Robinson and Grindrod, 1994), although there are differing opinions about the relative strengths of competing approaches to treating Type-B uncertainties (Dubois and

332 / 7. UNCERTAINTY, SENSITIVITY AND IMPORTANCE ANALYSIS Prade, 1992; Wu et al., 1990). Nonprobabilistic approaches are receiving attention in the literature (Cooper, 1994; Robinson and Grindrod, 1994), and they appear to merit further consideration. Regardless of the approach used, the result of an uncertainty analysis is essentially the same. If all uncertainties are addressed, the result would be a set of alternative possible outcomes (e.g., annual doses) that depend on a particular conceptualization of a disposal system and its evolution in time, mathematical models used to represent that conceptualization, and parameters used in the mathematical models. The primary difference between the various approaches is the manner in which results of individual deterministic analyses are integrated and interpreted for input to the decision-making process. Each type of uncertainty discussed in Section 7.4 (i.e., model uncertainty, uncertainty in the future, and parameter uncertainty) is qualitatively different. Consequently, propagation of each type of uncertainty is addressed separately in the following sections. Further discussion on combining and using results of each kind of analysis is given in Section 7.7. 7.5.2

Propagation of Model Uncertainty

Most performance assessments have been conducted using a single conceptual model of a disposal system that represents an analyst’s best judgment about its behavior, and little work has been done to address uncertainty in conceptual models (Kozak et al., 1991). However, regulatory agencies may choose to conduct independent performance assessments, rather than simply review an applicant’s analysis. Such an independent check is beneficial in building confidence that credible assumptions that could affect a decision about regulatory compliance, including different assumptions about conceptual models, have been identified and addressed. Independent assessments generally should result in different estimates of dose, but such differences are unimportant for purposes of regulatory decision making. Regardless of projected outcomes, if independent assessments based on different credible assumptions support a decision of compliance, additional confidence in a regulatory decision is provided. The only available approach to accounting for uncertainties in conceptual models is to conduct assessments that span a range of conceptual models that are reasonably consistent with site-specific data. Some investigators have suggested that formal elicitation of expert opinion may be a good way of spanning that range and developing an extensive list of possible alternative conceptual models that are consistent with available data (Chhibber et al., 1991a; Kerl

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et al., 1991). By broadening the base of expertise from which conceptual models are developed to include additional analysts, regulators, reviewers and public interest groups, there is an increased likelihood that uncertainties in conceptual models will be addressed to the extent possible and that conceptual models will be included that capture potentially adverse characteristics of a disposal system. Disadvantages of using highly structured formal elicitation processes include increases in cost and time and a reduction in flexibility (Bonano et al., 1990). In addition, such approaches involve consideration of increasing layers of judgment, which may make defense of a regulatory decision less transparent and, thus, more difficult. Chhibber et al. (1991a; 1991b) and Heger et al. (1991) have suggested that once alternative conceptual models have been elicited, a probability should be associated with each model. That probability is interpreted as a measure of the degree of belief that a model is appropriate for the given purpose. Chhibber et al. (1991a) also recognized a number of difficulties with that method. Perhaps the most important is that to apply probability theory, the different conceptual models should be defined such that they are exhaustive, mutually exclusive, and independent. This difficulty seems insurmountable since all conceptual models would be based on the same site-specific data. Other difficulties arise in combining and aggregating expert opinion, and in incorporating new information into probability estimates. Thus, although this approach represents an interesting area of research, many significant issues need to be addressed before it can be considered for use in performance assessment. Kozak et al. (1993) suggested an alternative approach to treating uncertainty in conceptual models in which all models consistent with available site-specific data are used in a performance assessment. If a particular model cannot be shown to be unlikely by acquiring additional data, it should be considered credible. A performance assessment is conducted using each credible conceptual model, and results are used to establish the model that is the most conservative. This approach avoids the problem of assigning probabilities to conceptual models, and it emphasizes choosing a model that permits the most comfortable regulatory decision. Kozak et al. (1993) also emphasize that model conservatism should be determined after an analysis, since there is no general way of establishing the relative degree of conservatism of different conceptual models prior to performing calculations. This approach can be used with either deterministic or probabilistic methods of uncertainty analysis.

334 / 7. UNCERTAINTY, SENSITIVITY AND IMPORTANCE ANALYSIS The degree of conservatism in a model is always relative to something. Ideally, when a conceptual model is intended to be conservative, it should be conservative with respect to the actual behavior of a disposal system. Unfortunately, an analyst rarely will have the luxury of establishing conservatism with respect to any single aspect of system behavior, let alone conservatism with respect to overall system performance. Thus, model conservatism usually is defined with respect to other possible models or combinations of parameters. In that sense, model intercomparisons play an important role in evaluating the conservatism of a model and, thus, its reliability in regulatory decision making. In performance assessment, uncertainty in mathematical models developed to represent assumed conceptual models usually is assumed to be negligible compared with other uncertainties, especially uncertainties in parameter values, and is not incorporated in model results (Davis and Olague, 1991). Careful quality assurance procedures should be followed during development of computer codes to avoid introducing coding errors. Quality assurance of computer codes also should include verification exercises to ensure that model equations are implemented correctly. Verification is accomplished by careful evaluation of a code and by comparing its output with analytical solutions or outputs of other verified codes that solve the same problem (i.e., by benchmarking). Once a computer program is in operation, a configuration management system should be followed to ensure that no haphazard modifications are made. Error reduction techniques should be followed, such as using a finer discretization of the domain or using more terms to represent an infinite series. The user will know that errors have been reduced to acceptable levels once a convergent solution is achieved (e.g., when a finer discretization does not change the solution). 7.5.3

Propagation of Future Uncertainty

A generalized approach to addressing uncertainty in future conditions at a disposal site is to span the range of credible high-risk situations. Identifying all important events and processes that might occur in the future is a creative task that may depend exclusively on expert judgment (Bonano et al., 1990). Such an identification can be accomplished most readily by formalized elicitation of expert opinion in a manner analogous to an elicitation of expert opinion about conceptual models. A significant amount of elicitation regarding scenarios of the future evolution of a disposal system has already been undertaken, and comprehensive lists of possible disruptive events and processes are available (Cranwell et al., 1990).

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Most performance assessments of low-level waste disposal facilities have assumed continuity of present-day conditions into the indefinite future. On the basis of data obtained over times ranging from a few years to tens of years, an analyst chooses reasonable values or ranges of climatic data, earthquake frequency and intensity, and other potentially important time-dependent phenomena, and conducts a performance assessment that includes those conditions. This approach has an advantage of simplicity, and deviations from assumed conditions probably will not be substantial for the first 100 to 200 y after closure of a facility. In addition, it is often argued that performance assessment is only a hypothetical projection of doses under assumed conditions, in which case an analyst should not be too speculative about future conditions at a site. This argument is reasonable in regard to human activities at a site, provided prudently conservative assumptions developed on the basis of current conditions are used, since any assumptions about future human behavior are largely speculative. The primary difficulty with an approach of choosing reasonable values or ranges of future conditions on the basis of historical data is an ironic one. In the past, disposal of low-level waste in unlined trenches was commonly used, and peak releases from this type of disposal facility often are projected to occur within a few hundred years. In such cases, an assumption that present-day conditions will be maintained into the future is easily justified. However, many modern disposal facilities include massive concrete vaults with design lifetimes of many hundreds to thousands of years. Over such time frames, an assumption that recent conditions will continue into the future is more difficult to justify, because the historical record shows that the global climate of a few hundred years ago was significantly different than it is now. It is therefore tempting to introduce increasingly broad assumptions about potential future conditions as the expected durability of a disposal facility increases. As a result, use of a concrete vault, which has numerous short-term benefits, increases the number of potentially adverse conditions that might influence releases of long-lived radionuclides over longer time frames. In a regulatory framework that includes single-valued performance objectives, this point raises a serious question. If an engineered disposal facility is expected to maintain its integrity for a time sufficiently long that the occurrence of extreme events becomes more likely, should projected outcomes associated with such events be required to comply with performance objectives? If so, to the extent that waste contains significant amounts of long-lived radionuclides, designing a facility to meet performance objectives when extreme events at far future

336 / 7. UNCERTAINTY, SENSITIVITY AND IMPORTANCE ANALYSIS times are postulated would not necessarily be cost-effective, practical, or desirable. If not, a heuristic decision must be made to include some future conditions but exclude others. Regulatory authorities and reviewers have been reluctant to explicitly exclude extreme conditions from performance assessment. In current practice, it is common for an analysis to be restricted to current conditions, but this approach is based on subjective judgment and logical argument rather than a defined regulatory position. The issue described above has led to development of other approaches to incorporating alternative future conditions at a disposal site in a performance assessment. Broadly speaking, other approaches are based on probability theory, in that a particular future condition is assigned a weighting factor that represents a judgment about its probability of occurrence. One such method is a scenario approach that was developed for use in performance assessments of high-level waste disposal facilities (Helton, 1993). The performance measure of concern was the cumulative (time-integrated) release of radionuclides to the accessible environment over 10,000 y (EPA, 1993a).38 In that approach, a suite of credible scenarios (i.e., assumed future conditions at a site that affect calculated cumulative releases) is selected, and each scenario is treated by modeling the disposal system at steady-state and weighting the consequences by its likelihood. An assumption of steady-state scenarios is intended to provide a much simpler approach than modeling the time-dependent behavior of a system. By weighting the consequences of a scenario by its likelihood, the contribution of that scenario to the overall performance measure is obtained. For example, consider an assumed change in climate that produces an increased infiltration lasting for 1,000 y. In a scenario analysis, the steady-state cumulative discharge over the regulatory time frame of 10,000 y would be modeled by assuming the higher infiltration rate, and the result would be weighted by 0.1 to incorporate the contribution due to increased infiltration over 1,000 y (i.e., one-tenth of the regulatory time frame). It is important to note that the likelihood associated with this scenario is not the probability of 38At the time the analysis by Helton (1993) was conducted, EPA regulations in 40 CFR Part 191 (EPA, 1993a) applied to all facilities for disposal of spent nuclear fuel, high-level waste, or transuranic waste. At the present time, however, those regulations apply only to disposal of DOE’s defense transuranic waste at the WIPP Facility in New Mexico, but they do not apply to the proposed facility for disposal of spent nuclear fuel and high-level waste at Yucca Mountain, Nevada.

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its occurrence over 10,000 y, because the probability of a wetter climate during that time is essentially unity. Rather, this likelihood identifies the contribution of the scenario to the performance measure specified in regulations on the basis of its assumed duration. An interpretation of the likelihood of scenarios described above cannot be applied to estimates of maximum annual dose of concern to low-level waste disposal, because that performance measure essentially is a point estimate in time. If transport of radionuclides from a site to an assumed receptor location were sufficiently rapid under specified conditions of increased infiltration, the maximum annual dose would occur during that time and releases during other times of lesser infiltration would be irrelevant. Furthermore, the effect of increased infiltration on the maximum annual dose could depend on when increased infiltration occurs, given the usual mix of shorter-lived and long-lived radionuclides in low-level waste. Thus, it does not seem reasonable to weight an annual dose produced during a wet period with a small likelihood, because the probability of that dose occurring would be near unity if all other conditions assumed in an analysis were met. A scenario approach to addressing uncertainties associated with the future does have the advantage that it considers assumptions about the future in a formal manner. However, there are problems associated with using that approach in low-level waste performance assessments, in addition to the problem described above that derives from the particular form of performance objectives. First, because most low-level waste disposal facilities are located near the land surface, surficial events and processes that reasonably could occur over long time periods may become important (e.g., flooding or erosion following significant changes in surficial conditions, glaciation). Considering such processes would greatly complicate a performance assessment, and it is questionable whether it is reasonable to do so. Continental glaciation is a particular example of an event whose consequences certainly would be much more disruptive of human society than the consequences of any release of radionuclides from a disposal facility. An alternative approach to considering future conditions at a site would be to describe the future evolution using a timedependent model. That approach has been adopted in the United Kingdom for the purpose of comparing projected outcomes with a deterministic regulatory performance objective expressed in terms of risk (Section 3.4.4). The rationale for using that approach, instead of the scenario approach described above, is that extreme but still credible conditions should not be modeled as steady-state phenomena for the entire time period of an analysis. Recall that in

338 / 7. UNCERTAINTY, SENSITIVITY AND IMPORTANCE ANALYSIS the scenario approach, a maximum annual rainfall over 10,000 y is modeled as a steady-state process that lasts for 10,000 y, but projected releases then are weighted by a likelihood (relative frequency) of 10–4, since the scenario is presumed to affect only 10–4 of the integrated discharge over that time. In a time-dependent analysis, climate events that occur with a frequency of 10–4 would affect the outcome only if significant radionuclide migration occurs as a result of a single year’s rainfall. Although the alternative approach described above attempts to treat time-dependent processes and events in a more realistic fashion, it often is difficult to justify detailed assumptions that go into those kinds of analyses. Another clear disadvantage to a timedependent modeling approach is that transient modeling is much more difficult and time-consuming than steady-state modeling. Thus, given that the future evolution of a site is unknown but that some extreme events can be expected to occur with a high probability, that approach may do little to improve the level of confidence in results with respect to a decision about regulatory compliance. 7.5.4

Propagation of Parameter Uncertainty

Parameter uncertainty generally is treated by propagating uncertainties in input parameters through a model to identify their effects on model output. For the purpose of importance analysis, the intent is to identify parameter values that produce outputs that exceed regulatory performance objectives. Methods of propagating parameter uncertainty through models are reviewed elsewhere (Doctor, 1989; Doctor et al., 1988; Maheras and Kotecki, 1990; NCRP, 1996c; Zimmerman et al., 1990) and are not discussed in detail here. Rather, the following discussion focuses on an evaluation of available methods for purposes of performance assessment. The result of an accounting of uncertainties in model input parameters is either a distribution or a series of outcomes. Those outcomes must be compared with a single-valued performance objective, which suggests that there must be a means of identifying parameter values that can cause the performance objective to be exceeded and choosing which outcome of an analysis will be compared with the performance objective. These issues present difficulties for all methods of parameter uncertainty analysis. 7.5.4.1 Deterministic Methods. A frequently used approach to parameter uncertainty analysis is to develop a suite of results, obtained from multiple runs of a model, that reflects an analyst’s best estimates (or, alternatively, those estimates believed to be

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conservative) of appropriate parameter values. A variety of discrete parameter realizations may be analyzed in a sensitivity study to identify potential conditions for which performance objectives may be exceeded. If credible cases that could be assumed to represent expected conditions are identified in which performance objectives may be exceeded, additional data collection, design modifications, or model development may be needed. Otherwise, the most representative case is selected for comparison with performance objectives, and a discussion of other cases that were considered is provided. This approach has an advantage of simplicity of implementation and transparency of interpretation, in that a relatively small set of deterministic calculations is performed and model behavior is interpreted on the basis of results. In addition, communicating a single value of an outcome to a skeptical public is easier than communicating more complex approaches involving probability distributions of outcomes. The approach described above also has some drawbacks, however. First, assumed parameter variations, which often are considered singly or a few at a time, do not necessarily yield extremes of model output. Extremes of output may be caused by unusual combinations of input parameters that are not included in an analysis. In general, there is no way to determine the response of a model to variations in input parameters prior to an analysis, and troublesome combinations of parameters can be overlooked. Second, the approach does not lend itself to development of a clear decision rule. At the end of an analysis, the decision maker is left with a set of undifferentiated results. If some results exceed performance objectives, there is no clear criterion that can be applied in determining whether a facility is in compliance. If such a criterion is developed, it inevitably is based on judgment. On the other hand, judgment will be a necessary part of the decision process regardless of the approach to uncertainty analysis and regardless of whether a clear decision criterion can be formulated. Thus, there is a need for importance analysis to openly admit and assess this judgment. The type of uncertainty analysis described above can be evaluated by imposing weighting factors on each set of input parameters. That approach begins to look like a Monte-Carlo analysis described in the following section, but without the generality of such analyses. Indeed, undifferentiated results can be viewed as equally likely realizations of the space of input parameters, which therefore approximate inputs in the form of uniform probability distributions of input variables. However, inputs to this type of analysis are not random, and they may not span the space of possible input parameter values.

340 / 7. UNCERTAINTY, SENSITIVITY AND IMPORTANCE ANALYSIS 7.5.4.2 Probabilistic Methods. An alternative approach to parameter uncertainty analysis is to allow input parameters to vary freely over an assumed range of values. A series of individual sets of input parameters are chosen from probabilistic (Kozak et al., 1993) or possibilistic (Robinson and Grindrod, 1994) distributions. Those sets of input parameters then are propagated through a model to produce a probabilistic or possibilistic distribution of model outputs.

7.5.4.2.1 Monte-Carlo analysis. The most common probabilistic method of treating parameter uncertainty is Monte-Carlo analysis. In that method, discrete sets of input parameter values are selected at random from probability distribution functions, each set is run through the model, and a probability distribution function of model output is constructed. That distribution represents the uncertainty in model output associated with uncertain input parameters. Zimmerman et al. (1990) concluded that Monte-Carlo analysis is the most versatile method of parameter uncertainty analysis because: (1) it facilitates consistent propagation of uncertainties; (2) it can easily be applied to a series of linked models, such as those used in performance assessment (Kozak et al., 1990); (3) it does not require modifications of original models and, therefore, is generally straightforward to use; and (4) it is capable of dealing with large uncertainties in input variables by allowing a full stratification over specified ranges of all variables. In addition, the method can be used with nonlinear models (Helton et al., 1991). The primary advantage of Monte-Carlo analysis is that it provides model results for a large number of sets of input parameters. Thus, uncertainty in model output is acknowledged, and there is some means of identifying whether uncertainty in the output due to uncertainties in input parameters has been bounded, which increases the level of confidence in using results in regulatory decision making. Another advantage of the method is that sensitivities of model output to variations in input parameters may be identified (Zimmerman et al., 1990). This allows an identification of important model parameters for which efforts at further data collection may reduce uncertainty in the output. However, as discussed in Section 7.6, interpretation of sensitivity analyses of complex models may not be straightforward. It also should be emphasized, however, that Monte-Carlo analysis has many of the same difficulties as deterministic analyses. If variations in all input parameters are not considered, the most troublesome combinations of parameter values may be missed.

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Thus, judgment about which parameters should be varied will always be an important part of an analysis. Monte-Carlo analysis also does not lend itself to a clear decision rule regarding regulatory compliance. Many realizations are obtained, and judgment must be used to identify a particular case that will be used to support a regulatory decision. In this regard, it is important to appreciate that a probability distribution of model output from a MonteCarlo analysis does not provide a statistically rigorous estimate of the output, because inputs to an analysis involve many judgments. Rather, output distributions provide qualitative information about the relative likelihood of various outputs that is based on judgments about shapes and ranges of distributions of input parameters. The importance of qualitative and judgmental inputs to a specific problem should be considered in determining whether the additional effort required to conduct a Monte-Carlo analysis is warranted. Additional disadvantages of Monte-Carlo analysis as applied to performance assessment include that (1) many realizations of input data are required to span the range of possible values and (2) input parameters usually are treated as uncorrelated. However, both of these problems have been addressed to some extent. The required number of realizations can be greatly reduced by using a stratified sampling strategy, such as Latin hypercube sampling (Iman et al., 1981), and techniques are available that allow assumed correlations among variables in implementing the Latin hypercube sampling method (Iman and Conover, 1982; Iman et al., 1981). However, even with Latin hypercube sampling, an extensive computing effort may be required in a performance assessment when complex models are used. Although measures of central tendency of an output distribution normally can be established using only a modest number of realizations, many more realizations (perhaps thousands) often are required to establish the behavior of the upper tail of an output distribution, which is generally where attention will be focused. Use of Monte-Carlo analysis in performance assessment can lead to a belief that all sources of uncertainty have been taken into account. This belief is mistaken because, as discussed previously, addressing parameter uncertainty does not mean that model uncertainty has been addressed. Furthermore, shapes of frequency distributions of parameters that must be assigned are themselves uncertain. Probabilistic analyses can be difficult to interpret, and they are difficult to communicate to general audiences. There is a danger that an analysis will be interpreted as being much more rigorous than it is.

342 / 7. UNCERTAINTY, SENSITIVITY AND IMPORTANCE ANALYSIS An issue that frequently is raised about Monte-Carlo analysis is the difficulty in assigning the precise shape of the probability distribution function of each parameter. There are two aspects of this issue that need to be addressed. The first, and the one most easily dealt with, is the large amount of data required to establish a distribution function. This concern is important only if a distribution is intended to represent a Type-A (stochastic) uncertainty. However, Type-A uncertainties are expected to be unimportant in performance assessment compared with Type-B (lack of knowledge) uncertainties (Section 7.4.1). Thus, the second concern of assigning shapes of parameter distribution functions to represent informational uncertainty requires more careful consideration. Judgment is important in assigning this type of uncertainty. Practitioners of Monte-Carlo analysis in performance assessment have used several approaches to specifying probability distribution functions of model input parameters. Some have attempted to identify distribution functions that should reproduce an expected central tendency of the actual behavior of a disposal system (DOE, 1996d; WIPP, 1992). The intent of this approach is to conduct an uncertainty analysis, rather than an importance analysis. Others have recognized the informational nature of input distributions and have proposed that Shannon’s entropy (Hamming, 1991; Shannon, 1948) be used to specify distributions on the basis of the amount of available information (Harr, 1987; Kaplan, 1991). Still others have used a less formal approach in an attempt to bias an analysis to produce a conservative result (Baer et al., 1994; Price et al., 1993). In that approach, some parameters may be assigned a single conservative value, when such a value can be identified in advance of an analysis, whereas other parameters are assigned conservatively biased probability distributions, when such a bias is believed to be identified, or unbiased distributions if a bias toward conservatism cannot be identified. All three approaches recognize that assumed distributions of model input parameters are subjective, and each approach represents an alternative way of constructing distributions to represent that subjectivity in results of an analysis. The foregoing discussion indicates that there is not a consensus about the most suitable approach to assigning probability distribution functions of model input parameters for use in performance assessment. If probabilistic approaches are used, distribution functions should be regarded as a means of representing the state of knowledge about input parameters, rather than actual distributions that would result from repeated measurement. In this sense, there is benefit in representing uncertainty using distributions,

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because certain kinds of judgmental information can be represented that are lost when undifferentiated deterministic analyses are used. Ranges of distributions often can be established by measurements, but their shapes generally will need to be established on the basis of judgment. For example, there may be information that a parameter is more likely to be at one end of its range than the other, and that information can justifiably be used to bias a result in that direction. Alternatively, such a bias in input distributions can justifiably be introduced if there is knowledge that model output will be biased conservatively. In general, however, caution must be used when attempting to establish that an assumed parameter value or distribution is conservative prior to conducting an analysis. 7.5.4.2.2 Perturbation analysis. Another proposed method of treating parameter uncertainty in performance assessment involves use of perturbation analyses (Gelhar, 1986), which also are referred to as analytical stochastic methods. The method is conceptually similar to Monte-Carlo analysis, in that distributions of model input parameters are used to estimate distributions of outputs. However, by using simplifying assumptions, probability distribution functions of model inputs are incorporated explicitly in model equations and, thus, are propagated directly into model outputs. This method has been used in groundwater flow and transport calculations, for example, by Polmann et al. (1988). Perturbation analyses are computationally more efficient than Monte-Carlo methods, and they can be used to develop an improved conceptual understanding of groundwater processes by obtaining closed-form solutions of governing equations. For example, it is well known that dispersivities obtained from laboratory column studies may differ by orders of magnitude from those observed in the field. Analytical perturbation models have been developed that can, in principle, be used to predict field-scale dispersivities from laboratory-scale measurements (Dagan, 1984; 1986; Gelhar and Axness, 1983; Naff et al., 1988). Other studies using perturbation methods (Vomvoris and Gelhar, 1990) have provided information on the expected variance of a contaminant concentration at a given sampling location. Still other examples relate to predictions of the moisture-dependent anisotropy of hydraulic conductivity in unsaturated heterogeneous media (Yeh et al., 1985). In general, analytical stochastic methods have been limited in their scope and flexibility, in that they typically are derived for a single type of input parameter distribution (e.g., lognormal) and for a limited set of parameters. The method thus is too restrictive at the present time to be of much use in performance assessment.

344 / 7. UNCERTAINTY, SENSITIVITY AND IMPORTANCE ANALYSIS 7.5.4.2.3 Possibilistic analysis. Treatment of uncertainty using possibilistic analysis is similar to the approach used in probabilistic analyses, but the significance of results is expressed differently (Robinson and Grindrod, 1994). Results of possibilistic analyses can be expressed qualitatively using such common terms as “very likely,” “reasonable,” or “unlikely.” Expressing results in this way eliminates a misperception of mathematical rigor that sometimes is associated with probabilistic analyses. The full implications of possibilistic theory for performance assessment have not yet been explored. Other related approaches to propagating parameter uncertainty, such as fuzzy set analysis, also are not well established for application to performance assessment. Those methods are not discussed in this Report. 7.6 Role of Sensitivity Analysis in Importance Analysis 7.6.1

Need for Sensitivity Analysis

The need for sensitivity analysis in performance assessment, or in any predictive modeling, has been widely recognized (IAEA, 1995a; 1999; Kozak, 1994a; NAS/NRC, 1990; NEA, 1991; Seitz et al., 1992a; Vovk and Seitz, 1995). Sensitivity analysis has two functions in the context of importance analysis: (1) it provides information to identify parameters and assumptions that can affect a decision about regulatory compliance; and (2) it provides information that can be used to identify parameters or assumptions for which additional data collection or design modifications should provide the most benefit in terms of building confidence in a decision about compliance. In the context of performance assessment, sensitivity analysis has a specific focus, in that it does not matter that a small change in an input variable results in a large change in model output (dose) if the change does not affect a regulatory decision. However, given the degree of conservatism that is often incorporated in performance assessment to compensate for a paucity of relevant site-specific data, some combinations of assumptions may result in outcomes that exceed performance objectives, especially in early stages of an analysis. It is in such cases that sensitivity analysis is important. In importance analysis, the purpose of sensitivity analysis is to identify assumptions about a parameter value or set of parameter values which, when changed, can result in a different regulatory decision. This purpose is more focused than the more traditional

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aim of quantifying a change in model output resulting from a change in one or more input parameters. This difference suggests that some methods used in traditional sensitivity analysis need to be modified or implemented in different ways in generating the kinds of information needed in importance analysis. Sensitivity analyses generally have focused only on parameter variabilities and, thus, have provided information only within the context and limitations of a particular model structure. Mathematical approaches to sensitivity analysis available for use in importance analysis are identical to ones developed for use in traditional sensitivity analysis. However, interpretation of results and the significance of an analysis differ in much the same way as uncertainty and importance analyses differ. 7.6.2

Methods of Parameter Sensitivity Analysis

Methods of parameter sensitivity analysis can be broadly classified as either differential or regression. Either type of analysis can be applied to raw data or to ranked (ordered) values. Differential analysis derives from the well-known method of expanding Taylor series solutions about a particular point in a parameter domain to provide an approximation of an output variable in the neighborhood of that point. An advantage of differential analysis is that it often provides an analytical expression of sensitivity on the basis of the form of underlying equations. The main disadvantage is that it is a local technique (i.e., it only provides information about sensitivity in the neighborhood of a particular point in the space of possible input parameter values). Regression analysis encompasses a broad range of methods, including step-wise regression and response-surface methods. In those approaches, previously generated outcomes from an uncertainty analysis are linked to input parameters through an approximate empirical equation. Most commonly, those equations are linear to permit application of standard regression and response-surface methods. Regression methods have an advantage that they provide global sensitivities; i.e., they can be applied over any range of input parameters and model outputs. However, outputs from performance assessment models frequently are complicated and nonlinear, and it thus is difficult to evaluate sensitivities using simple linear-response models that are approximations of models used in an analysis. Given the difficulties with more sophisticated techniques, it often is useful and defensible to conduct point sensitivity analyses, which are local analyses in which one or more parameters are varied around particular values and effects on model output are

346 / 7. UNCERTAINTY, SENSITIVITY AND IMPORTANCE ANALYSIS evaluated on the basis of an understanding of the physics of the problem and the mathematical model being used (IAEA, 1999). If some results exceed performance objectives, combinations of parameter values that produced those results are reviewed. Thus, sensitivity analysis becomes an effort to confirm which assumptions or parameters are driving results that exceed performance objectives. In such an analysis, it is not necessary to quantify sensitivities. Rather, it is only necessary to gain an understanding of the impact on model output of changes in different parameters. In summary, although techniques of sensitivity analysis are not difficult to apply, their interpretation may not be straightforward. Nonlinearities in system behavior may obscure sensitivities that are derived using linear, multivariate regression methods, and methods that provide only local estimates of sensitivities may not provide adequate information throughout the space of input parameters. No single technique of sensitivity analysis has been shown to be generally applicable to an interpretation of outputs of Monte-Carlo analyses. 7.7 Application of Uncertainty Analysis to Importance Analysis 7.7.1

General Structure of Uncertainty Analysis

A general structure of uncertainty analysis that involves separate treatments of future uncertainty, model uncertainty, and parameter uncertainty (Kozak et al., 1993) is shown in Figure 7.1. In this structure, particular conceptualizations of the future of a disposal site are embodied in different scenarios. Within each scenario of the future, it is possible to postulate alternative conceptual models of the behavior of the disposal system, each of which leads to a particular mathematical model to describe that behavior. For each model, it is possible to postulate alternative sets of input parameter values. In the context of performance assessment, the primary purpose of an uncertainty analysis is to support a decision about compliance of the disposal system with regulatory requirements, rather than to estimate the uncertainty in actual outcomes (e.g., maximum annual dose). As described previously, different approaches to uncertainty analysis that can be applied to the general structure in Figure 7.1 have been developed. Probabilistic and possibilistic analyses attempt to include many possible combinations of scenario, conceptual model, and sets of parameter values. Those methods compete with deterministic approaches that involve selected variations in one or more parameters.

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Fig. 7.1. General structure of an approach to uncertainty analysis for use in performance assessment (Kozak et al., 1993).

7.7.2

Evaluation of Different Methods of Uncertainty Analysis

Deterministic and probabilistic approaches to uncertainty analysis each face difficulties when applied in the process of regulatory decision making. To illustrate the difficulties, the following hypothetical (but common) decision process involving performance assessment is posed. First, several independent modeling teams conduct an assessment, and the different teams do not all use the same conceptual or mathematical models. Second, for each model, results of numerous deterministic realizations of sets of input parameters are obtained, as will be the case in probabilistic or deterministic methods. Finally, results of some of those realizations exceed regulatory performance objectives. In deterministic approaches, a decision maker is presented with an undifferentiated set of results (Section 7.5.4.1), meaning that an analysis gives no indication of whether some outcomes are more likely than others. The essence of the decision problem then is to identify a combination of model and set of parameter values that represents the “best” case for making a regulatory decision, and to evaluate potential conditions that deviate from those assumptions

348 / 7. UNCERTAINTY, SENSITIVITY AND IMPORTANCE ANALYSIS for which regulatory criteria may be exceeded. One such approach is to select a model and set of parameter values to be used as the primary basis for a regulatory decision prior to conducting an assessment, and to describe potential conditions that may result in higher projected outcomes only qualitatively. The primary advantage of that approach is its simplicity and transparency. However, there are three important drawbacks with a deterministic approach that involves selection of a single set of assumptions prior to conducting an assessment. First, given the complexity and spatial variability of disposal systems and nonlinearities in system performance, it is difficult to choose the “best case” conditions for a regulatory decision solely on the basis of judgment, and to justify that choice. Second, it is debatable whether “best case” conditions are the ones that should be used to support a regulatory decision. Finally, there is no clear rule for taking into account credible conditions other than the base case that could result in outcomes that exceed regulatory criteria. Problems described above could be addressed by choosing conditions for a regulatory decision after a full uncertainty analysis is conducted. That is, an uncertainty analysis provides a suite of potential deterministic results, one of which is chosen as the basis for a decision. That approach has an advantage that there is no need to prejudge which conditions (models and parameter values) provide the “best case” for describing system performance. Rather, a decision is made on the basis of a range of possible outcomes, and a decision maker can be as risk averse as desired. A disadvantage of the approach is that selection of any particular result from a large set of possible outcomes as the basis for a decision is largely arbitrary. In probabilistic approaches, a decision maker is presented with a seemingly improved situation, in that results of a performance assessment are differentiated by their estimated probability of occurrence. Therefore, it appears that if a sufficiently small portion of a probability distribution of projected outcomes exceeds the performance objectives, there would be a correspondingly high probability that the facility will be in compliance. Unfortunately, however, probability distributions of model outputs do not have this significance. Distributions of outputs represent an aggregation of many expert judgments about models and parameter values, and they are no more than a formal representation and documentation, to the extent possible, of the state of knowledge about the disposal system. Therefore, even though a regulator may specify that compliance would be indicated if an upper confidence limit of a probability distribution of model outputs did not exceed a performance

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objective, such an objective comparison may not be adequate to support a regulatory decision (Section 3.4.3.4). That is, in advance of an analysis and in the absence of other detailed considerations about an analysis, it is difficult to identify an upper confidence limit in a probability distribution of model outputs that would represent an appropriate degree of confidence in a regulatory decision. Instead, it may be more appropriate to take into account the believed conservatism of an analysis, and to assign an upper confidence limit for a decision on the basis of considerations of the specific site, facility design, and models. In this approach, judgments required in making a regulatory decision are similar to the judgments required in the case of a deterministic assessment, because a decision maker must choose a case that represents the desired degree of conservatism in an analysis. Probabilistic methods do have two advantages compared with deterministic analyses. First, the extent to which the space of input parameters is spanned in a Monte-Carlo analysis is greater than in an analysis that involves variations of selected parameters. Thus, there is greater confidence that potentially adverse conditions have been evaluated in an assessment. Second, a probabilistic analysis provides an estimate of the likely importance of any particular outcome (e.g., annual dose) by assigning a likelihood to each outcome. This characteristic of probabilistic analysis allows some differentiation among outcomes. However, it must be understood that this differentiation is qualitative and judgmental, rather than rigorously quantitative and statistical. Probabilistic methods also have serious limitations, however. Although such analyses reduce judgmental aspects of a regulatory decision somewhat, the need for judgment is not eliminated. Whereas in deterministic analyses judgment is applied to values of input parameters to be used in establishing the “best” assessment for purposes of regulatory decision making, in probabilistic analyses judgment is applied to ranges and shapes of distributions of input parameters and to the likelihood that a calculated outcome may exceed regulatory criteria. The judgmental element of a decision is not eliminated in probabilistic analyses but is posed in different terms. One of the important drawbacks of probabilistic methods is the common misperception that judgmental aspects have been eliminated, rather than simply reposed. A related drawback is the complexity of concepts of probability theory and the potential for misinterpretation of results. Most commonly, misinterpretations lead to a perception that probabilistic analyses are technically more “correct” than other types of analyses.

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Evaluation of Approaches to Importance Analysis

A flexible approach to importance analysis is most appropriate for performance assessment. At some sites or for some release, transport or exposure pathways at a particular site, it should be possible to conduct clearly conservative deterministic analyses and thereby demonstrate regulatory compliance without placing unreasonable restrictions on waste acceptance criteria at a facility. In such cases, there is no need to conduct more detailed assessments to make a confident decision. However, that type of clarity in an assessment generally is restricted to simple models. Once model complexity increases, it becomes difficult to determine which assumptions and parameter values are appropriate for use in regulatory decision making. Instead, effects of multiple lines of reasoning (i.e., conceptual models) and variations of parameters over reasonable ranges on the projected behavior of a system need to be investigated. There is no consensus within the waste management community on the issue of how alternative conceptual models or parameter sets should be chosen. Competing arguments in favor of different approaches are summarized as follows. Advocates of deterministic methods maintain that: (1) it is easier to choose a single value of a parameter than to choose an uncertain distribution; (2) results of an assessment are more transparent; (3) probabilistic methods are unduly complex; (4) probabilistic methods give an unwarranted appearance of rigor; and (5) subjectivity is important in either approach. Counterarguments include that: (1) it is impossible to make a supportable regulatory decision without adequately evaluating uncertainty in projected outcomes; (2) single values of input parameters cannot be chosen meaningfully when models are nonlinear; (3) it is impossible to understand how an outcome of one assessment relates to any other; and (4) a deterministic analysis represents an approach to ignoring uncertainty rather than accounting for it. Advocates of probabilistic methods maintain that: (1) they represent the best available approach to representing the degree of uncertainty (i.e., state of knowledge) in performance assessment; (2) other approaches tend to underestimate uncertainty; (3) they provide a mathematical representation (albeit judgmental) of the relative importance of differing deterministic analyses; and (4) they can provide a rational basis for decisions to exclude unusual combinations of conditions and parameter values in reaching a regulatory decision. Counterarguments include that: (1) probabilistic analyses are unduly complex; (2) they are difficult to interpret and convey to the public; and (3) comparisons of arbitrary points on a probability

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distribution of model outputs with performance objectives is as difficult to justify as any other way of choosing the “best” case to be used in reaching a decision about compliance. A possible alternative is to use a combination of deterministic and probabilistic methods. Advocates of such approaches suggest that there are benefits to conducting probabilistic analyses, but that they should be used for the benefit of analysts, rather than as the basis for a regulatory decision. That is, a probabilistic analysis should form the foundation from which a single deterministic analysis is chosen to provide the basis for a decision. However, it can be argued that this approach is really no different than choosing a particular confidence level of a probability distribution as the basis for a decision, and that a particular deterministic calculation that might be chosen cannot be defended using any objective criterion. There is not a clear technical basis for choosing from among alternative approaches to importance analysis summarized above. One possibility is to address uncertainty using Monte-Carlo analysis and use point sensitivity analyses to identify parameters that have the greatest influence on results. However, the fundamental problem with any approach is that decisions about compliance of waste disposal systems with regulatory criteria are not based primarily on rigorous and objective technical analysis. Rather, decision making in the presence of uncertainty is largely a qualitative and judgmental exercise (Section 3.5.3), and results of performance assessment can provide only part of the basis for regulatory decisions. Thus, the preferred approach to uncertainty analysis is one that provides the best representation of the state of knowledge about a disposal system with respect to achieving confidence in a regulatory decision. The particular choice will depend on characteristics of the disposal site, the design of the facility, properties of waste intended for disposal, and considerations of the defensibility of a decision. 7.8 Summary The primary reason for conducting performance assessments is to provide decision makers with information about a proposed disposal practice that is relevant to a regulatory decision. An appropriate treatment of uncertainties in performance assessment and sensitivities of results to changes in particular assumptions is an essential part of an assessment. Given the subjective nature of performance assessment, the variety of disposal sites for low-level waste and facility designs, and the variety of wastes requiring disposal, it is not reasonable to attempt to develop a prescriptive

352 / 7. UNCERTAINTY, SENSITIVITY AND IMPORTANCE ANALYSIS approach to achieving “reasonable assurance” of compliance of near-surface disposal systems with regulatory requirements in the presence of uncertainty that could be applied in a cost-effective manner at all sites. Rather, appropriate approaches to accounting for uncertainty and sensitivities should be considered on a sitespecific basis. Uncertainties associated with performance assessment arise from a number of sources, but in all cases the dominant type of uncertainty is lack of knowledge. Lack of knowledge can be about processes, model representations, parameterizations of models, values of model parameters, future conditions, or human behavior. In each of these areas, there is an irreducible level of uncertainty that must be addressed by analysts and decision makers. Several technical approaches have been developed to represent uncertainties and sensitivities mathematically. There is no consensus on the best approach; indeed, the most defensible approach depends on specific features of a disposal system of concern. However, of greater importance than a particular approach is the process for considering and documenting uncertainties and sensitivities. Regardless of the particular method chosen, an analyst must present a credible argument that uncertainties have been addressed in a manner that adequately supports the resulting regulatory decision. Approaches to uncertainty and sensitivity analysis described in this Section should be viewed in the context of the process of performance assessment described in Section 4. A properly documented assessment process and a credible treatment of uncertainty and sensitivity are what make a defensible product. This Section has emphasized that the purpose of uncertainty and sensitivity analysis in performance assessment is fundamentally different from the purpose of traditional approaches to such analyses. In performance assessment, the purpose of uncertainty and sensitivity analysis is to identify conditions that might result in outcomes that exceed applicable performance objectives. Indeed, results of performance assessment can be quite uncertain in regard to predicting actual outcomes and, at the same time, quite convincing in regard to supporting a regulatory decision that reasonable assurance of compliance has been achieved. To emphasize this difference, the term “importance analysis” is used in this Report to describe the kind of uncertainty and sensitivity analysis that is needed to supporting regulatory decision making.

8. Summary This Report has discussed performance assessments of near-surface facilities for disposal of low-level radioactive waste, including the purpose and scope of performance assessment and approaches to modeling the long-term performance of such facilities that experience has shown should be acceptable for use in developing and licensing new facilities. 8.1 Purpose and Scope of Performance Assessment Performance assessment is concerned with evaluating the capabilities of waste disposal systems (i.e., disposal units, including any engineered barriers, and the natural geologic environment surrounding disposal units) to limit potential radiological impacts on humans and the environment over long time periods in the future. As emphasized in this Report, the essential purpose of performance assessment is to investigate important characteristics of a disposal site, a disposal facility, and waste itself that can lead to a finding of reasonable assurance of compliance with regulatory performance objectives related to protection of human health and the environment. Thus, performance assessment is used to identify critical data, facility design, and model development needs for defensible and cost-effective licensing decisions and to develop and maintain operating limits (e.g., waste acceptance criteria) at specific disposal facilities. Performance assessment is fundamentally an exercise in subjective scientific judgment, essentially because it is concerned with evaluating unknown and unknowable outcomes of waste disposal on the basis of limited information on the actual long-term performance of various components of a disposal system. It is important that analysts, decision makers, and other stakeholders respect the limitations of performance assessment in predicting actual outcomes, and it is equally important that all parties appreciate that those limitations do not compromise the essential role of performance assessment in regulatory decision making. Another essential characteristic of performance assessment is that it is a process that is iterative in nature and involves interactions with other disposal activities (e.g., design and operation of disposal facilities, treatment and packaging of wastes) and with regulators. 353

354 / 8. SUMMARY Given the limitations of performance assessment, most assessments intend to incorporate some degree of conservatism in modeling. A central issue that must be confronted in performance assessment is an appropriate balance between use of simple and clearly conservative models and use of more complex and rigorous approaches. Although a desire for realism is understandable, there are many difficulties in attempting to model the long-term performance of disposal facilities realistically. The point of view taken in this Report is that (1) there is no single prescription for balancing conservatism and realism in performance assessment that is appropriate at all sites and (2) the goal at any site should be to provide a cost-effective and defensible performance assessment that is commensurate with the characteristics of the site and disposal facility and hazards posed by wastes that are intended for disposal. The use of models that underestimate the waste-isolation capabilities of disposal facilities is an appropriate way of compensating for a lack of knowledge about many important aspects of long-term performance, and it strengthens the case for compliance with performance objectives. It also should be emphasized, however, that it is not always easy to justify that a model is conservative, and such justification should always be provided as part of an assessment.

8.2 Basic Elements of Performance Assessment Performance assessment can be a complex undertaking, depending on such factors as the characteristics of a disposal site, the design of a disposal facility, the characteristics of waste intended for disposal at a site, and the demands of regulatory authorities and other parties with an interest in licensing decisions. However, the process of performance assessment can be viewed as consisting of three basic elements: 1. development of conceptual models of the long-term performance of different parts of a disposal system; 2. development and implementation of mathematical and, perhaps, physical models to describe the performance of different parts of a disposal system on the basis of assumed conceptual models; and 3. integration and interpretation of results of an assessment with respect to demonstrating compliance with applicable performance objectives.

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Proper attention to all these elements is essential to successful conduct of performance assessment. The greatest effort normally is expended on development and implementation of mathematical models used in an assessment. However, experience has shown that when an assessment is critically reviewed by technical peers and licensing authorities, greatest scrutiny often is given to the conceptual models underlying an analysis and to the integration and interpretation of results. 8.2.1

Development of Conceptual Models

Development of conceptual models of the long-term performance of waste disposal systems is the foundation of performance assessment. No amount of skill or ingenuity in developing and implementing mathematical models can overcome deficiencies in conceptual models. Therefore, any attempt to short-circuit the process of developing and documenting conceptual models likely will result in a judgment that a performance assessment is unacceptable. An example of an approach that likely would be judged unacceptable would be to choose available computer codes to perform calculations without investigating assumptions embodied in the codes and their applicability to a disposal site of concern. A conceptual model is a qualitative description of assumptions about the behavior of a disposal system. For example, a conceptual model of the component of performance assessment concerned with pathways of human exposure would include a description of assumptions about different activities that humans undertake and exposure pathways that result from those activities. Similarly, a conceptual model of the performance of engineered barriers would include a description of various mechanisms that are assumed to result in failure of barriers over time, and a conceptual model of groundwater flow would include a description of assumptions about the nature of flow (porous or fracture) and characteristics of an aquifer (e.g., stratigraphy, flow boundaries, areas of recharge and discharge, degree of isotropy and homogeneity). Conceptual models also are needed for aspects of the performance of a disposal system that will be ignored in an assessment. For example, the influence of biotic intrusion on release and transport of radionuclides is not often taken into account, but justification for ignoring biotic intrusion should be developed as part of an assessment. A conceptual model should be broadly consistent with available site-specific information or with generic information that reasonably can be applied, with proper justification, to a disposal site

356 / 8. SUMMARY of concern. It also is important to investigate if substantially different conceptual models are consistent with available information. This is especially the case, for example, in developing conceptual models of groundwater flow at specific sites. If conservative assumptions are used in developing conceptual models, they should be so identified. 8.2.2

Development and Application of Mathematical and Physical Models

Development of mathematical and physical models used in performance assessment should be based directly on assumed conceptual models. Results of observations of physical models (e.g., natural analogs, tests of leaching characteristics of waste, tests of concrete barriers) normally would be incorporated in mathematical models. Probably the most important issue in developing and implementing mathematical models is the need for quality assurance of the entire process. Data and assumptions used in models must be documented and traceable, and computer codes must be tested (e.g., by benchmarking and verifying that a code is free from significant numerical error). As an initial step in the process of developing and implementing mathematical models, simple and clearly conservative screening models can be used to foster an efficient approach to performance assessment. In principle, a large number of radionuclides, several release and transport pathways, and many exposure pathways must be considered. However, by use of properly documented screening models, unimportant radionuclides and pathways can be eliminated from further consideration, and an assessment then can focus on more important aspects of system performance. Most mathematical models used in performance assessment are implemented by means of computer codes. Given the many processes that normally are taken into account and the complexities of some processes, a suite of codes, each of which addresses a particular aspect of performance, normally is used in site-specific analyses, rather than a single code that incorporates models of all processes of concern. When a suite of codes is used, it is important to determine that outputs from one code are compatible with required inputs to the next, and that mass balance and consistency of units are maintained. Use of a suite of codes also offers some advantages. A code used to model a particular process can be replaced, when a better model becomes available, without having to make extensive changes to

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other codes. An ability to examine outputs of each code is an important means of quality assurance and of checking that intermediate results are reasonable. 8.2.3

Integration and Interpretation of Results

The desired endpoint of a performance assessment normally is a projected dose (usually a maximum annual dose) to a member of the public. Thus, projected maximum concentrations of radionuclides in groundwater or surface water usually are needed, even if separate requirements for protection of water resources are not imposed. Results of an assessment then are compared with applicable performance objectives for the purpose of making a judgment about whether a disposal facility is in compliance with those objectives. Without regard for the magnitude of projected doses or concentrations of radionuclides in the environment, it generally is not sufficient to present results of a performance assessment and a comparison with performance objectives and assume that licensing authorities and other interested parties will appreciate that a case for compliance with regulatory requirements has been made. Given the many assumptions embodied in a performance assessment and the many processes that need to be modeled, it also is necessary to identify those assumptions that were important in obtaining the results and whether they are sufficiently robust to support a decision of compliance. An assumption is robust if plausible changes would not affect a decision. A description of what a performance assessment means in regard to demonstrating compliance with performance objectives is referred to as an integration and interpretation of results. Proper integration and interpretation of results is an essential aspect of performance assessment. Given that performance assessment models often will incorporate a considerable degree of conservatism and, further, that plausible ranges of assumptions and input data normally will result in a wide range of projected outcomes (e.g., maximum annual doses), it often will be the case that some projections exceed performance objectives. In such cases, the essential role of an integration and interpretation of results is to identify the conditions (assumptions) that caused those outcomes, to discuss the significance of those outcomes in regard to rendering a licensing decision (e.g., whether they are plausible), to discuss needs for additional data to refine the calculations and reduce uncertainty about whether compliance has been demonstrated, and to discuss changes in facility design, waste emplacements, or waste acceptance criteria that would be needed to obtain reasonable assurance of compliance.

358 / 8. SUMMARY 8.3 Components of Performance Assessment Modeling The usual approach to performance assessment is to divide an analysis into components, also referred to as modules, that are conceptually or physically separate and describe different aspects of the performance of a disposal system. That approach is reflected in discussions of performance assessment models in Section 5. The following sections present summary comments on the different components of performance assessment. Technical challenges in modeling different aspects of the performance of a disposal system are particularly emphasized. 8.3.1

Cover Performance and Infiltration

Covers on near-surface disposal facilities are intended to inhibit infiltration of water into a disposal facility, as well as intrusion by plants and animals. The purpose of an analysis of cover performance is to estimate the flux of water that infiltrates into a disposal facility as the cover fails over time on the basis of assumptions about the flux of precipitation onto the ground surface and how properties of the cover that control infiltration will change over time. Many disposal concepts include an elaborate cover system that could greatly inhibit infiltration of water over long periods of time. However, available data on cover performance generally are insufficient to allow confident predictions of infiltration far into the future, and performance assessments often give minimal credit to a cover system. For example, a cover often is assumed to function as designed (and perhaps tested) during an assumed period of institutional control (e.g., 100 y) but to fail instantaneously at the end of the control period, at which time the infiltration rate often is assumed to be the same as the natural infiltration rate in undisturbed surface soil. When credit is taken for cover performance, it usually is based on generic, rather than site-specific, data. The natural infiltration rate at a site normally can be estimated on the basis of site-specific information. At sites with abundant precipitation, uncertainty in the infiltration rate usually is small compared with the infiltration rate itself. At arid sites, however, the infiltration rate relative to the precipitation rate can be highly uncertain. That difficulty often is addressed by using a conservative assumption about the fraction of precipitation that infiltrates through surface soil. At the present time, there is no generally applicable method of estimating infiltration through soil or engineered covers. Transient

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behavior of infiltration during episodic precipitation events, which can occur at any site, usually is ignored. It generally is difficult to justify an assumed performance of an intact cover over times much beyond a period of active institutional control. A potentially important concern, especially at moist sites, is the occurrence of preferential paths of infiltration into a facility following failure of an engineered cover that could result in infiltration rates that exceed normal levels in undisturbed soil. Increases in infiltration compared with natural levels are not often considered in performance assessments. 8.3.2

Performance of Concrete Barriers

Many disposal concepts employ engineered concrete structures that are intended to provide support for an earthen cover, delay and inhibit inflow of water to waste, supply material to adsorb radionuclides, and delay and inhibit release of leachate. There is historical evidence that some types of concrete can maintain their structural integrity for thousands of years. However, that evidence has provided little relevant information that can be used to predict the physical integrity and load-bearing capabilities of structures built using modern concrete formulations over times beyond a few hundred years and to predict performance under a variety of environmental conditions. Important chemical properties of concrete (e.g., pH, adsorption capabilities) should persist for much longer. Hydraulic properties of concrete that are important to controlling radionuclide transport also may persist beyond the period of structural stability, although firm evidence for this is lacking. Models have been developed to describe flow of water through concrete. This is a difficult modeling problem when concrete is unsaturated, as is normally the case when a disposal facility is constructed above the water table. An additional complication is that concrete soon develops cracks which, if they extend through a member, can greatly increase the infiltration rate of water. However, infiltration through fractured, unsaturated systems also is poorly understood. Models of degradation of concrete barriers are concerned mainly with predicting when a concrete member can no longer bear loads imposed upon it, at which time concrete usually is assumed to no longer function as a physical barrier. Processes that degrade concrete can be grouped into those that degrade the surface, those that affect bulk properties, and those that corrode reinforcing steel. Models often used in performance assessment, which are largely empirical and generic, indicate that reinforced concrete

360 / 8. SUMMARY may maintain its physical integrity for perhaps a few thousand years. However, those predictions have not been tested under field conditions. Models that consider the combined effects of multiple degradation mechanisms are becoming available, but are not yet widely used in performance assessment. Those models do not indicate that concrete would maintain its physical integrity for times longer than those normally assumed in performance assessment. In addition, there still can be considerable value in understanding individual mechanisms and the ability of particular design features to enhance the performance of concrete. Given the difficulties in justifying predictions of infiltration of water through concrete barriers over time and predictions of the time over which a concrete barrier will maintain its structural integrity, many performance assessments do not take credit for the performance of concrete at times beyond a few hundred years. Thus, concrete barriers usually are assumed to inhibit releases of shorter-lived radionuclides only (e.g., 3H, 90Sr, and 137Cs). Little credit is taken for the ability of concrete barriers to inhibit releases of longer-lived radionuclides, although there can be some benefit if the time period over which compliance with performance objectives must be demonstrated is relatively short (e.g., 1,000 y). However, claims that concrete will perform as intended for only a few hundred years still require proper justification. 8.3.3

Source Term

Modeling of the source term (i.e., rates of release of radionuclides from a disposal facility as a function of time after facility closure) is a key component of performance assessment for any disposal concept in any environment. The source term is determined by many factors including: inventories of radionuclides in disposed waste; physical and chemical processes that degrade engineered barriers, waste containers, and waste forms; and processes that alter chemical conditions within a disposal facility. Modeling of the source term involves consideration of the degradation of engineered barriers to an extent sufficient to permit fluids to contact waste containers, degradation of waste containers to permit access of fluids to wastes, release of radionuclides from waste forms, and transport of released radionuclides from a disposal facility. Releases of radionuclides in both liquid and gaseous forms must be considered. Models of the performance of waste containers and waste forms and models of transport of radionuclides in aqueous and gaseous phases through a disposal facility have been developed. However,

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there rarely will be site-specific data that can be used in implementing models, and there will be little, if any, data that directly address releases from a disposal facility after emplacement of waste. A major concern at many sites is modeling of the source term when a facility contains highly heterogeneous distributions of radionuclides and a multiplicity of waste forms. Since it is difficult to model the source term realistically under the best of conditions (e.g., when radionuclides are distributed fairly homogeneously and only a single waste form is used), it is imperative when wastes are distributed heterogeneously to use stylized and simplified representations of a disposal facility and different processes that affect the source term, such as assumptions of uniform distributions of radionuclides and uniform infiltration throughout a facility. Therefore, the emphasis in performance assessment should be placed on describing the source term using simple and defensible models that allow a finding of compliance with performance objectives. To this end, it often is appropriate to represent important processes conservatively, and even to ignore some processes if they do not need to be invoked in demonstrating compliance. Estimates of inventories of radionuclides in disposed waste are an important concern in evaluating the source term. Radionuclides that are easy to measure in low-level waste (i.e., radionuclides that are high-energy photon emitters) usually are unimportant in performance assessment, because they tend to have relatively short half-lives (e.g., ~30 y or less) and are expected to decay to innocuous levels before release and transport to locations of potential exposure could occur. Radionuclides that tend to be the most important to projected doses beyond a site boundary are low-energy beta emitters (e.g., 14C, 99Tc, and 129I) that are among the most difficult to measure reliably. Inventories of hard-to-measure radionuclides often are estimated by applying scaling factors to measured amounts of photon emitters. However, scaling factors may not be appropriate for some important radionuclides, and their use may be problematic when a waste has undergone substantial processing after it is generated. 8.3.4

Unsaturated Zone Flow and Transport

Modeling of flow of water and transport of radionuclides in the unsaturated zone provides the link between releases from a disposal facility in water (the source term) and transport in an underlying zone of saturation (aquifer). Inputs to unsaturated zone flow and transport models include the infiltration rate through the cover on a disposal facility, flow rate through the facility, and

362 / 8. SUMMARY release rate of radionuclides from the facility. Outputs include the flow (recharge) rate and rate of radionuclide migration to an underlying aquifer. Of all the processes of concern to performance assessment, flow and radionuclide transport in the unsaturated zone are perhaps the most difficult to model. If substantial credit is taken for the ability of the unsaturated zone to provide a natural barrier to radionuclide transport, the approach to modeling is likely to be the most difficult to defend. There are several reasons why modeling of unsaturated zone flow and transport and defending the results are difficult. Flow in the unsaturated zone depends greatly on soil characteristics and moisture content and, thus, is highly site-specific, but site-specific data of sufficient detail are unlikely to be available. Relationships between hydraulic head (suction pressure), moisture content, and hydraulic conductivity used in modeling flow are complex and highly nonlinear. The nature of those relationships means, for example, that empirical approximations obtained from the literature may not represent flow at a given site and, furthermore, that an average flow over time may depend significantly on the detailed time history of precipitation, which is not normally taken into account in performance assessment. Radionuclide transport in the unsaturated zone normally is described in terms of a retardation factor, which is assumed to be linearly related to the ratio of the solid/solution distribution coefficient (Kd) to the moisture content of soil. However, that assumption has not been demonstrated experimentally, and there is some evidence that the retardation factor can be directly, rather than inversely, dependent on moisture content. Finally, if flow in fractured media occurs in the unsaturated zone, there is not even a generally accepted conceptual model to describe flow and transport, let alone agreement on approaches to mathematical modeling. Given the difficulties in modeling flow and transport in the unsaturated zone, a graded approach may be used. The simplest and most conservative approach is to ignore those processes by linking the output of the source-term model directly to models of flow and transport in the saturated zone, with no time delay or dilution. This approach is particularly appropriate at moist sites where a disposal facility lies within a few meters of the water table and the unsaturated zone may not provide a significant natural barrier. At arid sites, however, the thickness of the unsaturated zone may be hundreds of meters, and taking credit for delay and dilution in radionuclide transport in the unsaturated zone may be

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an important factor in demonstrating compliance with performance objectives. The simplest approach to describing unsaturated flow and transport is to assume a unit-gradient model, in which steady-state flow and a hydraulic conductivity equal to the infiltration rate are assumed. However, an infiltration rate would need to be chosen conservatively, and it may be difficult to justify use of a retardation factor in a transport analysis. If more complex methods are used, or if the unsaturated zone consists of substantial thicknesses of fractured media, considerable effort may be required to develop a defensible model of flow and transport. 8.3.5

Aquifer Flow

The purpose of an analysis of groundwater flow is to generate a velocity field, which generally depends on location and time. A velocity field provides input to an analysis of radionuclide transport in groundwater. The principal challenge in modeling aquifer flow is that groundwater velocities are not measurable in the field and are not directly deducible from other field measurements. Rather, a velocity field must be generated from a site-specific model of the flow system that is calibrated using site-specific data on hydraulic head obtained from monitoring wells and data obtained from pump or slug tests or tests of hydraulic conductivity in core samples. The fundamental difficulty with that approach is that neither the assumed model of the flow system nor the velocity field generated from it are unique, and a variety of assumptions generally fit available data reasonably well. Therefore, justification of a velocity field selected for use in performance assessment is an important issue, and it may be necessary to investigate the consequences of various assumptions that are reasonably compatible with available data. Use of Darcy’s Law to describe the velocity of a fluid in a porous medium is nearly universal in analyses of groundwater flow, and a general equation describing flow in the saturated zone is obtained by combining Darcy’s Law with the principle of conservation of mass. The flow equation, which is time dependent, usually is simplified by assuming steady-state flow, which should be appropriate over time scales of concern, and the equation is solved by invoking other simplifying assumptions (e.g., that recharge to the aquifer system is uniform over the spatial domain). Important issues that arise in using a simplified flow equation include the validity of a steady-state analysis developed on the basis of intrinsically transient data obtained from pump and slug tests, the dependence of hydraulic conductivity on the scale of a groundwater system and

364 / 8. SUMMARY the presence of heterogeneities in the system, uncertainties in determining appropriate boundary conditions for a flow system, the applicability (scaling) of data obtained from core samples to field conditions, incompatibilities of different sources of data (e.g., data from tests that may be greatly influenced by preferential flow paths versus data that characterize average conditions in a flow field), and the potential importance of flow in fractured media, which greatly complicates the modeling problem. There is no generally applicable approach to addressing uncertainties in modeling groundwater flow. Rather, uncertainties must be evaluated on a site- and analysis-specific basis. Uncertainties are best treated by an analysis of multiple conceptual models that are reasonably consistent with available data. 8.3.6

Radionuclide Transport in Groundwater and Surface Water

In modeling transport of radionuclides in groundwater, results of an analysis of the source term and an analysis of flow and transport in the unsaturated zone are combined with results of a groundwater flow analysis. Inputs to a transport analysis include a groundwater velocity field and radionuclide-specific release rates into the aquifer. Since at least some performance objectives for low-level waste disposal facilities are expressed in terms of annual dose to individuals, the desired outputs are concentrations of radionuclides in water, either groundwater or surface water into which groundwater may discharge, as a function of location and time. Phenomena that influence transport of radionuclides in groundwater include sorption, advection, diffusion, dispersion and radioactive decay. Advection generally is modeled by assuming that transport of radionuclides in solution is described by a groundwater velocity field. Radioactive decay is well understood, but approximations may be required in treating decay and ingrowth when modeling transport of decay chains containing many radionuclides. Sorption generally is described in performance assessment using an equilibrium solid/solution distribution coefficient (Kd), which is assumed to be independent of concentrations of radionuclides in a soil/water system. The distribution coefficient is used to derive a retardation factor, which represents the water velocity relative to the velocity of a radionuclide. The primary issue for performance assessment is the need to justify assumed values of Kd. Site-specific data may not be available, and determinations of distribution coefficients under field conditions are rare. Therefore, generic data often are used, including data sets in which

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measurements are categorized according to soil type (e.g., sand, clay, loam, organic). Difficulties in justifying Kds may be addressed by selecting conservative values. However, the lowest plausible Kd does not always result in the highest estimate of dose, especially when a radionuclide has radiologically significant long-lived decay products whose activity increases over a long time period (e.g., 238 U). Diffusion and dispersion affect transport of dissolved contaminants in groundwater. Absent radioactive decay, these phenomena generally serve to reduce concentrations compared with assuming advective transport only. For radionuclides with a half-life less than the travel time to an assumed receptor location, however, longitudinal dispersion can increase the peak concentration at that location. Diffusion generally is represented by Fick’s First Law, which states that the diffusive flux is linearly proportional to the concentration gradient. Dispersion is commonly assumed to be proportional to the Darcy velocity multiplied by the concentration gradient. Other, non-Fickian formulations of dispersion have limited utility in performance assessment, because their use requires more extensive data than are typically available. Given the usual assumptions about diffusion and dispersion, the main challenge for performance assessment is the selection of diffusion and dispersion coefficients in a soil/water system. The uncertainty in a dispersion coefficient may be many orders of magnitude. An additional complication is that dispersion coefficients may depend on the spatial scale of a system. The importance of diffusion and dispersion generally will be site-specific, but most analyses will need to be based on generic data. Modeling of radionuclide transport in surface water is not often considered in performance assessments of low-level waste disposal facilities, but surface-water transport could be important at some sites (e.g., when significant surface runoff of radionuclides could occur or groundwater discharges to surface water at locations close to a facility) and may also need to be considered in conducting an ALARA analysis. At many sites, it should be acceptable to assume that discharges of radionuclides are uniformly mixed in a surface-water body (river, stream or lake), and that attachment of radionuclides to sediments is unimportant. Solutions of an advective-diffusion equation at steady-state and methods of modeling transport of radionuclides attached to sediments are available in the literature if greater sophistication is needed.

366 / 8. SUMMARY 8.3.7

Atmospheric Transport

Release of radionuclides to the atmosphere and transport in air to the accessible environment usually are considered to be of secondary importance in performance assessments of near-surface waste disposal facilities, compared with release and transport in groundwater. However, atmospheric releases of radionuclides in gaseous form (e.g., 3H, 14C, 129I, and radon) could be important at any site. Atmospheric releases of radionuclides in particulate form also may be a concern in analyses of some disposal systems and in assessments of dose due to radionuclides deposited on the ground surface. Suspension of particulates may be particularly important in assessments of potential exposures of inadvertent intruders. Available models and databases to describe atmospheric releases of radionuclides from a disposal facility or the ground surface in either gaseous or particulate form are largely generic. However, this is not normally an important deficiency in performance assessment, because releases, if they occur, should be chronic and compliance with applicable performance objectives for chronic releases to the atmosphere often can be demonstrated using simple and conservative models. Models of atmospheric transport under conditions of chronic releases generally are well established, especially compared with models of transport in groundwater, on the basis of comparisons of model predictions with environmental data. 8.3.8

Biotic Transport

The term “biotic transport” refers to actions of plants or animals that serve to transport waste materials from a disposal site to the accessible environment. Although numerous studies have shown that biotic transport has occurred at historic low-level waste disposal sites, those processes normally are considered to be of secondary importance in performance assessment, due mainly to a lack of relevant site-specific data. Most contemporary designs of low-level waste disposal facilities include a cover or other engineered barriers that are intended to minimize biotic intrusion and transport. Nonetheless, at least a qualitative discussion of the potential importance of biotic transport should be included in performance assessments. Three classes of biotic transport mechanisms have been identified and are referred to as transport enhancement, intrusion and active transport, and secondary transport. Transport enhancement occurs when plants or animals intrude into a disposal facility in a manner that increases the potential for transport of radionuclides

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in water or air (e.g., by creating preferential flow paths through a cover system). Intrusion and active transport occurs when plants or animals penetrate into waste and cause a redistribution of waste material (e.g., transport by root uptake in plants). Secondary transport occurs when radionuclides are available for additional displacement by biota after they have been mobilized by other processes (e.g., redistribution by plant leaves or fruit following root uptake). As yet, there are no generally accepted approaches to modeling various transport processes of concern. In performance assessments that have considered biotic transport, the usual approach has been to perform simple screening-level calculations in an effort to show that biotic transport is unlikely to be significant. 8.3.9

Exposure Pathways and Radiological Impacts

Models of exposure pathways and radiological impacts are used to convert estimates of concentrations of radionuclides in environmental media at assumed receptor locations, as obtained from models of transport in water or the atmosphere, to estimates of exposure of humans and dose. Of all the components of performance assessment, modeling of exposure pathways is unique in that models and data sets can be based directly on relevant environmental measurements. Standard data sets to convert estimates of exposure to dose (or risk, if so desired) also are well established. Thus, modeling of exposure pathways and radiological impacts is largely free from challenges in conducting other parts of a performance assessment that often arise from a lack of knowledge about appropriate models and relevant data. Analyses of exposure pathways involve the formulation of exposure scenarios, selection of relevant exposure pathways, and selection of data sets used in the pathway models. Analyses generally should be site-specific. However, in contrast to other components of a performance assessment, suitable scenarios, models and data sets should not differ greatly from one site to another. As a result, estimates of dose per unit concentration of radionuclides in the environment should not be highly site-specific. The main concern in performance assessment normally is the need to justify assumed exposure scenarios and associated exposure pathways. Once this is accomplished, use of generic data sets in modeling exposure pathways and standard sets of dosimetry data for reference adults should be acceptable in most cases. Given that largely generic scenarios, models and data sets normally should be acceptable for use in site-specific performance assessments with a minimum of justification, analyses of exposure

368 / 8. SUMMARY pathways and radiological impacts should be the least important component of performance assessment in regard to the amount of attention they warrant in reviews, essentially because this component is not concerned with the central issue of evaluating the capabilities of a disposal facility and site to inhibit releases of radionuclides to the accessible environment. It is that issue that warrants the greatest attention in the process of site selection and facility design. 8.3.10 Overview of Components of Performance Assessment An important theme in this Report is that there are potentially important uncertainties associated with formulating conceptual models for all components of performance assessment except analyses of atmospheric transport and exposure pathways and radiological impacts. In many respects, the future behavior of a disposal system is not well understood at a fundamental level. In many components of performance assessment, standard approaches to modeling phenomena of concern have been developed, and there may be widely accepted data sets. Nonetheless, it often is difficult, if not impossible, to determine if a conceptual model and its implementation by means of available mathematical models and data sets provide a reasonable representation of the behavior of a disposal system for purposes of performance assessment. Difficulties that arise from a basic lack of knowledge of many aspects of the future behavior of disposal systems are an important reason why reviewers of performance assessments usually pay considerably greater attention to conceptual models that provide the basis for an assessment and to an integration and interpretation of results of an analysis than to mathematical models used in an assessment and their implementation. Uncertainties in formulating conceptual models are alleviated, in part, by recognizing that the primary purpose of performance assessment is to provide a means of demonstrating compliance of a disposal facility with applicable performance objectives. Two related considerations then become important in conducting performance assessments. First, the approach to performance assessment should be intentionally conservative in some respects, although highly implausible assumptions usually are not warranted except for purposes of screening of pathways and radionuclides and use of too many highly conservative assumptions should be avoided if such assumptions result in overly restrictive requirements on acceptable waste disposals. An evaluation of conservative bias is likely to be an important part of any assessment. Second, a

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variety of conceptual models (multiple lines of reasoning) that are broadly consistent with existing knowledge should be considered in an assessment. Without investigating the consequences of different conceptual models, it is difficult to determine that a full range of plausible outcomes has been captured in an assessment and that an analysis that is used in making the case for compliance with performance objectives is conservatively biased. 8.4 Inadvertent Human Intrusion Analyses of inadvertent human intrusion are an important part of performance assessment at near-surface facilities for disposal of low-level waste. However, that type of analysis addresses a different issue than the components of performance assessment summarized above. For this reason, inadvertent intrusion is treated separately in this Report. Components of performance assessment discussed in the previous section are concerned essentially with the issue of determining limits on acceptable disposals of radionuclides that would ensure adequate protection of individual members of the public who reside beyond the boundary of a disposal site. Thus, the focus is on release of radionuclides from the disposal facility and transport to the accessible environment. In contrast, analyses of inadvertent intrusion are concerned with the issue of determining limits on acceptable disposals that would ensure adequate protection of individuals who might be exposed to waste remaining in a disposal facility after loss of institutional control. Thus, the essential role of analyses of inadvertent intrusion is to determine limits on acceptable near-surface disposals of low-level waste without regard for the capabilities of a disposal system to inhibit release of radionuclides to the accessible environment. Limits on acceptable disposals based on analyses of inadvertent intrusion may be applied uniformly to all disposal sites (NRC, 1982b), or they may be developed on a site-specific basis (DOE, 1988a; 1999b). The importance of analyses of inadvertent intrusion is evidenced by the fact that they often determine limits on acceptable near-surface disposals of most radionuclides, even at sites with abundant precipitation. At most sites, disposal limits are determined on the basis of analyses of release and transport only for relatively mobile and longer-lived radionuclides. Issues that arise in analyses of inadvertent human intrusion are essentially the same as those that arise in analyses of exposure pathways and radiological impacts following releases to the accessible environment. Thus, the primary focus is on formulation

370 / 8. SUMMARY of scenarios for inadvertent intrusion into disposed waste, selection of relevant exposure pathways, and selection of data sets to implement exposure pathway models. Largely generic scenarios and models, such as used by NRC in developing the waste classification system in 10 CFR Part 61 (NRC, 1981a; 1982a; 1982b), should be acceptable at most disposal sites. Nonetheless, consideration should be given to the effects of specific site characteristics, facility designs, waste forms, and intended waste emplacements on the development of plausible scenarios. When these factors are considered, credible scenarios for inadvertent intrusion can be highly site-specific.

8.5 Uncertainty, Sensitivity and Importance Analysis Uncertainty is an inherent aspect of performance assessment, because the future behavior of disposal systems is largely unknown and cannot be tested over time periods of concern. Therefore, an appropriate treatment of uncertainties in models and databases used in an assessment and the sensitivity of results to changes in particular assumptions is an essential part of an assessment. Only in this way can decision makers be provided with adequate information on what is known and unknown about a proposed disposal practice. Appropriate considerations of uncertainty are greatly influenced by the primary purpose of performance assessment. An analysis of uncertainty need not be unduly concerned with estimating uncertainty in projected radiological impacts. Rather, the more relevant concern is the uncertainty in a decision regarding compliance with regulatory performance objectives, which is made on the basis of results of performance assessment. This emphasis leads to a more tractable problem, because many important sources of uncertainty in predicting actual impacts at future times are largely unquantifiable, except in highly subjective terms, but they sometimes are not important to a regulatory decision. To emphasize the difference between the type of uncertainty analysis appropriate to performance assessment and more traditional approaches to quantitative uncertainty analysis of models, this Report has used the term “importance analysis” to describe the former. This distinction is important because, in spite of the seeming contradiction, there can be a high degree of assurance of compliance with performance objectives when, at the same time, there is a high degree of uncertainty in the actual performance of a disposal facility.

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Importance analysis can use tools of traditional uncertainty and sensitivity analysis. However, importance analysis differs from traditional methods that are concerned with uncertainties in projected outcomes in that it focuses on identifying assumptions about uncertain models and parameter values which, when varied over a range of plausible conditions, could affect a decision about regulatory compliance. The decision maker then must judge whether assumptions that result in outcomes that exceed performance objectives are sufficiently credible to warrant design changes, additional data collection, and refinements in models, or perhaps rejection of a site, or whether those assumptions are so implausible that a site can be licensed. It rarely, if ever, would be the case that there would be no assumptions within the realm of possibility that would result in projected outcomes that exceed performance objectives.

8.6 Iterative Nature of Performance Assessment A single performance assessment that is based on available site characterization data and a conceptual design of a disposal facility generally will not suffice in making the case for licensing, mainly because of the paucity of relevant information at initial stages of facility development. Experience has shown that performance assessment should be viewed as an iterative process that involves substantial interaction between model development and datacollection and design activities throughout the period prior to licensing. Site data and design information provide the basis for assumptions about models and parameter values, evaluation of modeling results can lead to an identification of additional data needs or design changes, and these in turn can lead to different assumptions and parameter values in modeling. Sensitivity analysis plays an important role in identifying parameters and design features for which further investigation is needed in subsequent iterations. These interactions can result in multiple iterations of an assessment. An iterative approach to performance assessment almost certainly will be dictated by results of technical peer reviews and reviews by regulatory authorities during the licensing process. Experience has shown that technical peers and regulators invariably raise concerns that can only be addressed by further analysis, and addressing those concerns is important in enhancing the credibility of an analysis for the purpose of regulatory decision making.

372 / 8. SUMMARY 8.7 Reasonable Assurance of Compliance with Performance Objectives The conduct of performance assessment has many qualitative and subjective aspects, even though mathematical models are used extensively and analyses generally produce quantitative results. The development of conceptual models of disposal systems and an integration and interpretation of results of an assessment with respect to the issue of compliance with performance objectives are basically qualitative and subjective, and there are many important uncertainties in performance assessment models that can only be quantified subjectively. Given the nature of performance assessment, a licensing decision itself is necessarily subjective. The decision process can be greatly aided by an importance analysis, but it cannot be reduced to a simple comparison of quantitative results with regulatory performance objectives. Rather, the standard of proof that a facility complies with performance objectives must be one of “reasonable assurance.” This concept acknowledges that proof of compliance is not to be had in the ordinary sense of the word, and it is based on consideration of the nature of uncertainties in evaluating the long-term performance of disposal systems. There may always be a residual level of doubt about the ability of any facility to comply with performance objectives, unless allowable disposals of radionuclides are severely limited. If, based on all information (both quantitative and qualitative) that has been assembled concerning the long-term performance of a disposal facility, regulatory authorities reach a judgment that there is reasonable assurance of compliance with performance objectives and any other regulatory requirements (or that a site should be rejected), the process of performance assessment for the purpose of qualifying a site to dispose of waste is complete. However, further performance assessment activities likely will continue during the period of waste emplacement, in order to support a decision that a facility can be closed, and during the institutional control period after closure. Design changes may be incorporated into a facility that require a revision of an assessment, or site monitoring or research activities may provide new information on the waste-isolation capabilities of a disposal system. A final performance assessment presumably will be required before a disposal site can be released from institutional control. Thus, performance assessment will be important in all phases of the life cycle of a disposal facility from initial siting and design activities through postclosure institutional control.

Glossary absorbed dose (D): Quotient of d∈ by dm, where d∈ is the mean energy imparted by ionizing radiation to matter in a volume element and dm is the mass of matter in that volume element: D = d∈/dm. For purposes of radiation protection and assessing dose or risk to humans in general terms, the quantity normally calculated is the mean absorbed dose in an organ or tissue (T): DT = ∈T/mT, where ∈ is the total energy imparted in an organ or tissue of mass mT. The SI unit of absorbed dose is the joule per kilogram (J kg–1), and its special name is the gray (Gy). In conventional units often used by federal and state agencies, absorbed dose is given in rad; 1 rad = 0.01 Gy. accessible environment: The atmosphere, land surfaces, surface waters, oceans, and all of the lithosphere that is beyond a controlled area and is generally accessible to the public. accuracy: As applied to models, agreement between a model prediction and actual occurrences. An accurate model should be precise and unbiased (see bias and precision). activation: Production of radionuclides by absorption of radiations (e.g., photons, neutrons or alpha particles) by atomic nuclei. activity: Rate of transformation (or “disintegration” or “decay”) of radioactive material. The SI unit of activity is the reciprocal second (s–1), and its special name is the becquerel (Bq). In conventional units often used by federal and state agencies, activity is given in curies (Ci); 1 Ci = 3.7 × 1010 Bq. adsorption: (see sorption). advection: Movement of material due to bulk flow of a medium (e.g., air or water) in which material is dissolved or suspended. Agreement State: Any state with which the U.S. Nuclear Regulatory Commission (NRC) has entered into an effective licensing agreement under Section 274(b) of the Atomic Energy Act of 1954, as amended, to enable the state to regulate source, special nuclear, and byproduct materials. alpha radiation: Energetic nuclei of helium atoms, consisting of two protons and two neutrons, emitted spontaneously from nuclei in decay of some radionuclides. Also called alpha particle and sometimes shortened to alpha (e.g., alpha-emitting radionuclide). amorphous: Lacking a definite crystalline form; not crystalline. anaerobic: Living, acting or occurring in the absence of free oxygen. analytical solution: Solution of a differential or integral equation that can be expressed in terms of known mathematical functions, either in closed form or as an infinite series. anion: Negatively charged ion.

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374 / GLOSSARY annual dose equivalent: Sum of the dose equivalent received in a year from external radiation and the committed dose equivalent due to intakes of radionuclides in a year. A similar definition applies to annual effective dose, annual equivalent dose, and annual effective dose equivalent. aquifer: Saturated geologic unit that is sufficiently permeable to transmit usable quantities of water under ordinary hydraulic gradients. In an unconfined aquifer, there is no restricting, less permeable material at the top of groundwater (i.e., the water table), and groundwater levels are free to rise or fall. In a confined aquifer, a water-bearing unit is confined between two aquitards (also see groundwater). aquitard: Saturated geologic unit that is sufficiently impermeable that it does not transmit usable quantities of water under ordinary hydraulic gradients. as low as reasonably achievable (ALARA): Principle of radiation protection that calls for every reasonable effort to maintain radiation exposures as far below specified limits as is practical, taking into account cost-benefit and any other societal concerns. Atomic Energy Act (AEA): Law passed originally in 1946 and extensively revised in 1954 that governs production and use of radioactive materials (i.e., byproduct material, source material, and special nuclear material) for defense and peaceful purposes and regulation of such radioactive materials to protect public health and safety. Act provides authority for licensing of nuclear activities by the U.S. Nuclear Regulatory Commission or Agreement States and regulation by the U.S. Department of Energy of its atomic energy defense, research and development activities. autocatalysis: Speeding up (or sometimes slowing down) of the rate of a chemical reaction by the presence of a substance that itself undergoes no permanent chemical change as a result of the reaction. background radiation: Ionizing radiation that occurs naturally in the environment, including cosmic radiation, radiation emitted by naturally occurring radionuclides in air, water, soil and rock, radiation emitted by naturally occurring radionuclides in tissues of organisms (e.g., due to ingestion or inhalation), radiation emitted by man-made materials containing incidental amounts of naturally occurring radionuclides (e.g., building materials), and radiation emitted by widespread fallout from atmospheric testing of nuclear weapons. In the United States, the average annual effective dose due to natural background radiation is ~1 mSv, excluding the dose due to indoor radon, and the average annual effective dose due to indoor radon is ~2 mSv. barrier: Any natural or man-made part of a disposal system that is intended to inhibit access to disposed waste by humans, plants, animals, air or water, or to inhibit migration of radionuclides from locations of disposed waste to the accessible environment. Bayesian probability theory: Method of estimating uncertainty on the basis of probabilities that represent subjective degrees of belief or inferences, rather than purely objective facts (direct observations).

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becquerel (Bq): Special name for the SI unit of activity; 1 Bq = 1 s–1. benchmarking: Intercomparison of outputs of different models or codes that purport to solve the same problem. beta radiation Energetic electrons or positrons (positively charged electrons) emitted spontaneously from nuclei in decay of some radionuclides. Also called beta particle and sometimes shortened to beta (e.g., beta-emitting radionuclide). bias: Tendency for an estimate to over- or under-predict an actual event (see also conservative bias). bioaccumulation factor: Ratio of the concentration of a radionuclide in a biological organism or tissue to the concentration in water or soil. biokinetic model: Model describing the time course of absorption, distribution, metabolism and excretion of a substance introduced into the body of an organism. biosphere: Zone of the Earth, including the lower part of the atmosphere, the hydrosphere, soil, and the lithosphere to a depth of ~2 km, that contains living organisms. boundary condition: Assumed value of a dependent variable or one of its spatial derivatives at boundaries defining the domain over which a differential equation is solved. byproduct material: (1) Any radioactive material (except special nuclear material) yielded in, or made radioactive by, exposure to radiation incident to the process of producing or utilizing special nuclear material, and (2) tailings or waste produced by extraction or concentration of uranium or thorium from any ore processed primarily for its source material content. Byproduct material does not include ore bodies depleted by uranium solution extraction operations and which remain underground. The Energy Policy Act of 2005 amended the definition of byproduct material to include (1) any radioactive material produced in an accelerator, (2) discrete sources of 226Ra, and (3) any discrete source of naturally occurring radioactive material, other than source material, that NRC determines would pose a threat similar to that posed by a discrete source of 226Ra. NRC has not yet developed regulations to address control of these materials. cap: (see cover). capillary fringe: Zone immediately above the water table in which water is held by surface tension and is under a pressure less than atmospheric. In the capillary fringe of a porous medium, the volumetric water content is constant. cation: Positively charged ion. cement: Substance capable of making objects adhere to each other. In construction, cement is a burned and finely pulverized substance containing alumina, silica, lime, iron oxide, and magnesia that is used to form concrete when mixed with water, sand and aggregate. characteristic curve: In an unsaturated porous medium, relationship between moisture content and suction pressure, hydraulic conductivity and suction pressure, or hydraulic conductivity and moisture content; also referred to as a moisture characteristic curve.

376 / GLOSSARY chelate: Chemical compound in which the central atom (usually a metal ion) is attached to neighboring atoms by at least two bonds in such a way as to form a ring structure. chemisorption: Adsorption in which forces involved are valence forces of the same kind as those operating in the formation of chemical compounds; also referred to as chemical adsorption. colloid: Small, insoluble and nondiffusible particle (as a single large molecule or mass of smaller molecules) in solid, liquid or gaseous form that remains in suspension in a surrounding solid, liquid or gaseous medium of different matter. commercial waste: Waste generated in any activity by a nongovernmental entity. Often refers to waste containing source, special nuclear, or byproduct material regulated by the U.S. Nuclear Regulatory Commission or an Agreement State, but also may refer to waste containing naturally occurring and accelerator-produced radioactive material (NARM) that is currently regulated only by the states. committed dose: Dose (absorbed dose, dose equivalent, effective dose, effective dose equivalent, or equivalent dose) delivered to specified organs or tissues over a specified period of time following an acute intake of a radionuclide by ingestion, inhalation or dermal absorption. Time period over which committed doses are calculated normally is 50 y for intakes by adults or from age at intake to age 70 for intakes by other age groups. Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA or “Superfund”): Law passed in 1980, and amended by Superfund Amendments and Reauthorization Act (SARA) of 1986 and later amendments, that governs federal response and compensation for unpermitted and uncontrolled releases, including threats of release, of hazardous substances, including radionuclides, to the environment. An “unpermitted” release is any release that is not properly regulated under other laws. An important focus of CERCLA/ SARA is remediation of old, unpermitted waste disposal sites that are closed or inactive. Objectives of the Superfund program are to protect human health and the environment in a cost-effective manner, maintain this protection over time, and minimize amounts of untreated waste in the environment. compressive strength: Compressive force per unit area that a material can withstand without significant change in properties. concentration factor, concentration ratio: (see bioaccumulation factor). conceptual model: Set of assumptions and qualitative descriptions about the behavior of a system that provides the basis for subsequent quantitative descriptions of the system and interpretations of results of mathematical or physical modeling. concrete: Material formed by coalescence of particles into one solid mass. In construction, concrete is a material made by mixing cement with water, sand and aggregate.

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conservative bias: Tendency of a model to overestimate, rather than underestimate, actual events. containment: Confinement of material within a designated boundary. controlled area: Surface location, including the location of a waste disposal facility, identified by active or passive institutional control that is intended to be used for monitoring and surveillance of a disposal facility and to restrict or discourage public access, and the subsurface underlying such a surface location. convection: Movement of material within a fluid at a nonuniform temperature due to the variation in its density and the action of gravity. coordination complex: Compound in which an atom or group of atoms is bound to the central atom by a shared pair of electrons supplied by the coordinated group and not by the central atom. corrosion: Wearing away of a material or degradation of material properties by the action of chemicals. cover: Layer of soil or other earthen or man-made materials installed above waste upon closure of a disposal facility. creep: Change in shape as a result of constant stress, temperature, or other external influences. critical group: Subgroup of an exposed or potentially exposed population that receives or is expected to receive the highest dose or experience the highest risk due to exposure. critical organ: Organ receiving the highest dose or experiencing the highest risk from exposure to radiation or radionuclides. curie (Ci): (see activity). Darcy velocity: Product of the hydraulic conductivity of a porous medium and the gradient in pressure head. defense waste: Radioactive waste from any activity performed in whole or in part in support of U.S. Department of Energy (DOE) atomic energy defense activities. Not all DOE waste is defense waste (e.g., waste from research and development activities not related to atomic energy defense). depassivation: Removal of protective coating on the surface of a material, resulting in an increase in chemical activity with surrounding materials. deposition velocity: Ratio of flux density of a contaminant from the atmosphere to the ground surface to the concentration in air above ground. desorption: Any mechanism that removes ions from the solid phase of a medium into the fluid phase (see sorption). deterministic effects: Effects in organisms for which the severity varies with the dose of radiation (or other toxic substance), and for which a threshold usually exists. deterministic methods: Methods of exercising mathematical models in which a single set of assumptions (i.e., scenario, model, and model input parameters) is used to calculate a single value of model output. detriment: Measure of stochastic effects from exposure to ionizing radiation that takes into account the probability of fatal cancers, probability

378 / GLOSSARY of severe hereditary effects in future generations, probability of nonfatal cancers weighted by the lethality fraction, and relative years of life lost per fatal health effect (ICRP, 1991). differential analysis: Method of parameter sensitivity analysis of a model in which a model solution near a point in the domain of parameter values is approximated using lower-order derivatives of the solution at that point. diffusion: Spreading out of a material in a medium, due to thermal or mechanical agitation, in response to a concentration gradient. diffusion coefficient: Measure of rate of diffusion of a material in a medium, given by the ratio of the diffusive flux of material crossing a defined boundary to the (negative) concentration gradient. diffusivity: Hydraulic conductivity divided by differential water capacity, or flux of water per unit gradient of water content in the absence of other force fields. dispersion: Spreading of a flowing substance in a medium due to random variations in the structure of the medium or random variations in the speed and direction of flow. dispersion coefficient: Measure of rate of dispersion of material in flow in a medium, given by the dispersive flux divided by the negative of the concentration gradient. dispersivity: Measure of dispersion characteristics of a porous medium defined as the ratio of the dispersion coefficient to the fluid velocity in the medium. disposal: Placement of waste in a facility designed to isolate waste from the accessible environment without an intention to retrieve the waste, irrespective of whether such isolation permits recovery of waste. disposal cell: (see disposal unit). disposal facility: Land, structures and equipment used for disposal of waste. disposal, geologic: Isolation of waste using a system of engineered and natural barriers at a depth of up to several hundred meters below ground in a geologically stable formation (see geologic repository). disposal, near-surface: Disposal of waste, with or without engineered barriers, on or below the ground surface, such that the final protective cover above the waste is on the order of a few meters thick, or in mined openings within a few tens of meters of the Earth’s surface (see land disposal facility). disposal site: Natural setting at the location of a disposal facility. disposal system: Disposal facility plus surrounding land in a controlled area. disposal unit: Discrete portion of a disposal facility into which waste is emplaced (e.g., single trench, vault, bunker or tumulus). distribution coefficient: Quantity of a contaminant sorbed by solid material per unit mass of the solid divided by the quantity of the contaminant dissolved in a fluid per unit volume of fluid. dose: General term used to denote mean absorbed dose, equivalent dose, effective dose, or effective equivalent dose, and to denote

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dose received or committed dose. Particular meaning of the term should be clear from context in which it is used. dose coefficient: (1) For ingestion or inhalation of radionuclides, committed dose per unit activity intake; or (2) for external exposure to radionuclides in the environment, dose rate per unit concentration in an environmental medium. dose commitment: (see committed dose). dose equivalent (H): Absorbed dose (D) at a point in tissue weighted by quality factor (Q) for type and energy of the radiation causing the dose: H = D × Q. For purposes of radiation protection and assessing health risks in general terms, and especially prior to introduction of the equivalent dose and as used by federal and state agencies, dose equivalent often refers to mean absorbed dose in an organ or tissue (T) weighted by average quality factor ( Q ) for the particular type of radiation: HT = DT × Q . The SI unit of dose equivalent is the joule per kilogram (J kg–1), and its special name is the sievert (Sv). In conventional units often used by federal and state agencies, dose equivalent is given in rem; 1 rem = 0.01 Sv. dose rate: Dose per unit time; often expressed as an average over some time period (e.g., a year). dosimetric model: (1) For intakes of radionuclides into the body, model that estimates the dose in various organs and tissues per disintegration of a radionuclide in a specified source organ (site of deposition or transit in the body). (2) For external exposure, model that estimates the dose rate in organs and tissues per unit activity concentration of a radionuclide in an environmental medium. dynamic: (1) Time-dependent or (2) in motion. effective dose (E): Sum over specified organs and tissues (T) of equivalent dose in each tissue weighted by tissue weighting factor (wT): E = ΣwT × HT, where ΣwT ≡ 1 (ICRP, 1991) (supersedes effective dose equivalent). effective dose equivalent (HE): Sum over specified organs and tissues (T) of mean dose equivalent in each tissue weighted by tissue weighting factor (wT): HE = ΣwT × HT, where ΣwT ≡ 1 (ICRP, 1977) (now superseded by effective dose, but often used by federal and state agencies). effective porosity: Porosity of that part of a medium through which water flows. Eh: Symbol for oxidation/reduction potential of a medium. element: Any substance that cannot be separated into different substances by ordinary chemical methods. Elements are distinguished by the number of protons in the nucleus of atoms. embrittlement: Increase in hardness or decrease in flexibility of a material, resulting in an increased tendency to break or shatter. encapsulation: Incorporation of waste into a solid waste form (e.g., grout, bitumen or polyethylene). enhancement factor: Ratio of activity of a radionuclide per unit mass of particulate material in air above ground to activity of the radionuclide

380 / GLOSSARY per unit mass of surface soil that provides the source of suspended or resuspended material. environment: Soil, rock, water, atmosphere and biosphere surrounding a waste disposal facility into which radionuclides may be released and transported. epoxy: Resin, containing groups of oxygen atoms joined to two carbon atoms to form a bridge, that polymerizes spontaneously when mixed with a diphenol, forming a strong, hard and chemically resistant adhesive. equilibrium: Condition of equal rates of processes or equal ratios of values of parameters. equivalent dose (H): Quantity developed for purposes of radiation protection and assessing health risks in general terms, defined as mean absorbed dose in an organ or tissue (T) weighted by radiation weighting factor (wR) for type and energy of the radiation causing the dose: HT = DT × wR (ICRP, 1991). The SI unit of equivalent dose is the joule per kilogram (J kg–1), and its special name is the sievert (Sv). In conventional units often used by federal and state agencies, equivalent dose is given in rem; 1 rem = 0.01 Sv (see dose equivalent). error: Difference between a computed or estimated result and the actual value. error function: Integral over normal or Gaussian probability function defined as: z

2 –t 2 erf ( z ) = ------- ∫ e dt . π 0

(G.1)

Error function gives the probability that the value of the parameter t does not exceed z. evaporation: Removal of water from the Earth’s surface into the atmosphere by vaporization. exfiltration: Process of upward movement of materials buried in the Earth’s surface into the atmosphere. Exfiltration normally occurs for gaseous materials but may also occur for liquids in arid environments. exposure: General term indicating human contact with radiation or radionuclides. external exposure: Exposure to radiation originating from a source outside the body. extrapolation: Use of a data set or model under conditions different from those for which it was established. fault tree analysis: Systematic approach used in analyzing the reliability of complex systems in which probabilities of failure of individual components of a system and resulting chains of cause-effect consequences within the system are estimated. finite-difference method: Method of solving systems of partial differential equations in which differentials are approximated by finite differences, resulting in a series of algebraic equations.

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finite-element method: Method of solving systems of partial differential equations in which independent variables are discretized, which creates an integral form of the differential equations and provides a system of linear algebraic equations. fission product: Atom, either stable or radioactive, produced by splitting apart of an atomic nucleus, either spontaneously or when induced by absorption of a neutron. flux: Volume of discharge from a given cross-sectional area per unit time. flux density: Flux per unit cross-sectional area. fuel reprocessing: Separation and extraction of various chemical elements in spent nuclear fuel. fuzzy set: Generalization of the concept of set; in particular, a subset F of a given set S for which the characteristic function χF:S → [0,1] has the entire interval as its range, rather than just the two values {0,1}. fuzzy set analysis: Application of the concept of fuzzy set for estimating the degree of certainty in conclusions. gamma radiation: Electromagnetic radiation emitted in de-excitation of atomic nuclei, and frequently occurring in decay of radionuclides. Also called gamma ray and sometimes shortened to gamma (e.g., gammaemitting radionuclide) (see photon and x ray). Gaussian plume model: Pollutant diffusion model that is based on an assumption of stationary, homogeneous turbulent flow. Distribution of material in a plume is assumed to be Gaussian in shape. generic: Of, applied to, or referring to a whole kind, class or group; not site-specific. Often refers to a collection of data representing measurements made under a variety of conditions, or general assumptions intended to be broadly applicable to any site. geologic repository: System intended for disposal of radioactive waste in excavated geologic media; includes a subterranean mined facility for disposal of waste and portion of the geologic setting that provides a barrier to movement of radionuclides in waste. groundwater: Water below the land surface in a zone of saturation that is under a pressure equal to or greater than atmospheric pressure. grout: Any concrete-like material used to encapsulate waste or provide fill in waste disposal units. half-life (T1/2): Time over which half the atoms of a particular radionuclide decay to another nuclear form. hazard: Act or phenomenon that has the potential to produce harm or other undesirable consequences to humans (e.g., ionizing radiation). heuristic: Proceeding along empirical lines, using rules of thumb, to find solutions or answers; not rigorous. high-level radioactive waste: (A) Highly radioactive material resulting from reprocessing of spent nuclear fuel, including liquid waste produced directly in reprocessing and any solid material derived from such waste that contains fission products in sufficient concentrations, and (B) other highly radioactive material that the U.S. Nuclear Regulatory Commission, consistent with existing law, determines by rule requires permanent isolation. In most countries other than the United

382 / GLOSSARY States, high-level waste also includes waste from any source that contains high concentrations of shorter-lived radionuclides and high concentrations of long-lived, alpha-emitting radionuclides. At the present time, however, high-level waste in the United States includes only waste produced directly in chemical reprocessing of spent nuclear fuel. homolog: Chemical element belonging to same group of the periodic table as another element (e.g., halogens, actinides). humic: Of or derived from organic materials (partially decayed plant or animal matter) in soil. hydraulic conductivity: Volume of water that will move per unit time under a unit hydraulic gradient through a unit cross-sectional area perpendicular to direction of flow. hydraulic head: (see pressure head). hysteresis: Dependence of the state of a system on its previous history, generally a retardation or lagging of an effect behind the cause of the effect. importance analysis: Analysis of performance assessment models for the purpose of identifying assumptions and parameter values which, when changed within credible bounds, can affect a decision about compliance of a waste disposal facility with applicable regulatory performance objectives. inadvertent intruder: Hypothetical individual who might occupy a waste disposal site after facility closure and engage in normal activities, such as agriculture, dwelling construction, permanent residence, or other pursuits, that might result in the individual being unknowingly exposed to waste materials. infiltration: Process of downward movement of water from the ground surface into underlying materials. initial condition: Assumed value of a dependent variable or one of its time derivatives at the time the solution of a differential equation or set of equations is begun. institutional control: Control of a waste disposal site by an authority or institution designated under laws of a country, state or local authority. Institutional control may be active (e.g., monitoring of effluents, surveillance, remedial activities, fences, or guards) or passive (e.g., records or warning signs). integration: Process of presenting results of a performance assessment and important assumptions that determined the results in a manner that provides the basis for identifying conditions under which a disposal facility is expected to comply with applicable regulatory performance objectives. interception fraction: Fraction of material deposited from the atmosphere onto the ground surface that is intercepted and immediately retained on surfaces of vegetation. intermediate-level radioactive waste: Class of radioactive waste defined and used in many countries other than the United States that

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contains concentrations of shorter-lived and long-lived radionuclides intermediate between those in low-level and high-level wastes. internal exposure: Exposure to radiation originating from a source within the body (e.g., as a result of intakes of radionuclides into the body by inhalation or ingestion). intrinsic permeability: Property of a porous medium defined as the ratio of the product of the hydraulic conductivity and the absolute viscosity of the fluid to the product of the density of the fluid and the acceleration due to gravity. inventory: Quantity of a radionuclide placed in a waste package, disposal unit, or facility, usually expressed in terms of activity. ionizing radiation: Any radiation capable of displacing electrons from atoms or molecules, thereby producing ions. Examples include alpha radiation, beta radiation, gamma radiation or x rays, and cosmic rays. Minimum energy of ionizing radiation is a few electron volts (eV); 1 eV = 1.6 × 10–19 J. isolation: Disposal of waste in a manner that is expected to provide adequate protection of human health and the environment. isomorphic: Having identical or similar structure or form. isotope: Form of a particular chemical element determined by the number of neutrons in the atomic nucleus. An element may have many stable or unstable (radioactive) isotopes. isotropic: Characteristic of a medium denoting that properties at any point within the medium are the same in all directions. kriging: Geostatistical interpolation method for predicting, without bias and with minimum variance, spatial distributions of properties of geologic systems on the basis of observations at discrete locations. land disposal facility: Land, buildings and equipment intended to be used for disposal of wastes in a subsurface facility located within a few tens of meters of the Earth’s surface or in an above-grade facility. A geologic repository is not considered a land disposal facility. Latin hypercube sampling method: Technique of stratified random sampling from specified probability distributions of variables in which probability distributions are divided into intervals of equal probability and one sample is taken at random from each interval (see MonteCarlo analysis). leaching: Extraction or removal of a soluble substance from a solid material by contact with water. ligand: Atom, group of atoms with similar chemical properties, ion, radical, or molecule that forms a coordination complex with a central atom or ion. linear energy transfer [LET (L)]: Quotient of dE by d , where dE is the energy lost by a charged particle in traversing a distance d in a material: L = dE/d . The SI unit of LET is the joule per meter (J m–1). For purposes of radiation protection, LET normally is specified in water and given in units of keV µm–1. low-level radioactive waste: Radioactive waste that (A) is not high-level radioactive waste, spent nuclear fuel, transuranic waste, or

384 / GLOSSARY byproduct material as defined in Section 11(e)(2) of the Atomic Energy Act, and (B) the U.S. Nuclear Regulatory Commission, consistent with existing law, classifies as low-level radioactive waste. Byproduct material referred to in Clause (A) essentially is uranium or thorium mill tailings. Low-level radioactive waste does not include waste containing naturally occurring and accelerator-produced radioactive material (NARM), although such waste may be managed as low-level waste. In most countries other than the United States, low-level waste includes only radioactive waste that contains the lowest concentrations of radionuclides, but low-level waste in the United States can contain high concentrations of shorter-lived radionuclides and high concentrations of long-lived radionuclides other than long-lived, alpha-emitting transuranium radionuclides. Low-Level Radioactive Waste Policy Act (LLRWPA): Law passed in 1980 and amended in 1985 that governs disposal of low-level radioactive waste by states or State Compacts. Act does not govern disposal of low-level waste generated at U.S. Department of Energy sites unless such waste is sent to a disposal facility established under the Act. All waste disposals under the Act will be licensed by the U.S. Nuclear Regulatory Commission or an Agreement State. lysimeter: Tank containing soil or other earthen materials, which is placed into the ground so that its surface is on the same level as that of the surrounding land, that may be used to measure infiltration of precipitation or leaching and transport of contaminants contained in a waste form placed inside the device. mass loading: Concentration of particulate material in air resulting from suspension or resuspension of surface soil. matric potential: (see suction pressure). maximally exposed individual: Individual assumed to receive the highest dose or to be at greatest risk from exposure to radiation. meat transfer coefficient (Ff): Fraction of the amount of an element ingested daily by an herbivore that is deposited in 1 kg of meat. migration: Movement of substances in the environment (e.g., by means of air, surface water, groundwater). Migration may be controlled by several phenomena including, for example, advection, diffusion and sorption. milk transfer coefficient (Fm): Fraction of the amount of an element ingested daily by a cow or goat that is secreted in 1 L of milk. mill tailings: Residues from chemical processing of uranium or thorium ores for their source material content. Mill tailings are a form of byproduct material, as defined in Section 11(e)(2) of the Atomic Energy Act. mixed-bed: Mixtures of materials in ion-exchange media that sorb anions and cations. mixed waste: Waste containing radionuclides (i.e., source, special nuclear, or byproduct materials), as defined in the Atomic Energy Act, and hazardous chemical waste regulated under the Resource Conservation and Recovery Act (RCRA). Mixed waste also may include (1) waste

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containing radionuclides as defined in the Atomic Energy Act and hazardous chemical waste regulated under the Toxic Substances Control Act (TSCA) and (2) waste containing naturally occurring and accelerator-produced radioactive material and hazardous chemical waste regulated under RCRA or TSCA. model: Mathematical or physical representation of an environmental or biological system, sometimes including specific numerical values for parameters of the system. model calibration: Process of fitting a model to represent site-specific conditions using data for that site. model structure: With reference to mathematical models, the conceptual design, mathematical equations, and set of algorithms that determine results or predictions produced from a given set of input data or assumptions. model validation: Investigation of the extent of discrepancy (or agreement) between model predictions and data sets that were not used in developing the model or its parameters and were obtained over a range of conditions that represent the extent of intended application of the model. modulus of elasticity: Ratio of stress placed on a material to resulting strain. moisture content: Volumetric fraction of a porous medium, including solid medium and pore spaces, that is occupied by liquid water. monitoring: Measurement of radiation levels or quantities of radionuclides in environmental media. Monte-Carlo analysis: Computation of a probability distribution of an output of a model on the basis of repeated calculations using random sampling of input variables from specified probability distributions. multiplicative-chain model: Model in which the dependent variable is expressed as simple products (or quotients) of independent variables. naturally occurring and accelerator-produced radioactive material (NARM): Any naturally occurring radioactive material other than source, special nuclear, or byproduct material, or any radioactive material produced in an accelerator. near-surface disposal: Disposal of waste, with or without engineered barriers, on or below the ground surface, such that the final protective covering is on the order of a few meters thick, or in mined openings within a few tens of meters of the Earth’s surface. nuclear fuel cycle: Activities associated with production, utilization and disposition of fuel for nuclear reactors, including power reactors, research reactors, and defense and isotope production reactors, and byproducts related to such activities (see uranium fuel cycle). Nuclear Waste Policy Act (NWPA): Law passed in 1982 and amended in 1987 that governs U.S. Department of Energy (DOE) program for disposal of commercial spent nuclear fuel and high-level radioactive waste in a geologic repository. Act also governs disposal of DOE’s spent nuclear fuel and high-level waste when such waste is co-disposed in

386 / GLOSSARY the same facility as commercial waste. All waste disposals under the Act will be licensed by the U.S. Nuclear Regulatory Commission. numerical solution: Solution of a differential or integral equation that is not expressed in algebraic form but is expressed in terms of numerical values of dependent variables corresponding to specified values of independent variables. osmosis: Passage of a solvent through a semipermeable membrane or porous partition in response to a concentration gradient. oxidation: Any reaction in which one or more electrons are removed from a chemical species. parameter: Any of a set of independent variables in a model whose values determine the characteristics or behavior of the model. partition coefficient: Ratio of mass of a contaminant in solution per unit volume of water to the mass of the contaminant in a waste form per unit mass of the waste form at equilibrium. pathway: Route or means of release of contaminants from a disposal facility, transport of contaminants in the environment, or exposure of humans. PCBs (polychlorinated biphenyls): Family of chemicals composed of biphenyl molecules that have been chlorinated to varying degrees. Peclet number: Average groundwater flow velocity multiplied by a characteristic length (e.g., average particle diameter) divided by the diffusion coefficient. Used to identify whether mechanical dispersion or diffusion is the controlling transport mechanism in a flow system. For example, when the Peclet number is >10, mechanical dispersion is dominant, and when the Peclet number is <1, diffusion is dominant. percolation: (see infiltration). performance assessment: Iterative process involving site-specific, prospective modeling evaluations of the postclosure time phase of a waste disposal system for the purpose of (1) determining whether reasonable assurance of compliance with regulatory performance objectives can be demonstrated, and (2) identifying critical data, facility design, and model development needs for defensible and cost-effective licensing decisions and developing and maintaining operating limits (waste acceptance criteria) for specific disposal facilities. performance objective: Standard that defines acceptable projected exposures of individuals or populations or concentrations of radionuclides in the accessible environment following closure of a waste disposal facility. permeability: Same as hydraulic conductivity, except definition applies to any fluid. perturbation analysis: Method of investigating effects of parameter uncertainty on output of a model in which probability distribution functions of parameters are incorporated explicitly in model equations. pH: Symbol for degree of acidity or alkalinity of a solution. photon: Quantum of electromagnetic radiation, having no charge or mass, that exhibits both particle and wave behavior, especially a gamma or x ray.

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physisorption: Adsorption in which forces involved are intermolecular forces (i.e., van der Waals forces) of the same kind as those responsible for the imperfection of rare gases and condensation of vapors, and which do not involve a significant change in electronic orbital patterns of species involved. piezometric surface: In a confined groundwater system, surface obtained by connecting equilibrium water levels in tubes penetrating the aquifer and open to the atmosphere. In an unconfined aquifer, the piezometric surface is the same as the water table. Also called the potentiometric surface. plant-to-soil concentration ratio (Bv): Ratio of concentration of a radionuclide in vegetation to concentration in dry soil. pore velocity: Average rate of flow in pores of a groundwater medium, which is approximated as the flux divided by the effective porosity. porosity: Property of voids or openings in a porous medium; total porosity is the ratio of the volume of all openings to the total volume of the material, and effective porosity refers to the porosity of that part of the material through which flow occurs. possibilistic, possibility theory: Generalization of probability theory developed to account for limitations of classical probability theory in dealing with imprecision (Walley, 1991). postclosure: Times subsequent to cessation of waste emplacement activities at a disposal facility and actions to prepare disposal site for long-term waste isolation (e.g., construction of caps and covers, seals, and surface markers). precipitation: Rain, snow, sleet or any other form of water deposited from the atmosphere onto the Earth’s surface. precision: Acceptable degree of uncertainty of an estimate with respect to an actual event. pressure head: Height above a given point in an aquifer to which water will rise when an open, vertical tube is inserted down to that point. probabilistic: (see stochastic). probabilistic methods: (see stochastic methods). probabilistic risk assessment: Type of risk assessment in which probabilistic methods are used to describe processes and events acting on a natural or engineered system and their consequences and to derive a distribution of risk based on repeated random sampling of distributions of input variables. probability distribution: Estimate of the likelihood of occurrence of different possible values of a model parameter or model output. pump test: Method of estimating hydraulic properties of aquifers in which a well is pumped at a constant rate and drawdown of the piezometric surface or water table is observed in wells at some distance from the pumped well. In a steady-state pump test, pumping is continued until water levels in observation wells approach equilibrium, which enables a calculation of the transmissivity of the aquifer. In a transient pump test, drops in water levels in observation wells are

388 / GLOSSARY measured as a function of pumping time, which enables a calculation of specific storage as well as transmissivity. quality assurance: Process of ensuring proper documentation of data, interpretations of data, which are embodied in assumptions, and computer codes. quality factor (Q): Dimensionless factor developed for purposes of radiation protection and assessing health risks in general terms that accounts for relative biological effectiveness of different radiations in producing stochastic effects and is used to relate absorbed dose (D) at a point in tissue to dose equivalent (H): H = D × Q. The quality factor is a specified function of unrestricted linear energy transfer (L) in water (ICRP, 1991), and is defined with respect to the particular type and energy of radiation incident on tissue at the point of interest. Prior to introduction of the radiation weighting factor and as often used by federal and state agencies, a mean quality factor ( Q ) for any energy of a particular radiation type (e.g., 1 for all photons and electrons, 20 for all alpha particles) is used to relate a mean absorbed dose in an organ or tissue (DT) to the mean dose equivalent in that organ or tissue (HT): HT = DT × Q . radiation weighting factor (wR): Dimensionless weighting factor developed for purposes of radiation protection and assessing health risks in general terms that accounts for relative biological effectiveness of different types (and, in some cases, energies) of radiations in producing stochastic effects and is used to relate mean absorbed dose in an organ or tissue (T) to equivalent dose: HT = DT × wR (ICRP, 1991). The radiation weighting factor is intended to supersede the mean quality factor ( Q ) and is defined with respect to the type and energy of the radiation incident on the body or, in the case of sources within the body, emitted by the source. Values of wR include 1 for all photons and electrons and 20 for all alpha particles. Radiation weighting factor is independent of tissue weighting factor (wT). radioactive waste: Solid, liquid or gaseous materials of no value that contain radionuclides, either man-made or naturally occurring, and are regulated as hazardous material due to the presence of radionuclides. radioactivity: Property or characteristic of an unstable atomic nucleus to spontaneously transform with emission of energy in the form of ionizing radiation. radiolysis: Chemical decomposition brought about by radiation. radionuclide: Naturally occurring or artificially produced radioactive element or isotope. radon: Colorless, odorless, naturally occurring, and radioactive gaseous element formed by radioactive decay of isotopes of radium. radon progeny: Short-lived decay products of 222Rn (i.e., 218Po, 214Pb, 214 Bi, and 214Po) or 220Rn (i.e., 216Po, 212Pb, 212Bi, and 212Po). reasonable assurance: Concept that acknowledges that proof of the future performance of waste disposal systems over time periods of many hundreds to thousands of years is not to be had in the ordinary

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sense of the word. In judging the long-term performance of waste disposal systems on the basis of model predictions or studies of physical systems including natural analogs, what is required is reasonable assurance, making allowance for the time period, hazards, and uncertainties involved, that the outcome of waste disposal will be in compliance with applicable regulatory performance objectives. redox: Related to or producing processes of oxidation and reduction. reduction: Chemical change that involves a gain of electrons, either by removal of oxygen or an addition of hydrogen, or simply by addition of electrons. reference individual: Stylized representation of a human of specified age with defined anatomical and physiological characteristics, which is used for purposes of radiation protection. regression analysis: Analysis, which is based on empirical data, of relationship between a variable and one or more other variables that takes into account the degree of correlation among the variables. relative biological effectiveness: For a specific radiation (A), ratio of absorbed dose of a reference radiation required to produce a specific level of a response in a biological system to absorbed dose of radiation (A) required to produce an equal response. Reference radiation normally is gamma rays or x rays with an average linear energy transfer of 3.5 keV µm–1 or less. Relative biological effectiveness generally depends on dose, dose per fraction if the dose is fractionated, dose rate, and biological endpoint. rem: (see dose equivalent or equivalent dose). Resource Conservation and Recovery Act (RCRA): Law passed in 1976 as amendment to Solid Waste Disposal Act of 1975, and amended in 1980 and again in 1984 by Hazardous and Solid Waste Amendments, that governs generation, transport, treatment, storage and disposal of solid hazardous waste and disposal of nonhazardous solid waste in municipal/industrial landfills. Solid hazardous wastes regulated under RCRA are defined in 40 CFR Part 261, Subpart A, and specifically exclude source, special nuclear, and byproduct materials as defined in the Atomic Energy Act. Objectives of RCRA include protection of human health and the environment, expeditious reduction or elimination of generation of hazardous waste, and conservation of energy and natural resources (i.e., material recycling and recovery). response surface: Parametric approximation of a multiparameter (usually linear) function; coefficients of the function are fit to the parametric approximation using a limited set of responses, and the function is used to approximate other responses in the domain. resuspension: Transfer of material that has been deposited on the ground surface to the atmosphere; also commonly used to mean suspension for material on the ground surface that was not deposited from the atmosphere. resuspension factor: Ratio of concentration of a radionuclide in air, often at height of 1 m above ground, to its areal concentration on the ground surface; usually given in units of m–1.

390 / GLOSSARY resuspension rate: Fractional rate of release of radionuclides from ground surface into air; usually given in units of s–1. retardation coefficient, retardation factor: Measure of capability of a porous medium to impede by sorption the movement of a contaminant being carried by a fluid, given by ratio of the average velocity of the contaminant to the average velocity of the fluid. risk: Probability of harm, combined with potential severity of that harm. For example, in regard to impacts on human health resulting from disposal of radioactive waste, risk is the probability of a response (e.g., cancer) in an individual or frequency of a response in a population taking into account (1) the probability of occurrence of processes and events that could result in release of radionuclides to the environment and the magnitude of such releases, (2) the probability that individuals or populations would be exposed to radionuclides released to the environment and the magnitude of such exposures, and (3) the probability than a given exposure would produce a response. risk assessment: Analysis of potential adverse impacts of an event (e.g., radioactive waste disposal) upon the well-being of an individual or a population. Risk assessment is a process by which information or experience concerning causes and effects under a given set of circumstances is integrated with the extent of those circumstances to quantify or otherwise describe risk. risk coefficient: (1) Probability of a cancer (fatal cancer or cancer incidence) per unit radiation dose; or (2) probability of a cancer per unit activity intake of a radionuclide or per disintegration per unit volume, area or mass of a radionuclide in the environment. robust: Insensitive to changes in assumptions, data, or other information, or resistant to a broad range of external conditions. Safe Drinking Water Act (SDWA): Law passed in 1974 and amended several times since, most recently by the Safe Drinking Water Act Amendments of 1996, that addresses protection of the nation’s drinking water supplies and resources. Act provides authority for National Primary Drinking Water Regulations for hazardous contaminants in drinking water, including radionuclides, and national requirements for State Underground Injection Control programs. saturated zone: Portion of subsurface zone in a geohydrologic system in which all openings (void spaces or pores) are filled with water. saturation: Fraction of openings (void spaces or pores) of subsurface zone that are filled with water. scaling factor: Factor used to infer the value of an unknown quantity from a known or measured quantity; often used to estimate the inventory of a radionuclide that is difficult to measure in low-level waste on the basis of the inventory of an easy-to-measure radionuclide. scenario: Set of assumptions about the future behavior of a disposal system or future exposures of individuals. scenario dose conversion factor: For a specified exposure scenario, dose or dose rate per unit concentration of a radionuclide in a specified source region.

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screening: Process of rapidly identifying potentially important radionuclides or release, transport or exposure pathways by eliminating those of known lesser significance. screening models: Simple models employing conservative assumptions for the expressed purpose of eliminating radionuclides and pathways of negligible importance. seepage: (see infiltration). sensitivity analysis: Analysis of effects of perturbations in input parameters or other assumptions on output of a model, without regard for their possible variability or uncertainty. shallow-land disposal: (see near-surface disposal). Shannon’s entropy: Expected value of the amount of information available to associate with a probability. Using Shannon’s entropy, the least-biased probability density function consistent with a given set of information can be defined (Hamming, 1991). sievert (Sv): Special name for the SI unit of dose equivalent, equivalent dose, and effective dose. site-specific: Pertaining to a particular location (e.g., of a disposal facility). slug test: Method of estimating hydraulic conductivity in aquifers in which a slug of water is removed from a well and the rate at which the piezometric surface or water table returns to its original level is observed. Specific storage normally cannot be evaluated using this method. solubility limit: Maximum amount of a substance that can be dissolved in a unit volume of a solvent under specified conditions. sorption: Any mechanism that removes ions from fluid phase of a medium to solid phase, including adsorption (attraction to a surface), absorption (incorporation into the interior of a solid), and ion exchange (adsorption, with a charge-for-charge replacement of ionic species on a surface by other ionic species in solution). Desorption includes the opposite of each of these reactions. sorption isotherm: Curve of sorption as a function of concentration of a contaminant in a solid/fluid system with all other parameters held constant (e.g., temperature, pressure, ratio of amount of fluid to amount of solid). source material: (1) Uranium or thorium or any combination of uranium or thorium in any physical or chemical form, or (2) ores that contain, by weight, 0.05 % or more of uranium, thorium, or any combination of uranium or thorium. Source material does not include special nuclear material. source term: Rate of release of radionuclides from a waste disposal facility, usually over time. spallation: Breaking off of material in layers parallel to a surface. special nuclear material: (1) Plutonium, uranium enriched in the isotope 233 or 235, and any other material that the U.S. Nuclear Regulatory Commission determines to be special nuclear material, or (2) any

392 / GLOSSARY material artificially enriched in any of the foregoing. Special nuclear material does not include source material. speciation: Particular chemical compounds or forms and valence states in which a chemical element can exist in a given environment. specific activity: Activity of a radionuclide per unit mass of the radionuclide; also may refer to activity of a radionuclide per unit mass of material in which the radionuclide is dispersed. specific activity model: Model that estimates dose from exposure to a radionuclide by assuming that the specific activity in tissues of an exposed individual is equal to, or a fraction of, the specific activity in air or water at a particular location. specific storage: Volume of water released from a unit volume of a saturated medium under a unit decline in hydraulic head. spent nuclear fuel: Fuel that has been withdrawn from a nuclear reactor following irradiation, the constituent elements of which have not been separated by reprocessing. State Compacts: Groups of two or more states that, when approved by Congress under authority of the Low-Level Radioactive Waste Policy Act of 1980, as amended, join together to develop a disposal facility for low-level waste generated within those states. steady-state: Constant in time. step-wise regression: Procedure that fits a multiple linear regression model to data in an iterative manner. Procedure begins by including variables that explain most of the variation in outcomes, and progressively adds more variables until a specified significance level is achieved. stochastic: Of, pertaining to, or arising from chance; involving probability; random. stochastic effects: Adverse effects in biological organisms for which the probability, but not the severity, is assumed to be a function of dose of ionizing radiation (or other contaminant) without threshold. stochastic methods: Methods of exercising mathematical models in which distributions of input parameters are propagated through a model to estimate distributions in outputs. stochastic variation: Random variability attributed to a property of a system based on repeated measurement. storage: Retention of waste with the intent to retrieve it for subsequent use, processing or disposal. strain: Relative change in dimensions of an object in response to an applied force. stress: External force acting on a material that tends to change dimensions of the material by compressing it, stretching it, or causing it to shear. suction pressure: Negative pressure head in an unsaturated medium. Superfund: (see Comprehensive Environmental Response, Compensation, and Liability Act). surface water: Water on or above the land surface (e.g., rivers, streams, lakes, ponds, oceans).

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surveillance: Inspection of all or part of a waste disposal facility, usually for the purpose of ascertaining its integrity or normal or expected functioning. suspension: Transfer of material from the Earth’s surface, including surface water and the land surface, to the atmosphere. tensile strength: Greatest stress that a material can bear without failure. tension: Negative pressure compared with atmospheric pressure. tissue weighting factor (wT): Dimensionless factor that represents ratio of the stochastic risk attributable to a specific organ or tissue (T) to total stochastic risk attributable to all organs and tissues when the whole body receives a uniform exposure to ionizing radiation. When calculating effective dose equivalent, tissue weighting factor represents the risk of fatal cancers or severe hereditary effects (ICRP, 1977). When calculating effective dose, tissue weighting factor represents total detriment (ICRP, 1991). tortuosity: Measure of deviation of individual flow paths from average direction of flow, given by difference between actual length of a flow path and a theoretical straight course. Toxic Substances Control Act (TSCA): Law passed in 1976 that governs regulation of toxic substances in commerce, with objective of preventing human health and environmental problems before they occur. Manufacturing, processing or distribution in commerce of toxic substances may be limited or banned if U.S. Environmental Protection Agency finds, based on results of toxicity testing and exposure assessments, that there is an unreasonable risk of injury to health or the environment. Hazardous chemicals regulated under TSCA include, for example, dioxins, PCBs and asbestos. transfer factor: Factor representing intake of a radionuclide by an exposed individual per unit concentration in an environmental medium (air, water or soil); not relevant for external exposure pathways. translocation factor: Fraction of material deposited from the atmosphere onto surfaces of vegetation that is transferred to edible parts of the plant. transmissivity: Rate at which water is transmitted through a unit width of an aquifer under a unit hydraulic gradient, given by product of the hydraulic conductivity and aquifer thickness. transpiration: Giving off of water through surfaces of leaves and other parts of plants. transport: (see migration). transuranic waste: Radioactive waste containing more than 4 kBq g–1 of alpha-emitting transuranium isotopes, with half-lives >20 y, except for (1) high-level waste, (2) radioactive waste that the Secretary of Energy has determined, with concurrence of the Administrator of the U.S. Environmental Protection Agency, does not need the degree of isolation required by disposal regulations in 40 CFR Part 191, or (3) radioactive waste that the U.S. Nuclear Regulatory Commission has

394 / GLOSSARY approved for near-surface disposal on a case-by-case basis in accordance with 10 CFR Part 61. transuranium element: Chemical element with an atomic number greater than that of uranium (92). Same as transuranic element. treatment: Any method, technique or process designed to change the physical or chemical character of a hazardous material to render it less hazardous, safer to transport, store or dispose of, or to reduce its volume. tumulus: An artificial mound. The tumulus concept for waste disposal involves construction of a pad (normally concrete) on or below the ground surface, placement of individual waste containers or packages on the pad (normally in stacks), and placement of earthen materials above the waste to form a cap. uncertainty: Lack of sureness or confidence in predictions of models or results of measurements. Uncertainties may be categorized as Type A (or aleatory), which are those due to stochastic variation, or Type B (or epistemic), which are those due to lack of knowledge founded on an incomplete characterization, understanding or measurement of a system. Type-B uncertainties generally are the most important in assessing the long-term performance of waste disposal systems. uncertainty analysis: Analysis of variability in model predictions due to uncertainty in input parameters or other assumptions. undisturbed performance: Projected performance of a waste disposal system absent human intrusion. unsaturated zone: Portion of subsurface zone, usually starting at the land surface, in which openings (voids and pores) are not completely filled with water. uranium fuel cycle: Normal life cycle of uranium used as fuel in nuclear reactors including: mining of uranium ore; milling of uranium ore to produce U3O8 concentrate; refining of that concentrate to remove impurities; chemical conversion of that concentrate to UF6; enrichment of the 235U content by gaseous diffusion; fabrication of nuclear fuel (usually by chemical conversion to UO2 and sintering into fuel rods); burning of fuel in a nuclear reactor for electricity generation, defense plutonium production, production of other isotopes, or research and development; chemical reprocessing of spent nuclear fuel to recover the remaining uranium and plutonium (commercial spent nuclear fuel is not current reprocessed); consolidation of spent nuclear fuel rods or encapsulation of liquid reprocessing wastes in borosilicate glass or other suitable waste form preparatory to disposal; and storage and disposal of spent nuclear fuel and solidified reprocessing waste. usage factor: Factor representing intakes of contaminated material by humans or livestock or residence time in a contaminated area for external exposure. vadose zone: (see unsaturated zone). verification: Determination that a computer (or any other) implementation of a mathematical equation or set of equations is without significant error.

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viscosity: Internal friction of a fluid, caused by molecular attraction, that makes it resist a tendency to flow. waste classification: Any grouping of wastes having similar attributes. waste classification system: (1) System for classifying waste arising from operations of nuclear fuel cycle including spent nuclear fuel (if it is declared to be waste), high-level waste, transuranic waste, low-level waste, and uranium or thorium mill tailings; or (2) system for classifying radioactive waste that is generally acceptable for near-surface disposal developed by U.S. Nuclear Regulatory Commission in 10 CFR Part 61. waste dilution factor: Fraction of a disposal facility that is occupied by waste following closure of the facility. waste form: Radioactive waste materials and any encapsulating or stabilizing matrix. waste management: Activities associated with disposition of waste products after their generation, including treatment, storage, transportation and disposal, as well as actions to minimize production of waste. waste package: Waste form and any containers, shielding, packing and other absorbent materials immediately surrounding a waste form. water table: Level in saturated zone at which water pressure is equal to atmospheric pressure. weathering half-time: Time required for half the material deposited onto surfaces of vegetation from the atmosphere to be lost from plants (e.g., by actions of wind or rain). x ray: (1) Electromagnetic radiation emitted in de-excitation of bound atomic electrons, and frequently occurring in decay of radionuclides, referred to as characteristic x rays, or (2) electromagnetic radiation produced in deceleration of energetic charged particles (e.g., beta radiation) in passing through matter, referred to as continuous x rays or bremsstrahlung (see gamma ray and photon). yield strength: Stress level at which a material attains a specified strain.

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432 / REFERENCES SELANDAR, W.N., WILKINSON, S.R. and ROWAT, J.H. (1991). A Vault Mass Transfer Model for LLRW, Atomic Energy of Canada, Ltd., Report TR-509 (Atomic Energy of Canada, Ltd., Chalk River, Ontario, Canada). SENES (2005). SENES Oak Ridge, Inc. “International data sets for use in testing environmental transport models,” http://www.cdc.gov/nceh/ radiation/brochure/pdf/CDC_Introduction/pdf (accessed December 31, 2005) (SENES Oak Ridge, Inc., Oak Ridge, Tennessee). SERNE, R.J., ARTHUR, R.C. and. KRUPKA, K.M. (1990). Review of Geochemical Processes and Codes for Assessment of Radionuclide Migration Potential at Commercial LLW Sites, U.S. Nuclear Regulatory Commission Report NUREG/CR-5548 (National Technical Information Service, Springfield, Virginia). SHANNON, C.E. (1948). “A Mathematical Theory of Communication,” Bell Sys. Tech. J. 27, 379–423, 623–656. SHEPPARD, M.I. and THIBAULT, D.H. (1990). “Default soil solid/liquid partition coefficients, Kds, for four major soil types: A compendium,” Health Phys. 59, 471–482. SHEPPARD, M.I., SHEPPARD, S.C. and AMIRO, B.D. (1991). “Mobility and plant uptake of inorganic 14C and 14C-labelled PCB in soils of high and low retention,” Health Phys. 61, 481–492. SHINN, J.H. (1991). “Enhancement factors for resuspended aerosol radioactivity: Effects of topsoil disturbance,” pages 1183 to 1193 in Precipitation Scavenging and Atmosphere-Surface Exchange, Schwartz, S.E. and Slinn, W.G.N., Eds. (National Technical Information Service, Springfield, Virginia). SHINN, J.H., HOMAN, D.N. and HOFMANN, C.B. (1986). A Summary of Plutonium Aerosol Studies: Resuspension at the Nevada Test Site, Lawrence Livermore National Laboratory Report UCRL-90746 (National Technical Information Service, Springfield, Virginia). SHINN, J.H., HOMAN, D.N. and ROBINSON, W.L. (1989). Resuspension Studies at Bikini Atoll, Lawrence Livermore National Laboratory Report UCID-18538, Rev. 1 (National Technical Information Service, Springfield, Virginia). SHIPERS, L.R. (1989). Background Information for the Development of a Low-Level Waste Performance Assessment Methodology: Identification of Potential Exposure Pathways, Vol. 1, U.S. Nuclear Regulatory Commission Report NUREG/CR-5453 (National Technical Information Service, Springfield, Virginia). SHIPERS, L.R. and HARLAN, C.P. (1989). Background Information for the Development of a Low-Level Waste Performance Assessment Methodology: Assessment of Relative Significance of Migration and Exposure Pathways, Vol. 2, U.S. Nuclear Regulatory Commission Report NUREG/CR-5453 (National Technical Information Service, Springfield, Virginia).

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The NCRP

The National Council on Radiation Protection and Measurements is a nonprofit corporation chartered by Congress in 1964 to: 1. Collect, analyze, develop and disseminate in the public interest information and recommendations about (a) protection against radiation and (b) radiation measurements, quantities and units, particularly those concerned with radiation protection. 2. Provide a means by which organizations concerned with the scientific and related aspects of radiation protection and of radiation quantities, units and measurements may cooperate for effective utilization of their combined resources, and to stimulate the work of such organizations. 3. Develop basic concepts about radiation quantities, units and measurements, about the application of these concepts, and about radiation protection. 4. Cooperate with the International Commission on Radiological Protection, the International Commission on Radiation Units and Measurements, and other national and international organizations, governmental and private, concerned with radiation quantities, units and measurements and with radiation protection. The Council is the successor to the unincorporated association of scientists known as the National Committee on Radiation Protection and Measurements and was formed to carry on the work begun by the Committee in 1929. The participants in the Council’s work are the Council members and members of scientific and administrative committees. Council members are selected solely on the basis of their scientific expertise and serve as individuals, not as representatives of any particular organization. The scientific committees, composed of experts having detailed knowledge and competence in the particular area of the committee's interest, draft proposed recommendations. These are then submitted to the full membership of the Council for careful review and approval before being published. The following comprise the current officers and membership of the Council:

Officers President Senior Vice President Secretary and Treasurer Assistant Secretary

439

Thomas S. Tenforde Kenneth R. Kase David A. Schauer Michael F. McBride

440 / THE NCRP Members John F. Ahearne Sally A. Amundson Benjamin R. Archer Mary M. Austin-Seymour Steven M. Becker Joel S. Bedford Eleanor A. Blakely William F. Blakely John D. Boice, Jr. Wesley E. Bolch Thomas B. Borak Andre Bouville Leslie A. Braby David J. Brenner James A. Brink Antone L. Brooks Jerrold T. Bushberg John F. Cardella Stephanie K. Carlson Polly Y. Chang S.Y. Chen Kelly L. Classic Mary E. Clark Michael L. Corradini J. Donald Cossairt Allen G. Croff Francis A. Cucinotta Paul M. DeLuca David A. Eastmond Stephen A. Feig John R. Frazier Donald P. Frush Thomas F. Gesell

Robert L. Goldberg Andrew J. Grosovsky Raymond A. Guilmette Roger W. Harms Kathryn Held John W. Hirshfeld, Jr. F. Owen Hoffman Roger W. Howell Kenneth R. Kase Ann R. Kennedy William E. Kennedy, Jr. David C. Kocher Ritsuko Komaki Amy Kronenberg Susan M. Langhorst Edwin M. Leidholdt Howard L. Liber James C. Lin Jill A. Lipoti John B. Little Paul A. Locke Jay H. Lubin C. Douglas Maynard Debra McBaugh Cynthia H. McCollough Barbara J. McNeil Fred A. Mettler, Jr. Charles W. Miller Donald L. Miller Jack Miller Kenneth L. Miller William H. Miller William F. Morgan David S. Myers

Bruce A. Napier Gregory A. Nelson Carl J. Paperiello R. Julian Preston Jerome C. Puskin Allan C.B. Richardson Henry D. Royal Michael T. Ryan Jonathan M. Samet Thomas M. Seed Stephen M. Seltzer Roy E. Shore Edward A. Sickles Steven L. Simon Paul Slovic Christopher G. Soares Daniel J. Strom Thomas S. Tenforde Julie E.K. Timins Richard E. Toohey Lawrence W. Townsend Lois B. Travis Fong Y. Tsai Richard J. Vetter Chris G. Whipple Stuart C. White J. Frank Wilson Susan D. Wiltshire Gayle E. Woloschak Shiao Y. Woo Andrew J. Wyrobek Marco A. Zaider Pasquale D. Zanzonico

Honorary Members Warren K. Sinclair, President Emeritus; Charles B. Meinhold, President Emeritus S. James Adelstein, Honorary Vice President W. Roger Ney, Executive Director Emeritus William M. Beckner, Executive Director Emeritus Roger O. McClellan William P. Dornsife Seymour Abrahamson Dade W. Moeller Patricia W. Durbin Edward L. Alpen A. Alan Moghissi Keith F. Eckerman Lynn R. Anspaugh Wesley L. Nyborg Thomas S. Ely John A. Auxier John W. Poston, Sr. Richard F. Foster William J. Bair Andrew K. Poznanski R.J. Michael Fry Harold L. Beck Genevieve S. Roessler Ethel S. Gilbert Bruce B. Boecker Marvin Rosenstein Joel E. Gray Victor P. Bond Lawrence N. Rothenberg Robert O. Gorson Robert L. Brent Eugene L. Saenger Arthur W. Guy Reynold F. Brown William J. Schull Eric J. Hall Melvin C. Carter John E. Till Naomi H. Harley Randall S. Caswell Robert L. Ullrich William R. Hendee Frederick P. Cowan Arthur C. Upton Donald G. Jacobs James F. Crow F. Ward Whicker Bernd Kahn Gerald D. Dodd Marvin C. Ziskin Charles E. Land Sarah S. Donaldson

THE NCRP

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Lauriston S. Taylor Lecturers Robert L. Brent (2006) Fifty Years of Scientific Investigation: The Importance of Scholarship and the Influence of Politics and Controversy John B. Little (2005) Nontargeted Effects of Radiation: Implications for Low-Dose Exposures Abel J. Gonzalez (2004) Radiation Protection in the Aftermath of a Terrorist Attack Involving Exposure to Ionizing Radiation Charles B. Meinhold (2003) The Evolution of Radiation Protection: From Erythema to Genetic Risks to Risks of Cancer to ? R. Julian Preston (2002) Developing Mechanistic Data for Incorporation into Cancer Risk Assessment: Old Problems and New Approaches Wesley L. Nyborg (2001) Assuring the Safety of Medical Diagnostic Ultrasound S. James Adelstein (2000) Administered Radioactivity: Unde Venimus Quoque Imus Naomi H. Harley (1999) Back to Background Eric J. Hall (1998) From Chimney Sweeps to Astronauts: Cancer Risks in the Workplace William J. Bair (1997) Radionuclides in the Body: Meeting the Challenge! Seymour Abrahamson (1996) 70 Years of Radiation Genetics: Fruit Flies, Mice and Humans Albrecht Kellerer (1995) Certainty and Uncertainty in Radiation Protection R.J. Michael Fry (1994) Mice, Myths and Men Warren K. Sinclair (1993) Science, Radiation Protection and the NCRP Edward W. Webster (1992) Dose and Risk in Diagnostic Radiology: How Big? How Little? Victor P. Bond (1991) When is a Dose Not a Dose? J. Newell Stannard (1990) Radiation Protection and the Internal Emitter Saga Arthur C. Upton (1989) Radiobiology and Radiation Protection: The Past Century and Prospects for the Future Bo Lindell (1988) How Safe is Safe Enough? Seymour Jablon (1987) How to be Quantitative about Radiation Risk Estimates Herman P. Schwan (1986) Biological Effects of Non-ionizing Radiations: Cellular Properties and Interactions John H. Harley (1985) Truth (and Beauty) in Radiation Measurement Harald H. Rossi (1984) Limitation and Assessment in Radiation Protection Merril Eisenbud (1983) The Human Environment—Past, Present and Future Eugene L. Saenger (1982) Ethics, Trade-Offs and Medical Radiation James F. Crow (1981) How Well Can We Assess Genetic Risk? Not Very Harold O. Wyckoff (1980) From “Quantity of Radiation” and “Dose” to “Exposure” and “Absorbed Dose”—An Historical Review Hymer L. Friedell (1979) Radiation Protection—Concepts and Trade Offs Sir Edward Pochin (1978) Why be Quantitative about Radiation Risk Estimates? Herbert M. Parker (1977) The Squares of the Natural Numbers in Radiation Protection Currently, the following committees are actively engaged in formulating recommendations:

442 / THE NCRP Program Area Committee 1: Basic Criteria, Epidemiology, Radiobiology, and Risk SC 1-7 Information Needed to Make Radiation Protection Recommendations for Travel Beyond Low-Earth Orbit SC 1-8 Risk to Thyroid from Ionizing Radiation SC 1-13 Impact of Individual Susceptibility and Previous Radiation Exposure on Radiation Risk for Astronauts SC 1-15 Radiation Safety in NASA Lunar Missions SC 85 Risk of Lung Cancer from Radon

Program Area Committee 2: Operational Radiation Safety SC 46-17 Radiation Protection in Educational Institutions

Program Area Committee 3: Nonionizing Radiation SC 89-5 Study and Critical Evaluation of Radiofrequency Exposure Guidelines

Program Area Committee 4: Radiation Protection in Medicine SC 4-1 Management of Persons Contaminated with Radionuclides SC 4-2 Population Monitoring and Decontamination Following a Nuclear/ Radiological Incident SC 91-1 Precautions in the Management of Patients Who Have Received Therapeutic Amounts of Radionuclides

Program Area Committee 5: Environmental Radiation and Radioactive Waste Issues SC 64-22 Design of Effective Effluent and Environmental Monitoring Programs SC 64-23 Cesium in the Environment

Program Area Committee 6: Radiation Measurements and Dosimetry SC 6-1 Uncertainties in the Measurement and Dosimetry of External Radiation Sources SC 6-2 Radiation Exposure of the U.S. Population SC 6-3 Uncertainties in Internal Radiation Dosimetry SC 6-4 Fundamental Principles of Dose Reconstruction SC 57-17 Radionuclide Dosimetry Models for Wounds

Advisory Committee 1: Public Policy and Risk Communication

In recognition of its responsibility to facilitate and stimulate cooperation among organizations concerned with the scientific and related aspects of radiation protection and measurement, the Council has created a category of NCRP Collaborating Organizations. Organizations or groups of organizations that are national or international in scope and are concerned with scientific problems involving radiation quantities, units, measurements and effects, or radiation protection may be admitted to collaborating status by the Council. Collaborating Organizations provide a means by which NCRP can gain input into its activities from a wider segment of society. At the same time, the relationships with the Collaborating Organizations facilitate wider dissemination of information about the Council's activities, interests and concerns. Collaborating Organizations have the opportunity to comment on draft reports (at the time that these are submitted to the members of the Council). This is intended to capitalize on the fact that Collaborating Organizations are in an excellent posi-

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tion to both contribute to the identification of what needs to be treated in NCRP reports and to identify problems that might result from proposed recommendations. The present Collaborating Organizations with which NCRP maintains liaison are as follows: American Academy of Dermatology American Academy of Environmental Engineers American Academy of Health Physics American Association of Physicists in Medicine American College of Medical Physics American College of Nuclear Physicians American College of Occupational and Environmental Medicine American College of Radiology American Conference of Governmental Industrial Hygienists American Dental Association American Industrial Hygiene Association American Institute of Ultrasound in Medicine American Medical Association American Nuclear Society American Pharmaceutical Association American Podiatric Medical Association American Public Health Association American Radium Society American Roentgen Ray Society American Society for Therapeutic Radiology and Oncology American Society of Emergency Radiology American Society of Health-System Pharmacists American Society of Radiologic Technologists Association of Educators in Radiological Sciences, Inc. Association of University Radiologists Bioelectromagnetics Society Campus Radiation Safety Officers College of American Pathologists Conference of Radiation Control Program Directors, Inc. Council on Radionuclides and Radiopharmaceuticals Defense Threat Reduction Agency Electric Power Research Institute Federal Communications Commission Federal Emergency Management Agency Genetics Society of America Health Physics Society Institute of Electrical and Electronics Engineers, Inc. Institute of Nuclear Power Operations International Brotherhood of Electrical Workers National Aeronautics and Space Administration National Association of Environmental Professionals National Center for Environmental Health/Agency for Toxic Substances National Electrical Manufacturers Association National Institute for Occupational Safety and Health National Institute of Standards and Technology Nuclear Energy Institute Office of Science and Technology Policy

444 / THE NCRP Paper, Allied-Industrial, Chemical and Energy Workers International Union Product Stewardship Institute Radiation Research Society Radiological Society of North America Society for Risk Analysis Society of Chairmen of Academic Radiology Departments Society of Nuclear Medicine Society of Radiologists in Ultrasound Society of Skeletal Radiology U.S. Air Force U.S. Army U.S. Coast Guard U.S. Department of Energy U.S. Department of Housing and Urban Development U.S. Department of Labor U.S. Department of Transportation U.S. Environmental Protection Agency U.S. Navy U.S. Nuclear Regulatory Commission U.S. Public Health Service Utility Workers Union of America NCRP has found its relationships with these organizations to be extremely valuable to continued progress in its program. Another aspect of the cooperative efforts of NCRP relates to the Special Liaison relationships established with various governmental organizations that have an interest in radiation protection and measurements. This liaison relationship provides: (1) an opportunity for participating organizations to designate an individual to provide liaison between the organization and NCRP; (2) that the individual designated will receive copies of draft NCRP reports (at the time that these are submitted to the members of the Council) with an invitation to comment, but not vote; and (3) that new NCRP efforts might be discussed with liaison individuals as appropriate, so that they might have an opportunity to make suggestions on new studies and related matters. The following organizations participate in the Special Liaison Program: Australian Radiation Laboratory Bundesamt fur Strahlenschutz (Germany) Canadian Nuclear Safety Commission Central Laboratory for Radiological Protection (Poland) China Institute for Radiation Protection Commonwealth Scientific Instrumentation Research Organization (Australia) European Commission Health Council of the Netherlands Institut de Radioprotection et de Surete Nucleaire International Commission on Non-ionizing Radiation Protection International Commission on Radiation Units and Measurements Japan Radiation Council Korea Institute of Nuclear Safety National Radiological Protection Board (United Kingdom)

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Russian Scientific Commission on Radiation Protection South African Forum for Radiation Protection World Association of Nuclear Operations World Health Organization, Radiation and Environmental Health NCRP values highly the participation of these organizations in the Special Liaison Program. The Council also benefits significantly from the relationships established pursuant to the Corporate Sponsor's Program. The program facilitates the interchange of information and ideas and corporate sponsors provide valuable fiscal support for the Council's program. This developing program currently includes the following Corporate Sponsors: Duke Energy Corporation GE Healthcare Global Dosimetry Solutions, Inc. Landauer, Inc. Nuclear Energy Institute The Council's activities have been made possible by the voluntary contribution of time and effort by its members and participants and the generous support of the following organizations: 3M Health Physics Services Agfa Corporation Alfred P. Sloan Foundation Alliance of American Insurers American Academy of Dermatology American Academy of Health Physics American Academy of Oral and Maxillofacial Radiology American Association of Physicists in Medicine American Cancer Society American College of Medical Physics American College of Nuclear Physicians American College of Occupational and Environmental Medicine American College of Radiology American College of Radiology Foundation American Dental Association American Healthcare Radiology Administrators American Industrial Hygiene Association American Insurance Services Group American Medical Association American Nuclear Society American Osteopathic College of Radiology American Podiatric Medical Association American Public Health Association American Radium Society American Roentgen Ray Society American Society of Radiologic Technologists American Society for Therapeutic Radiology and Oncology American Veterinary Medical Association

446 / THE NCRP American Veterinary Radiology Society Association of Educators in Radiological Sciences, Inc. Association of University Radiologists Battelle Memorial Institute Canberra Industries, Inc. Chem Nuclear Systems Center for Devices and Radiological Health College of American Pathologists Committee on Interagency Radiation Research and Policy Coordination Commonwealth Edison Commonwealth of Pennsylvania Consolidated Edison Consumers Power Company Council on Radionuclides and Radiopharmaceuticals Defense Nuclear Agency Defense Threat Reduction Agency Eastman Kodak Company Edison Electric Institute Edward Mallinckrodt, Jr. Foundation EG&G Idaho, Inc. Electric Power Research Institute Electromagnetic Energy Association Federal Emergency Management Agency Florida Institute of Phosphate Research Florida Power Corporation Fuji Medical Systems, U.S.A., Inc. Genetics Society of America Global Dosimetry Solutions Health Effects Research Foundation (Japan) Health Physics Society ICN Biomedicals, Inc. Institute of Nuclear Power Operations James Picker Foundation Martin Marietta Corporation Motorola Foundation National Aeronautics and Space Administration National Association of Photographic Manufacturers National Cancer Institute National Electrical Manufacturers Association National Institute of Standards and Technology New York Power Authority Philips Medical Systems Picker International Public Service Electric and Gas Company Radiation Research Society Radiological Society of North America Richard Lounsbery Foundation Sandia National Laboratory Siemens Medical Systems, Inc. Society of Nuclear Medicine Society of Pediatric Radiology Southern California Edison Company

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U.S. Department of Energy U.S. Department of Labor U.S. Environmental Protection Agency U.S. Navy U.S. Nuclear Regulatory Commission Victoreen, Inc. Westinghouse Electric Corporation Initial funds for publication of NCRP reports were provided by a grant from the James Picker Foundation. NCRP seeks to promulgate information and recommendations based on leading scientific judgment on matters of radiation protection and measurement and to foster cooperation among organizations concerned with these matters. These efforts are intended to serve the public interest and the Council welcomes comments and suggestions on its reports or activities.

NCRP Publications NCRP publications can be obtained online in both hard- and soft-copy (downloadable PDF) formats at http://NCRPpublications.org. Professional societies can arrange for discounts for their members by contacting NCRP. Additional information on NCRP publications may be obtained from the NCRP website (http://NCRPonline.org) or by telephone (800-229-2652, ext. 25) and fax (301-907-8768). The mailing address is: NCRP Publications 7910 Woodmont Avenue Suite 400 Bethesda, MD 20814-3095

Abstracts of NCRP reports published since 1980, abstracts of all NCRP commentaries, and the text of all NCRP statements are available at the NCRP website. Currently available publications are listed below. NCRP Reports No. 8 22 25 27 30 32 35 36 37 38 40 41 42 44 46 47

Title Control and Removal of Radioactive Contamination in Laboratories (1951) Maximum Permissible Body Burdens and Maximum Permissible Concentrations of Radionuclides in Air and in Water for Occupational Exposure (1959) [includes Addendum 1 issued in August 1963] Measurement of Absorbed Dose of Neutrons, and of Mixtures of Neutrons and Gamma Rays (1961) Stopping Powers for Use with Cavity Chambers (1961) Safe Handling of Radioactive Materials (1964) Radiation Protection in Educational Institutions (1966) Dental X-Ray Protection (1970) Radiation Protection in Veterinary Medicine (1970) Precautions in the Management of Patients Who Have Received Therapeutic Amounts of Radionuclides (1970) Protection Against Neutron Radiation (1971) Protection Against Radiation from Brachytherapy Sources (1972) Specification of Gamma-Ray Brachytherapy Sources (1974) Radiological Factors Affecting Decision-Making in a Nuclear Attack (1974) Krypton-85 in the Atmosphere—Accumulation, Biological Significance, and Control Technology (1975) Alpha-Emitting Particles in Lungs (1975) Tritium Measurement Techniques (1976)

448

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Structural Shielding Design and Evaluation for Medical Use of X Rays and Gamma Rays of Energies Up to 10 MeV (1976) Environmental Radiation Measurements (1976) Cesium-137 from the Environment to Man: Metabolism and Dose (1977) Medical Radiation Exposure of Pregnant and Potentially Pregnant Women (1977) Protection of the Thyroid Gland in the Event of Releases of Radioiodine (1977) Instrumentation and Monitoring Methods for Radiation Protection (1978) A Handbook of Radioactivity Measurements Procedures, 2nd ed. (1985) Physical, Chemical, and Biological Properties of Radiocerium Relevant to Radiation Protection Guidelines (1978) Radiation Safety Training Criteria for Industrial Radiography (1978) Tritium in the Environment (1979) Tritium and Other Radionuclide Labeled Organic Compounds Incorporated in Genetic Material (1979) Influence of Dose and Its Distribution in Time on Dose-Response Relationships for Low-LET Radiations (1980) Management of Persons Accidentally Contaminated with Radionuclides (1980) Radiofrequency Electromagnetic Fields—Properties, Quantities and Units, Biophysical Interaction, and Measurements (1981) Radiation Protection in Pediatric Radiology (1981) Dosimetry of X-Ray and Gamma-Ray Beams for Radiation Therapy in the Energy Range 10 keV to 50 MeV (1981) Nuclear Medicine—Factors Influencing the Choice and Use of Radionuclides in Diagnosis and Therapy (1982) Radiation Protection and Measurement for Low-Voltage Neutron Generators (1983) Protection in Nuclear Medicine and Ultrasound Diagnostic Procedures in Children (1983) Biological Effects of Ultrasound: Mechanisms and Clinical Implications (1983) Iodine-129: Evaluation of Releases from Nuclear Power Generation (1983) Radiological Assessment: Predicting the Transport, Bioaccumulation, and Uptake by Man of Radionuclides Released to the Environment (1984) Exposures from the Uranium Series with Emphasis on Radon and Its Daughters (1984) Evaluation of Occupational and Environmental Exposures to Radon and Radon Daughters in the United States (1984) Neutron Contamination from Medical Electron Accelerators (1984) Induction of Thyroid Cancer by Ionizing Radiation (1985) Carbon-14 in the Environment (1985) SI Units in Radiation Protection and Measurements (1985) The Experimental Basis for Absorbed-Dose Calculations in Medical Uses of Radionuclides (1985) General Concepts for the Dosimetry of Internally Deposited Radionuclides (1985) Biological Effects and Exposure Criteria for Radiofrequency Electromagnetic Fields (1986)

450 / NCRP PUBLICATIONS 87 Use of Bioassay Procedures for Assessment of Internal Radionuclide Deposition (1987) 88 Radiation Alarms and Access Control Systems (1986) 89 Genetic Effects from Internally Deposited Radionuclides (1987) 90 Neptunium: Radiation Protection Guidelines (1988) 92 Public Radiation Exposure from Nuclear Power Generation in the United States (1987) 93 Ionizing Radiation Exposure of the Population of the United States (1987) 94 Exposure of the Population in the United States and Canada from Natural Background Radiation (1987) 95 Radiation Exposure of the U.S. Population from Consumer Products and Miscellaneous Sources (1987) 96 Comparative Carcinogenicity of Ionizing Radiation and Chemicals (1989) 97 Measurement of Radon and Radon Daughters in Air (1988) 99 Quality Assurance for Diagnostic Imaging (1988) 100 Exposure of the U.S. Population from Diagnostic Medical Radiation (1989) 101 Exposure of the U.S. Population from Occupational Radiation (1989) 102 Medical X-Ray, Electron Beam and Gamma-Ray Protection for Energies Up to 50 MeV (Equipment Design, Performance and Use) (1989) 103 Control of Radon in Houses (1989) 104 The Relative Biological Effectiveness of Radiations of Different Quality (1990) 105 Radiation Protection for Medical and Allied Health Personnel (1989) 106 Limit for Exposure to “Hot Particles” on the Skin (1989) 107 Implementation of the Principle of As Low As Reasonably Achievable (ALARA) for Medical and Dental Personnel (1990) 108 Conceptual Basis for Calculations of Absorbed-Dose Distributions (1991) 109 Effects of Ionizing Radiation on Aquatic Organisms (1991) 110 Some Aspects of Strontium Radiobiology (1991) 111 Developing Radiation Emergency Plans for Academic, Medical or Industrial Facilities (1991) 112 Calibration of Survey Instruments Used in Radiation Protection for the Assessment of Ionizing Radiation Fields and Radioactive Surface Contamination (1991) 113 Exposure Criteria for Medical Diagnostic Ultrasound: I. Criteria Based on Thermal Mechanisms (1992) 114 Maintaining Radiation Protection Records (1992) 115 Risk Estimates for Radiation Protection (1993) 116 Limitation of Exposure to Ionizing Radiation (1993) 117 Research Needs for Radiation Protection (1993) 118 Radiation Protection in the Mineral Extraction Industry (1993) 119 A Practical Guide to the Determination of Human Exposure to Radiofrequency Fields (1993) 120 Dose Control at Nuclear Power Plants (1994) 121 Principles and Application of Collective Dose in Radiation Protection (1995) 122 Use of Personal Monitors to Estimate Effective Dose Equivalent and Effective Dose to Workers for External Exposure to Low-LET Radiation (1995)

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123 Screening Models for Releases of Radionuclides to Atmosphere, Surface Water, and Ground (1996) 124 Sources and Magnitude of Occupational and Public Exposures from Nuclear Medicine Procedures (1996) 125 Deposition, Retention and Dosimetry of Inhaled Radioactive Substances (1997) 126 Uncertainties in Fatal Cancer Risk Estimates Used in Radiation Protection (1997) 127 Operational Radiation Safety Program (1998) 128 Radionuclide Exposure of the Embryo/Fetus (1998) 129 Recommended Screening Limits for Contaminated Surface Soil and Review of Factors Relevant to Site-Specific Studies (1999) 130 Biological Effects and Exposure Limits for “Hot Particles” (1999) 131 Scientific Basis for Evaluating the Risks to Populations from Space Applications of Plutonium (2001) 132 Radiation Protection Guidance for Activities in Low-Earth Orbit (2000) 133 Radiation Protection for Procedures Performed Outside the Radiology Department (2000) 134 Operational Radiation Safety Training (2000) 135 Liver Cancer Risk from Internally-Deposited Radionuclides (2001) 136 Evaluation of the Linear-Nonthreshold Dose-Response Model for Ionizing Radiation (2001) 137 Fluence-Based and Microdosimetric Event-Based Methods for Radiation Protection in Space (2001) 138 Management of Terrorist Events Involving Radioactive Material (2001) 139 Risk-Based Classification of Radioactive and Hazardous Chemical Wastes (2002) 140 Exposure Criteria for Medical Diagnostic Ultrasound: II. Criteria Based on all Known Mechanisms (2002) 141 Managing Potentially Radioactive Scrap Metal (2002) 142 Operational Radiation Safety Program for Astronauts in Low-Earth Orbit: A Basic Framework (2002) 143 Management Techniques for Laboratories and Other Small Institutional Generators to Minimize Off-Site Disposal of Low-Level Radioactive Waste (2003) 144 Radiation Protection for Particle Accelerator Facilities (2003) 145 Radiation Protection in Dentistry (2003) 146 Approaches to Risk Management in Remediation of Radioactively Contaminated Sites (2004) 147 Structural Shielding Design for Medical X-Ray Imaging Facilities (2004) 148 Radiation Protection in Veterinary Medicine (2004) 149 A Guide to Mammography and Other Breast Imaging Procedures (2004) 150 Extrapolation of Radiation-Induced Cancer Risks from Nonhuman Experimental Systems to Humans (2005) 151 Structural Shielding Design and Evaluation for Megavoltage X- and Gamma-Ray Radiotherapy Facilities (2005) 152 Performance Assessment of Near-Surface Facilities for Disposal of Low-Level Radioactive Waste (2005)

452 / NCRP PUBLICATIONS Binders for NCRP reports are available. Two sizes make it possible to collect into small binders the “old series” of reports (NCRP Reports Nos. 8–30) and into large binders the more recent publications (NCRP Reports Nos. 32–152). Each binder will accommodate from five to seven reports. The binders carry the identification “NCRP Reports” and come with label holders which permit the user to attach labels showing the reports contained in each binder. The following bound sets of NCRP reports are also available: Volume I. NCRP Reports Nos. 8, 22 Volume II. NCRP Reports Nos. 23, 25, 27, 30 Volume III. NCRP Reports Nos. 32, 35, 36, 37 Volume IV. NCRP Reports Nos. 38, 40, 41 Volume V. NCRP Reports Nos. 42, 44, 46 Volume VI. NCRP Reports Nos. 47, 49, 50, 51 Volume VII. NCRP Reports Nos. 52, 53, 54, 55, 57 Volume VIII. NCRP Report No. 58 Volume IX. NCRP Reports Nos. 59, 60, 61, 62, 63 Volume X. NCRP Reports Nos. 64, 65, 66, 67 Volume XI. NCRP Reports Nos. 68, 69, 70, 71, 72 Volume XII. NCRP Reports Nos. 73, 74, 75, 76 Volume XIII. NCRP Reports Nos. 77, 78, 79, 80 Volume XIV. NCRP Reports Nos. 81, 82, 83, 84, 85 Volume XV. NCRP Reports Nos. 86, 87, 88, 89 Volume XVI. NCRP Reports Nos. 90, 91, 92, 93 Volume XVII. NCRP Reports Nos. 94, 95, 96, 97 Volume XVIII. NCRP Reports Nos. 98, 99, 100 Volume XIX. NCRP Reports Nos. 101, 102, 103, 104 Volume XX. NCRP Reports Nos. 105, 106, 107, 108 Volume XXI. NCRP Reports Nos. 109, 110, 111 Volume XXII. NCRP Reports Nos. 112, 113, 114 Volume XXIII. NCRP Reports Nos. 115, 116, 117, 118 Volume XXIV. NCRP Reports Nos. 119, 120, 121, 122 Volume XXV. NCRP Report No. 123I and 123II Volume XXVI. NCRP Reports Nos. 124, 125, 126, 127 Volume XXVII. NCRP Reports Nos. 128, 129, 130 Volume XXVIII. NCRP Reports Nos. 131, 132, 133 Volume XXIX. NCRP Reports Nos. 134, 135, 136, 137 Volume XXX. NCRP Reports Nos. 138, 139 Volume XXXI. NCRP Report No. 140 Volume XXXII. NCRP Reports Nos. 141, 142, 143 Volume XXXIII. NCRP Report No. 144 Volume XXXIV. NCRP Reports Nos. 145, 146, 147 Volume XXXV. NCRP Reports Nos. 148, 149 (Titles of the individual reports contained in each volume are given previously.)

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NCRP Commentaries No. 1

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5 6 7 8 9 10 11 12 13 14 15

16 17 18 19

Title Krypton-85 in the Atmosphere—With Specific Reference to the Public Health Significance of the Proposed Controlled Release at Three Mile Island (1980) Guidelines for the Release of Waste Water from Nuclear Facilities with Special Reference to the Public Health Significance of the Proposed Release of Treated Waste Waters at Three Mile Island (1987) Review of the Publication, Living Without Landfills (1989) Radon Exposure of the U.S. Population—Status of the Problem (1991) Misadministration of Radioactive Material in Medicine—Scientific Background (1991) Uncertainty in NCRP Screening Models Relating to Atmospheric Transport, Deposition and Uptake by Humans (1993) Considerations Regarding the Unintended Radiation Exposure of the Embryo, Fetus or Nursing Child (1994) Advising the Public about Radiation Emergencies: A Document for Public Comment (1994) Dose Limits for Individuals Who Receive Exposure from Radionuclide Therapy Patients (1995) Radiation Exposure and High-Altitude Flight (1995) An Introduction to Efficacy in Diagnostic Radiology and Nuclear Medicine (Justification of Medical Radiation Exposure) (1995) A Guide for Uncertainty Analysis in Dose and Risk Assessments Related to Environmental Contamination (1996) Evaluating the Reliability of Biokinetic and Dosimetric Models and Parameters Used to Assess Individual Doses for Risk Assessment Purposes (1998) Screening of Humans for Security Purposes Using Ionizing Radiation Scanning Systems (2003) Pulsed Fast Neutron Analysis System Used in Security Surveillance (2003) Biological Effects of Modulated Radiofrequency Fields (2003) Key Elements of Preparing Emergency Responders for Nuclear and Radiological Terrorism (2005)

Proceedings of the Annual Meeting No. 1 3

4

5

Title Perceptions of Risk, Proceedings of the Fifteenth Annual Meeting held on March 14-15, 1979 (including Taylor Lecture No. 3) (1980) Critical Issues in Setting Radiation Dose Limits, Proceedings of the Seventeenth Annual Meeting held on April 8-9, 1981 (including Taylor Lecture No. 5) (1982) Radiation Protection and New Medical Diagnostic Approaches, Proceedings of the Eighteenth Annual Meeting held on April 6-7, 1982 (including Taylor Lecture No. 6) (1983) Environmental Radioactivity, Proceedings of the Nineteenth Annual Meeting held on April 6-7, 1983 (including Taylor Lecture No. 7) (1983)

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7 8

9

10 11

12

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14

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23

Some Issues Important in Developing Basic Radiation Protection Recommendations, Proceedings of the Twentieth Annual Meeting held on April 4-5, 1984 (including Taylor Lecture No. 8) (1985) Radioactive Waste, Proceedings of the Twenty-first Annual Meeting held on April 3-4, 1985 (including Taylor Lecture No. 9)(1986) Nonionizing Electromagnetic Radiations and Ultrasound, Proceedings of the Twenty-second Annual Meeting held on April 2-3, 1986 (including Taylor Lecture No. 10) (1988) New Dosimetry at Hiroshima and Nagasaki and Its Implications for Risk Estimates, Proceedings of the Twenty-third Annual Meeting held on April 8-9, 1987 (including Taylor Lecture No. 11) (1988) Radon, Proceedings of the Twenty-fourth Annual Meeting held on March 30-31, 1988 (including Taylor Lecture No. 12) (1989) Radiation Protection Today—The NCRP at Sixty Years, Proceedings of the Twenty-fifth Annual Meeting held on April 5-6, 1989 (including Taylor Lecture No. 13) (1990) Health and Ecological Implications of Radioactively Contaminated Environments, Proceedings of the Twenty-sixth Annual Meeting held on April 4-5, 1990 (including Taylor Lecture No. 14) (1991) Genes, Cancer and Radiation Protection, Proceedings of the Twenty-seventh Annual Meeting held on April 3-4, 1991 (including Taylor Lecture No. 15) (1992) Radiation Protection in Medicine, Proceedings of the Twenty-eighth Annual Meeting held on April 1-2, 1992 (including Taylor Lecture No. 16) (1993) Radiation Science and Societal Decision Making, Proceedings of the Twenty-ninth Annual Meeting held on April 7-8, 1993 (including Taylor Lecture No. 17) (1994) Extremely-Low-Frequency Electromagnetic Fields: Issues in Biological Effects and Public Health, Proceedings of the Thirtieth Annual Meeting held on April 6-7, 1994 (not published). Environmental Dose Reconstruction and Risk Implications, Proceedings of the Thirty-first Annual Meeting held on April 12-13, 1995 (including Taylor Lecture No. 19) (1996) Implications of New Data on Radiation Cancer Risk, Proceedings of the Thirty-second Annual Meeting held on April 3-4, 1996 (including Taylor Lecture No. 20) (1997) The Effects of Pre- and Postconception Exposure to Radiation, Proceedings of the Thirty-third Annual Meeting held on April 2-3, 1997, Teratology 59, 181–317 (1999) Cosmic Radiation Exposure of Airline Crews, Passengers and Astronauts, Proceedings of the Thirty-fourth Annual Meeting held on April 1-2, 1998, Health Phys. 79, 466–613 (2000) Radiation Protection in Medicine: Contemporary Issues, Proceedings of the Thirty-fifth Annual Meeting held on April 7-8, 1999 (including Taylor Lecture No. 23) (1999) Ionizing Radiation Science and Protection in the 21st Century, Proceedings of the Thirty-sixth Annual Meeting held on April 5-6, 2000, Health Phys. 80, 317–402 (2001) Fallout from Atmospheric Nuclear Tests—Impact on Science and Society, Proceedings of the Thirty-seventh Annual Meeting held on April 4-5, 2001, Health Phys. 82, 573–748 (2002)

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Where the New Biology Meets Epidemiology: Impact on Radiation Risk Estimates, Proceedings of the Thirty-eighth Annual Meeting held on April 10-11, 2002, Health Phys. 85, 1–108 (2003) Radiation Protection at the Beginning of the 21st Century–A Look Forward, Proceedings of the Thirty-ninth Annual Meeting held on April 9–10, 2003, Health Phys. 87, 237–319 (2004) Advances in Consequence Management for Radiological Terrorism Events, Proceedings of the Fortieth Annual Meeting held on April 14–15, 2004, Health Phys. 89, 415–588 (2005)

Lauriston S. Taylor Lectures No. 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16

Title The Squares of the Natural Numbers in Radiation Protection by Herbert M. Parker (1977) Why be Quantitative about Radiation Risk Estimates? by Sir Edward Pochin (1978) Radiation Protection—Concepts and Trade Offs by Hymer L. Friedell (1979) [available also in Perceptions of Risk, see above] From “Quantity of Radiation” and “Dose” to “Exposure” and “Absorbed Dose”—An Historical Review by Harold O. Wyckoff (1980) How Well Can We Assess Genetic Risk? Not Very by James F. Crow (1981) [available also in Critical Issues in Setting Radiation Dose Limits, see above] Ethics, Trade-offs and Medical Radiation by Eugene L. Saenger (1982) [available also in Radiation Protection and New Medical Diagnostic Approaches, see above] The Human Environment—Past, Present and Future by Merril Eisenbud (1983) [available also in Environmental Radioactivity, see above] Limitation and Assessment in Radiation Protection by Harald H. Rossi (1984) [available also in Some Issues Important in Developing Basic Radiation Protection Recommendations, see above] Truth (and Beauty) in Radiation Measurement by John H. Harley (1985) [available also in Radioactive Waste, see above] Biological Effects of Non-ionizing Radiations: Cellular Properties and Interactions by Herman P. Schwan (1987) [available also in Nonionizing Electromagnetic Radiations and Ultrasound, see above] How to be Quantitative about Radiation Risk Estimates by Seymour Jablon (1988) [available also in New Dosimetry at Hiroshima and Nagasaki and its Implications for Risk Estimates, see above] How Safe is Safe Enough? by Bo Lindell (1988) [available also in Radon, see above] Radiobiology and Radiation Protection: The Past Century and Prospects for the Future by Arthur C. Upton (1989) [available also in Radiation Protection Today, see above] Radiation Protection and the Internal Emitter Saga by J. Newell Stannard (1990) [available also in Health and Ecological Implications of Radioactively Contaminated Environments, see above] When is a Dose Not a Dose? by Victor P. Bond (1992) [available also in Genes, Cancer and Radiation Protection, see above] Dose and Risk in Diagnostic Radiology: How Big? How Little? by Edward W. Webster (1992) [available also in Radiation Protection in Medicine, see above]

456 / NCRP PUBLICATIONS 17 18 19 20 21 22 23 24 25 26 27 28

Science, Radiation Protection and the NCRP by Warren K. Sinclair (1993) [available also in Radiation Science and Societal Decision Making, see above] Mice, Myths and Men by R.J. Michael Fry (1995) Certainty and Uncertainty in Radiation Research by Albrecht M. Kellerer. Health Phys. 69, 446–453 (1995) 70 Years of Radiation Genetics: Fruit Flies, Mice and Humans by Seymour Abrahamson. Health Phys. 71, 624–633 (1996) Radionuclides in the Body: Meeting the Challenge by William J. Bair. Health Phys. 73, 423–432 (1997) From Chimney Sweeps to Astronauts: Cancer Risks in the Work Place by Eric J. Hall. Health Phys. 75, 357–366 (1998) Back to Background: Natural Radiation and Radioactivity Exposed by Naomi H. Harley. Health Phys. 79, 121–128 (2000) Administered Radioactivity: Unde Venimus Quoque Imus by S. James Adelstein. Health Phys. 80, 317–324 (2001) Assuring the Safety of Medical Diagnostic Ultrasound by Wesley L. Nyborg. Health Phys. 82, 578–587 (2002) Developing Mechanistic Data for Incorporation into Cancer and Genetic Risk Assessments: Old Problems and New Approaches by R. Julian Preston. Health Phys. 85, 4–12 (2003) The Evolution of Radiation Protection–From Erythema to Genetic Risks to Risks of Cancer to ? by Charles B. Meinhold, Health Phys. 87, 240–248 (2004) Radiation Protection in the Aftermath of a Terrorist Attack Involving Exposure to Ionizing Radiation by Abel J. Gonzalez, Health Phys. 89, 418–446 (2005)

Symposium Proceedings No. 1 2 3 4 5

Title The Control of Exposure of the Public to Ionizing Radiation in the Event of Accident or Attack, Proceedings of a Symposium held April 27-29, 1981 (1982) Radioactive and Mixed Waste—Risk as a Basis for Waste Classification, Proceedings of a Symposium held November 9, 1994 (1995) Acceptability of Risk from Radiation—Application to Human Space Flight, Proceedings of a Symposium held May 29, 1996 (1997) 21st Century Biodosimetry: Quantifying the Past and Predicting the Future, Proceedings of a Symposium held February 22, 2001, Radiat. Prot. Dosim. 97(1), (2001) National Conference on Dose Reduction in CT, with an Emphasis on Pediatric Patients, Summary of a Symposium held November 6-7, 2002, Am. J. Roentgenol. 181(2), 321–339 (2003)

NCRP Statements No. 1 2

Title “Blood Counts, Statement of the National Committee on Radiation Protection,” Radiology 63, 428 (1954) “Statements on Maximum Permissible Dose from Television Receivers and Maximum Permissible Dose to the Skin of the Whole

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Body,” Am. J. Roentgenol., Radium Ther. and Nucl. Med. 84, 152 (1960) and Radiology 75, 122 (1960) X-Ray Protection Standards for Home Television Receivers, Interim Statement of the National Council on Radiation Protection and Measurements (1968) Specification of Units of Natural Uranium and Natural Thorium, Statement of the National Council on Radiation Protection and Measurements (1973) NCRP Statement on Dose Limit for Neutrons (1980) Control of Air Emissions of Radionuclides (1984) The Probability That a Particular Malignancy May Have Been Caused by a Specified Irradiation (1992) The Application of ALARA for Occupational Exposures (1999) Extension of the Skin Dose Limit for Hot Particles to Other External Sources of Skin Irradiation (2001) Recent Applications of the NCRP Public Dose Limit Recommendation for Ionizing Radiation (2004)

Other Documents The following documents were published outside of the NCRP report, commentary and statement series: Somatic Radiation Dose for the General Population, Report of the Ad Hoc Committee of the National Council on Radiation Protection and Measurements, 6 May 1959, Science 131 (3399), February 19, 482–486 (1960) Dose Effect Modifying Factors in Radiation Protection, Report of Subcommittee M-4 (Relative Biological Effectiveness) of the National Council on Radiation Protection and Measurements, Report BNL 50073 (T-471) (1967) Brookhaven National Laboratory (National Technical Information Service, Springfield, Virginia) Residential Radon Exposure and Lung Cancer Risk: Commentary on Cohen's County-Based Study, Health Phys. 87(6), 656–658 (2004)

Index 222, 311, 355, 361, 362, 363, 364, 374, 387, 388, 391, 393 As low as reasonably achievable (ALARA) 35, 54, 374 Atomic Energy Act (AEA) 32, 57, 373, 374, 384, 385, 389 Autocatalysis 375

Absorbed dose 73, 167, 284, 373, 376, 378, 379, 380, 388, 389 Accessible environments 202, 316, 336, 366, 368, 369, 371, 374, 386 Accuracy 78, 80, 100, 196, 218, 219, 320, 373, 375, 377, 378, 383 bias 78, 80, 373, 375, 377, 383 precision 320, 373, 378 Activation 33, 373 Activity 15, 16, 29, 34, 47, 48, 51, 55, 57, 58, 65, 72, 73, 145, 146, 147, 157, 159, 168, 173, 174, 177, 178, 207, 214, 217, 223, 230, 252, 254, 256, 257, 261, 263, 266, 268, 269, 271, 274, 276, 278, 284, 292, 308, 315, 365, 373, 375, 377, 379, 383, 388, 391, 392 becquerel (Bq) 373, 375 curie (Ci) 373, 377 AEA 32, 34, 47, 57, 70, 72, 73, 374 Agreement State 19, 30, 34, 47, 54, 72, 145, 293, 304, 373, 374, 376, 384 ALARA 35, 50, 51, 54, 56, 57, 68, 221, 365, 373, 374, 384, 385, 389 Amorphous 181, 373 Anaerobic 167, 373 Analytical solution 79, 161, 163, 164, 169, 196, 216, 219, 327, 334, 373 Anion 146, 373, 384 Annual dose equivalent 49, 50, 51, 52, 58, 374 Aquaitard 203, 374 Aquatic foodchain pathways 258, 272, 274, 275, 278, 279 Aquifer 4, 7, 91, 115, 117, 171, 184, 185, 195, 197, 198, 199, 200, 201, 202, 203, 205, 213, 220, 221,

Background radiation 374 Barrier 2, 4, 7, 8, 13, 16, 17, 20, 21, 40, 42, 44, 48, 56, 59, 60, 69, 72, 75, 86, 87, 88, 93, 117, 118, 119, 121, 123, 127, 130–143, 144, 148, 151, 153, 156–158, 168–172, 176, 179, 181, 235, 242, 243, 245, 286, 293, 296–299, 301, 302, 305, 306, 307, 309, 310, 315, 316, 318, 319, 330, 353, 355, 356, 359, 360, 362, 366, 374, 378, 381, 385 Bayesian probability theory 374 Becequerel 375 Benchmarking 79, 334, 356, 375 Beta radiation 375, 383, 395 Bias 78, 80, 81, 82, 88, 100, 101, 106, 196, 215, 218, 219, 248, 259, 260, 280, 306, 320, 342, 343, 368, 369, 373, 375, 377, 378, 383, 391 accuracy 100, 196, 218, 219, 373 conservative bias 78, 80, 81, 82, 106, 215, 248, 259, 280, 306, 342, 343, 368, 369, 375, 377 imprecision 378 precision 320, 373, 378 Bioaccumulation factor 78, 275, 276, 277, 278, 375, 376 concentration factor 275, 278, 376

458

INDEX /

concentration ratio 275, 276, 277, 376 Biokinetic model 261, 262, 263, 265–267, 272, 375 Biosphere 38, 39, 44, 69, 112, 114, 117, 253, 375 Boundary condition 375 Byproduct material 32, 70, 373, 374, 375, 376, 384, 389 Cap 3, 16, 148, 154, 155, 296, 315, 375, 377, 387, 394 cover 148, 154, 155, 377 tumulus 394 Capillary fringe 186, 375 volumetric water content 375 Cation 145, 375 Cement 131, 135, 148, 159, 164, 170, 176, 177, 181, 375, 377 (see concrete) CERCLA 34, 57, 64, 68, 302, 309, 376, 393 SARA 376 Superfund 34, 302, 309, 376, 393 Characteristic curve 188, 189, 190, 191, 195, 196, 328, 375 Chelate 181, 376 Chemisorption 206, 376 Colloid 151, 152, 178, 179, 181, 182, 215, 376 Commercial waste 376 Committed dose 3, 49, 50, 51, 52, 54, 57, 58, 73, 167, 254, 261–264, 266, 268–272, 283, 284, 309, 373, 374, 376, 378, 379, 380, 388, 389, 391, 393 absorbed dose 73, 167, 284, 373, 378, 379, 380, 388, 389 dose equivalent 3, 49, 50, 51, 52, 54, 57, 58, 254, 261, 262, 268, 270, 374, 379, 380, 388, 389, 391 effective dose equivalent 3, 50, 51, 52, 54, 57, 254, 261–264, 266, 268–272, 309, 374, 376, 379, 393

459

equivalent dose 3, 261, 262, 264, 266, 269, 270, 374, 376, 378, 379, 380, 388, 389, 391 Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA) 34, 376, 393 Compressive strength 132, 134, 136, 376 Concentration factor 275, 278, 376 Concentration ratio 275, 276, 277, 376 Conceptual model 2, 8, 21, 22, 23, 24, 78, 84, 86, 88, 89, 91, 92, 93, 95, 99, 103–108, 144, 159, 160, 161, 171, 173, 196, 204, 205, 323, 325, 326, 327, 332, 333, 334, 346, 350, 354, 355, 356, 362, 368, 369, 372, 376 Concrete 4, 8, 16, 34, 48, 56, 93, 110, 115, 117, 128, 130–143, 147, 148, 153–156, 169, 171, 181, 241, 296, 299, 307, 329, 335, 356, 358, 360, 376, 378, 381, 394 (see cement) grout 8, 16, 34, 48, 128, 134, 169, 378, 381 Conservative bias 80, 106, 215 Containment 4, 59, 60, 61, 66, 130, 286, 377 Controlled area 373, 377, 378 Convection 198, 377 Coordination complex 377, 383 Corrosion 135, 137, 138, 139, 141, 143, 150, 156, 157, 158, 165, 167, 181, 235, 277 Cover 3, 16, 148, 154, 155, 296, 315, 375, 377, 387, 394 cap 3, 16, 148, 296, 315, 375, 387, 394 Creep 156, 158, 377 Critical group 36, 37, 38, 39, 67, 377 Critical organ 377 Curie (Ci) 15, 16, 29, 34, 47, 48, 51, 55, 57, 58, 65, 72, 73, 145, 146, 147, 157, 159, 168, 173, 174, 177,

460 / INDEX 178, 207, 214, 217, 223, 230, 252, 254, 257, 261, 253, 266, 268, 269, 271, 274, 276, 278, 284, 292, 308, 315, 365, 373, 375, 377, 379, 383, 388, 391, 392 activity 15, 16, 29, 34, 47, 48, 51, 55, 57, 58, 65, 72, 73, 145, 146, 147, 157, 159, 168, 173, 174, 177, 178, 207, 214, 217, 223, 230, 252, 254, 257, 261, 253, 266, 268, 269, 271, 274, 276, 278, 284, 292, 308, 315, 365, 373, 375, 377, 379, 383, 388, 391, 392 Darcy velocity 80, 127, 194, 198, 199, 208, 365, 377 Defense waste 377 Depassivation 137, 138, 139, 143, 377 Deposition velocity 278, 377 Desorption 180, 377, 391 Deterministic effects 377 Deterministic methods 331, 338, 347, 350, 377 Detriment 35, 37, 264, 283, 284, 377, 393 Differential analysis 345, 378 Diffusion 9, 136, 138, 139, 141, 159, 160, 162, 163, 164, 165, 167–171, 183, 206, 209, 210, 211, 213, 214, 220–227, 234, 235, 236, 237, 364, 365, 378, 381, 384, 386 Diffusion coefficient 135, 159, 163, 164, 165, 209, 210, 212, 235, 236, 378, 386 Diffusivity 138, 139, 378 Dispersion 8, 49, 168, 169, 171, 183, 184, 191, 197, 206, 210–216, 218–225, 238, 239, 325, 330, 364, 365, 378 Dispersion coefficient 211, 212, 213, 214, 220, 224, 365, 378 Dispersivity 212, 213, 225, 325, 329, 330, 378 Disposal cell 147, 149, 168, 170, 172, 378

Disposal facility 3, 4, 7, 11, 13, 16–20, 22, 24, 25, 26, 32, 38, 40, 41, 45, 46, 52, 54, 55, 56, 60, 62, 65, 68, 71, 74, 75, 76, 81, 82, 83, 86, 87, 88, 90, 91, 94, 95, 97, 99, 100, 103, 104, 106, 107, 108, 111, 112, 114, 115, 117, 118, 119, 121, 123, 131, 140, 143–154, 157, 167–185, 189, 192, 195, 197, 203, 215, 218, 223, 226, 235, 240, 241, 242, 245, 346, 350, 251, 253, 259, 281, 285, 289, 292, 295, 296, 297, 298, 299, 300, 302–316, 318, 319, 321, 324, 331, 335, 337, 353, 354, 357, 358, 359, 360, 361, 362, 366–372, 377, 380, 382, 383, 384, 386, 387, 391, 392, 393, 395 Disposal site 5, 6, 12, 15, 16, 20, 29, 34, 37, 42, 43, 46, 48, 50, 51, 52, 55, 61, 62, 64, 65, 66, 71, 72, 74, 75, 78, 80, 91, 99, 103, 104, 105, 110, 114, 123, 155, 156, 159, 176, 177, 196, 202, 205, 207, 210, 212, 223, 227, 237, 238, 239, 241–246, 248, 249, 250, 251, 252, 258, 260, 278, 281, 282, 286–297, 299–309, 317, 318, 324, 327, 334, 336, 346, 351, 353, 354, 355, 366, 369, 370, 372, 376, 378, 382, 387 Disposal system 1, 2, 3, 6, 7, 8, 9, 13, 14, 16–22, 24, 25, 26, 27, 28, 30, 31, 38, 39, 42, 43, 44, 45, 47, 48, 55, 59, 66, 67, 68, 69, 75, 76, 77, 78, 80, 81, 82, 83, 84, 85, 86, 87, 88, 89, 90, 92, 93, 94, 96, 97, 101, 103, 104, 106–117, 121, 130, 131, 140, 154, 182, 183, 206, 208, 215, 219, 226, 241, 243, 244, 250, 280, 283, 286, 288, 291, 299, 300, 301, 305, 309, 315, 321, 322, 324, 326, 328, 329, 330, 332, 333, 334, 336, 342, 346, 348, 351, 352, 353, 354, 355, 358, 366, 368, 369, 370, 372, 374, 378, 386, 389, 390, 394 Disposal unit 7, 19, 30, 31, 46, 48, 59, 60, 103, 118, 127, 128, 149,

INDEX /

150, 155, 196, 291, 310, 311, 353, 378, 381, 383, 385 geologic repository 19, 30, 31, 46, 48, 59, 60, 196, 291, 378, 381, 383, 385 land disposal facility 378 near-surface disposal 378, 383 Distribution coefficient 7, 91, 92, 152, 180, 181, 206, 207, 209, 217, 222, 223, 275, 277, 328, 362, 364, 378 Dose 3, 6, 7, 14, 15, 23, 25, 35, 36, 37, 38, 40–46, 49, 50, 51, 52, 54, 56–69, 73, 75, 76, 81, 87, 88, 91, 100, 101, 105, 113, 114, 115, 146, 147, 152, 153, 158, 167, 173, 174, 175, 177, 202, 205, 208, 218, 221, 225, 226, 227, 229, 232, 235, 240, 242, 245, 246, 248, 249, 250, 251, 252, 253, 254, 255, 256, 257, 259, 260, 261, 262, 263, 264, 265, 266, 267, 268, 269, 270, 271, 272, 273, 275, 276, 277, 278, 279, 280, 281, 282, 283, 284, 285, 286, 287, 288, 289, 290, 291, 293, 294, 295, 300, 301, 302, 303, 304, 306, 307, 308, 309, 311, 312, 313, 314, 315, 316, 317, 318, 319, 321, 323, 324, 326, 332, 335, 337, 344, 346, 349, 357, 361, 364, 365, 366, 367, 373, 374, 376, 378, 379, 380, 384, 388, 389, 390, 391, 392, 393 absorbed dose 73, 167, 284, 373, 376, 378, 379, 380, 388, 389 collective dose 36, 40, 56, 57, 221 committed dose 58, 261, 263, 283, 374, 376, 379 dose assessments 14, 229, 235, 240, 248, 249, 250, 251, 254, 256, 267, 271, 275, 276, 277, 278, 279, 280, 285, 286, 287, 290, 293, 300, 301, 302, 306, 308, 309, 311, 312, 313, 314, 318, 319 dose coefficient 58, 88, 91, 101, 249, 253, 254, 255, 259, 260, 261, 262, 264, 265, 266, 267,

461

268, 269, 270, 271, 272, 273, 279, 282, 283, 285, 286, 379 dose commitment (see committed dose) dose equivalent 49, 50, 51, 52, 58, 261, 262, 268, 374, 378, 379, 380, 388, 389, 391 effective dose 15, 36, 37, 42, 65, 254, 262, 263, 264, 265, 266, 267, 268, 271, 272, 283, 284, 285, 313, 374, 376, 378, 379, 391, 393 effective dose equivalent 3, 50, 51, 52, 54, 57, 254, 261, 262, 263, 264, 266, 268, 269, 270, 271, 272, 283, 309, 374, 376, 379, 393 equivalent dose 3, 49, 50, 51, 52, 58, 261, 262, 268, 374, 376, 378, 379, 380, 388, 389, 391, 393 quality factor 379, 388 tissue weighting factor 264, 284, 379, 388, 393 Dose rate 115, 254, 256, 257, 269, 270, 284, 379, 389, 391 Dosimetric model 284, 379 Dynamic 21, 87, 126, 133, 236, 256, 266, 326, 379 Effective porosity 208, 209, 213, 379, 387 Eh 180, 183, 379 Embrittlement 158, 379 Encapsulation 48, 379, 394 Enhancement factor 228, 379 Environment 1, 3, 4, 5, 6, 11, 14, 16, 17, 18, 19, 25, 26, 28, 29, 34, 35, 41, 43, 47, 49, 50, 51, 55, 56, 57, 60, 62, 64, 68, 69, 70, 71, 72, 73, 75, 76, 91, 93, 94, 95, 97, 101, 103, 104, 110, 112, 115, 117, 118, 119, 130, 131, 133, 134, 140, 144, 147, 151, 152, 154, 156, 157, 158, 167, 177, 178, 180, 181, 182, 183, 197, 199, 202, 205, 208, 221, 229, 232, 234, 235, 243, 244, 245,

462 / INDEX 246–259, 268, 269, 270, 272, 280–286, 295, 301, 303, 305, 306, 307, 313, 315, 316, 318, 329, 330, 336, 353, 357, 359, 360, 366, 367, 368, 369, 373, 374, 376, 378, 379, 380, 383, 384, 385, 386, 389, 390, 392, 393 Environmental Protection Agency (EPA) 12, 49, 303, 393, 394 EPA 12, 49, 51–61, 63, 64, 65, 66, 68, 69, 71, 72, 74, 88, 126, 145, 219, 223, 224, 259, 261, 262, 263, 265, 266, 269, 270, 277, 284, 285, 303, 309, 316, 317, 336 Epoxy 139, 141, 380 Equilibrium 7, 92, 105, 141, 159, 160, 161, 162, 168, 217, 255, 256, 268, 277, 328, 364, 380, 386, 387 Error 79, 105, 125, 126, 185, 196, 197, 204, 239, 263, 265, 270, 327, 328, 329, 334, 356, 380, 395 Error function 239, 327, 380 Evaporation 118, 125, 126, 380 Exfiltration 151, 380 Exposure pathways 3, 5, 9, 14, 15, 17, 50, 51, 52, 57, 59, 91, 104, 112, 113, 117, 204, 205, 240, 244, 245, 246, 247, 248, 249, 250, 251–259, 261, 264, 269, 272, 273, 279, 280, 281, 283, 284, 285, 286, 297, 299, 300, 301, 303, 304, 305, 306, 308, 318, 350, 355, 356, 367, 368, 369, 370, 391, 393 External exposure 153, 245, 249, 252, 253, 254, 259, 260, 268, 269, 270, 271, 272, 273, 280, 282, 284, 285, 286, 296, 297, 298, 299, 300, 379, 380, 383, 394 Extrapolation 80, 158, 311, 380 Fault tree analysis 97, 380 Finite-difference method 217, 380 Finite-element method 217, 381 Fission product 30, 31, 145, 176, 381, 382

Flux 4, 69, 87, 114, 118, 139, 185, 209, 211, 213, 220, 222, 235, 237, 278, 358, 365, 378, 381, 387 Flux density 163, 164, 235, 377, 381 Foodchain pathways 254, 258, 260, 272, 274, 275, 277, 278, 279 aquatic foodchain pathways 258, 272, 274, 275, 278, 279 terrestrial foodchain pathways 254, 258, 260, 274, 277, 278 Fuel reprocessing 28, 30, 31, 381 Fuzzy set 381 Fuzzy set analysis 344, 381

Gamma radiation 381, 383 Gaussian plume model 238, 240, 381 Generic 11, 14, 19, 27, 42, 51, 52, 71, 72, 78, 85, 86, 88, 90, 91, 93, 94, 100, 113, 228, 229, 230, 240, 241, 249, 258, 259, 260, 272, 273, 274, 275, 277, 278, 279, 280, 285, 287, 288, 290, 293, 301, 305, 355, 358, 359, 364, 365, 366, 367, 370, 381 analysis 11, 52, 301 data 27, 78, 90, 91, 94, 100, 240, 258, 259, 260, 272, 273, 274, 275, 277, 278, 279, 280, 285, 287, 355, 358, 364, 365, 366, 367 models 93, 100, 113, 359, 366, 367, 370 parameters 228, 230 pathways 113 performance assessments 19, 51, 71, 72, 88, 113, 229, 241, 249, 272, 288, 293 reference level 42 scenarios 290, 293, 305, 367, 370 screening models 14 Geologic repository 19, 30, 31, 46, 48, 59, 60, 196, 291, 378, 381, 383, 385

INDEX /

Groundwater 3, 4, 5, 8, 9, 13, 14, 15, 41, 45, 55, 57, 59, 60, 64, 65, 66, 71, 72, 79, 80, 91, 92, 105, 112, 115, 116, 117, 121, 127, 140, 141, 144, 146, 154, 166, 177, 184, 187, 197–211, 213, 214, 215, 216, 217, 218, 219, 220, 221, 222, 225, 226, 227, 235, 244, 245, 246, 251, 252, 257, 281, 299, 300, 301, 306, 308, 330, 343, 355, 356, 357, 363, 364, 365, 366, 374, 381, 384, 386, 387 Grout 8, 16, 34, 46, 128, 134, 169, 378, 381 Half-life 126, 169, 220, 314, 365, 381 Hazard 9, 11, 26, 27, 29, 42, 43, 45, 67, 69, 70, 71, 72, 73, 145, 154, 249, 292, 354, 376, 381, 385, 388, 389, 390, 393, 394 Heuristic 327, 336, 381 High-level radioactive waste 246, 250, 381, 383, 386 Homolog 286, 382 Humic 181, 382 Hydraulic conductivity 8, 9, 16, 80, 119, 121, 123, 127, 132, 133, 136, 137, 141, 142, 143, 186, 188, 189, 190, 191, 193–199, 201, 202, 236, 237, 292, 343, 362, 363, 375, 377, 378, 382, 383, 386, 387, 388, 391, 393 permeability 16, 119, 121, 132, 136, 137, 141, 142, 143, 236, 237, 383, 386 pump test 198, 199, 387 slug test 201, 292, 363, 391 transmissivity 388, 393 Hydraulic head 7, 9, 79, 80, 116, 127, 133, 186, 187, 188, 190, 192, 193, 194, 198, 199, 203, 208, 362, 363, 377, 382, 387, 392 pressure head 7, 116, 127, 186, 187, 188, 190, 192, 193, 194, 203, 377, 382, 387, 392 Hysteresis 119, 179, 382

463

IAEA 33, 35, 56, 69, 85, 87, 104, 246, 253, 256, 278 Importance analysis 6, 14, 23, 320, 321, 322, 323, 324, 326, 330, 338, 339, 342, 344, 345, 350, 351, 352, 370, 371, 372, 382 Inadvertent intruder 5, 6, 45, 46, 48–56, 64, 76, 146, 173, 174, 226, 235, 240, 246, 248, 249, 250, 268, 288–314, 317, 318, 319, 366, 382 Infiltration 4, 7, 8, 48, 79, 81, 116–130, 132, 133, 134, 139–143, 148, 149, 150, 151, 153, 155, 169, 171, 182, 185, 186, 189, 191, 192, 195, 197, 242, 336, 337, 358, 359, 360, 361, 363, 382, 384, 386, 391 percolation 118, 386 Initial condition 7, 128, 147, 172, 199, 382 Institutional control 5, 6, 8, 15, 17, 22, 28, 38, 40, 41, 46, 49, 50, 51, 52, 62, 72, 74, 75, 76, 108, 125, 153, 244, 248, 289, 291, 292, 293, 294, 295, 301, 302, 303, 304, 305, 306, 307, 309, 310, 317, 358, 359, 369, 372, 377, 382 Integration 2, 3, 6, 17, 21, 23, 24, 92, 101, 109, 113, 161, 166, 205, 322, 354, 355, 357, 368, 372, 373, 379, 382, 390 Interception fraction 278, 382 Intermediate-level radioactive waste 154, 382 Internal exposure 261, 272, 282, 383 Intrinsic permeability 236, 383 Inventory 105, 114, 143, 144, 148, 150, 159, 160, 162, 164, 166, 168, 169, 172, 173, 174, 176, 177, 178, 179, 226, 383, 390 Isolation 36, 71, 154, 283, 286, 328, 354, 372, 378, 381, 383, 387, 394 Isomorphic 383 Isotropic 7, 214, 383 Kriging 201, 383

464 / INDEX Land disposal facility 378, 383 Latin hypercube sampling method 341, 343, 383, 385 Monte-Carlo method 343, 385 Leaching 135, 136, 137, 140, 143, 160, 277, 356, 383, 384 Ligand 151, 180, 183, 242, 383 Linear-energy transfer (LET) 383, 388, 389 LLRWPA 29, 30, 47, 384 Low-level radioactive waste 1, 11, 28–33, 47, 61, 70, 72, 110, 144, 353, 383, 384, 392 Low-Level Radioactive Waste Policy Act (LLRWPA) 1, 47, 61, 384, 392 Lysimeter 76, 127, 384 Mass loading 227, 228, 229, 230, 234, 240, 384 Matric potential 186, 384 suction pressure 186, 392 Maximally exposed individual 57, 384 Meat transfer coefficient 274, 275, 384 Migration 16, 89, 95, 166, 172, 181, 184, 185, 236, 245, 262, 374, 384, 393 Milk transfer coefficient 254, 255, 274, 275, 276, 279, 384 Mill tailings 1, 12, 29, 30, 31, 32, 34, 57, 64, 72, 123, 384, 395 Mixed bed 145, 384 Mixed waste 70, 72, 384 Models 1, 2, 4, 5, 6, 7, 8, 9, 12, 13, 14, 15, 17–28, 34, 43, 44, 45, 60, 61, 66, 67, 74, 76, 77, 78, 79, 80, 82, 84–287, 290, 304, 305, 306, 308, 312, 313, 315, 318, 319, 320, 321, 322, 323, 325–334, 336–372, 373, 375, 377, 378, 379, 380, 381, 382, 385, 386, 387, 380, 391, 392, 394, 395 benchmarking 79, 334, 356, 375 calibration 78, 80, 385

multiplicative-chain 95, 246, 255, 256, 285, 385 parameter 26, 76, 87, 249, 250, 259, 260, 274, 286, 304, 305, 306, 308, 318, 320, 325, 328, 330, 340, 352, 387 screening 14, 15, 17, 86, 103, 104, 198, 221, 232, 259, 356, 391 sensitivity analysis 6, 22, 320, 321, 322, 344, 345, 346, 352, 371, 378, 391 structure 26, 345, 385 uncertainty analysis 6, 23, 44, 45, 96, 197, 218, 320, 323, 324, 330, 331, 332, 333, 338, 339, 340, 342, 345, 346, 347, 348, 351, 370, 394 validation 28, 74, 76, 77, 253, 395 Modified mass loading model 227, 229, 234 Modulus of elasticity 385 Moisture content 8, 132, 162, 186–195, 197, 209, 213, 214, 230, 236, 362, 375, 385 Monitoring 2, 9, 18, 19, 22, 49, 72, 74, 75, 76, 106, 108, 170, 184, 198, 204, 364, 372, 377, 382, 385 Monte-Carlo analysis 241, 343, 383, 385 Latin hypercube sampling method 241, 383 Multiplicative-chain model 95, 246, 255, 256, 285, 385 NARM 31, 32, 72, 73, 376, 384, 385 Naturally occurring and accelerator-produced radioactive material (NARM) 32, 376, 384, 385, 385 Near-surface disposal 378, 383 Nuclear fuel cycle 1, 31, 32, 34, 385, 395 Nuclear Waste Policy Act (NWPA) 29, 30, 32, 33, 64, 385 Numerical solution 196, 218, 327, 386

INDEX /

Osmosis 215, 386 Oxidation 158, 379, 386, 389 Parameter 26, 76, 87, 249, 250, 259, 260, 274, 278, 304, 305, 308, 318, 320, 325, 328, 330, 332, 338, 340, 343, 344, 346, 352, 387 uncertainty 278, 325, 328, 330, 332, 338, 340, 343, 344, 346, 387 Partition coefficient 160, 161, 171, 176, 177, 386 Pathways 3, 5, 6, 9, 13, 14, 15, 17, 50, 51, 52, 56, 57, 59, 78, 91, 104, 112, 113, 114, 115, 117, 123, 142, 154, 155, 170, 204, 205, 207, 226, 227, 237, 240–261, 263, 264, 269, 272, 274, 275, 276, 277, 278, 279, 280, 281, 383, 284, 285, 286, 288, 290, 296, 297, 298, 299, 300, 301, 303, 304, 305, 306, 308, 318, 350, 355, 356, 367, 368, 369, 370, 386, 391, 393 Peclet number 220, 221, 386 Percolation 118, 386 (see infiltration) Performance assessment 1–10, 16–27, 88–109, 110–143, 386 models 110–143 Performance objective 1, 2, 3, 6, 8, 9, 11, 13, 18, 19, 20, 21, 22, 23, 24, 27, 28, 46, 49, 50, 51, 52, 54, 55, 56, 57, 59–69, 71, 73, 74, 81, 82, 83, 84, 85, 86, 92, 103, 105, 106, 109, 113, 153, 174, 182, 202, 205, 218, 246, 248, 249, 259, 261, 267, 268, 271, 272, 273, 291, 283, 286, 300, 315, 316, 317, 323, 329, 330, 331, 335, 337, 338, 339, 344, 346, 347, 348, 351, 352, 353, 354, 357, 360, 361, 363, 365, 366, 368, 369, 370, 371, 372, 382, 386, 389 Perturbation analysis 343, 386 Physisorption 206, 387 Piezometric surface 199, 387 Plant-to-soil concentration ratio 256, 274, 276, 277, 387

465

Pore velocity 194, 195, 220, 387 Porosity 127, 141, 142, 186, 192, 193, 208, 209, 210, 213, 214, 329, 379, 387 effective porosity 208, 209, 213, 379, 387 Possibilistic theory 344, 387 Possibility theory 387 Post closure 1, 2, 18, 19, 22, 42, 71, 108, 172, 178, 372, 386, 387 Precision in calculations 320, 373, 387 Pressure head 7, 9, 79, 80, 116, 127, 133, 186, 187, 188, 190, 192, 193, 194, 198, 199, 203, 208, 362, 363, 377, 382, 387, 392 hydraulic head 7, 116, 127, 186, 187, 188, 190, 192, 193, 194, 203, 377, 382, 387, 392 Probabilistic 39, 51, 66, 67, 97, 142, 153, 249, 283, 322, 324, 325, 331, 332, 341, 342, 343, 344, 346, 347, 348, 349, 350, 377, 387, 388, 392, 393, 394 stochastic 249, 283, 324, 325, 342, 343, 377, 388, 392, 393, 394 Probabilistic methods 331, 333, 340, 343, 349, 350, 351, 374, 387, 392 Bayesian probability theory 374 stochastic methods 343, 387, 392 Probabilistic risk assessment 97, 387 Probability distribution 51, 66, 277, 280, 328, 332, 339, 340, 341, 342, 343, 348, 349, 351, 383, 384, 385, 386, 387 Pump test 198, 199, 387 Quality assurance 42, 76, 78, 79, 84, 87, 96, 101, 143, 334, 356, 357, 399 Quality factor 379, 388 Radiation weighting factor 284, 380, 388

466 / INDEX Radiolysis 167, 388 Radon 3, 50, 52, 55, 58, 63, 64, 144, 154, 155, 167, 170, 235, 236, 237, 251, 263, 268, 269, 281, 366, 374, 388 progeny 388 RCRA 70, 71, 72, 73, 123, 384, 385, 389 Reasonable assurance 1, 2, 18, 20, 21, 22, 28, 44, 67, 74, 81, 82, 83, 84, 85, 86, 87, 89, 106, 107, 109, 146, 286, 321, 352, 353, 357, 372 386, 389 Redox 151, 180, 389 Reduction 39, 57, 68, 138, 148, 161, 208, 310, 312, 319, 326, 333, 334, 379, 389 Reference individual 249, 261, 262, 266, 271, 283, 389 Regression analysis 345, 389 Resource Recovery and Recovery Act (see RCRA) Response surface 345, 389 Resuspension factor model 227, 229, 230, 231, 232, 233, 240, 390 Resuspension rate model 227, 233, 234, 240, 390 Retardation coefficient 390 Risk assessment 37, 71, 97, 283, 387, 390 Risk coefficient 37, 254, 265, 283, 284, 285, 390 Robust 6, 41, 43, 45, 56, 87, 305, 357, 390 Safe Drinking Water Act (SDWA) 65, 390 Saturated zone 7, 9, 116, 186, 198, 199, 203, 219, 220, 305, 362, 363, 390 Saturation 4, 132, 133, 136, 184, 186, 188, 189, 198, 255, 361, 381, 390 Scaling factor 174, 361, 390 Scenario 6, 15, 38, 40, 41, 45, 46, 50, 52, 61, 67, 76, 101–109, 119, 123, 146, 154, 227, 229, 232, 234,

237, 242, 245, 246, 248, 249, 250, 251, 252, 253, 256, 258, 259, 267, 268, 269, 271, 272, 273, 285, 286, 287, 288, 289, 290, 292, 293, 294, 295–319, 331, 334, 336, 337, 338, 346, 367, 370, 377, 390, 391 Screening models 14, 15, 17, 86, 103, 104, 198, 221, 232, 259, 356, 391 Sensitivity analysis 6, 22, 320, 321, 322, 344, 345, 346, 352, 371, 378, 391, 392 step-wise regression 345, 392 Shallow land disposal 154, 291, 391 Shannon’s entropy 342, 391 Sievert (Sv) 379, 380, 391 Site-specific 1, 5, 8, 9, 14, 18, 19, 26, 27, 50, 54, 55, 56, 62, 64, 72, 73, 77, 78, 79, 90–95, 100, 103, 108, 113, 114, 119, 126, 171, 177, 180, 183, 189, 196, 197, 202, 212, 223, 226, 228, 230, 233, 241, 245, 249, 258, 259, 260, 272, 274, 277, 279, 285, 287, 288, 290, 293, 294, 301, 304, 305, 308, 312, 317, 318, 319, 326, 332, 333, 344, 355, 356, 358, 361, 362, 363, 364, 365, 366, 367, 369, 370, 381, 385, 386, 391 Sorption 7, 151, 152, 168, 169, 171, 173, 179, 180, 181, 183, 191, 206, 207, 209, 213, 214, 216, 221, 222, 225, 261, 266, 267, 282, 324, 325, 328, 359, 364, 373, 375, 376, 377, 381, 384, 387, 390, 391 isotherm 7, 214, 328, 391 Source material 31, 374, 384, 391, 392 Source term 5, 7, 8, 60, 61, 114, 115, 116, 117, 119, 121, 143–183, 185, 205, 217, 218, 234, 280, 312, 319, 360, 361, 362, 364, 391 Spallation 391 Special nuclear material 374, 375, 391, 392 Speciation 171, 180, 207, 392

INDEX /

Specific activity 174, 256, 257, 278, 281, 292 model 256, 257, 278, 281, 292 Specific storage 198, 388, 391, 392 Spent nuclear fuel 316, 327, 336, 381, 382, 383, 385, 392, 394, 395 State compacts 47, 61, 384, 392 Steady state 7, 9, 117, 121, 125, 127, 171, 179, 185, 191, 192, 195, 200, 203, 224, 226, 238, 240, 326, 336, 337, 338, 363, 365, 387, 392 Step-wise regression 345, 392 Stochastic 249, 283, 324, 325, 342, 343, 377, 378, 387, 388, 392, 394 effects 378, 388, 392 methods 343, 387, 392 variation 324, 392, 394 Strain 385, 392 Stress 132, 137, 157, 158, 199, 231, 232, 234, 377, 385, 392, 393, 395 Suction pressure 186, 392 Superfund 34, 57, 64, 68, 302, 309, 376, 393 CERCLA 34, 57, 64, 68, 302, 309, 376 Surveillance 2, 377, 383, 393 Suspension 181, 221, 227, 228, 229, 230, 231, 232, 233, 234, 240, 366, 376, 384, 389, 390, 393

Tensile strength 141, 393 Terrestrial foodchain pathways 254, 258, 260, 274, 277, 278 Tortuosity 210, 393 Toxic Substance Control Act (TSCA) 70, 385, 393 Transfer factors 253, 254, 255, 256, 258, 259, 260, 274, 275, 276, 278, 279, 280, 281, 285, 393 Translocation factor 278, 393 Transmissivity 388, 393 (see hydraulic conductivity) Transpiration 118, 119, 123, 125, 126, 151, 203, 251, 393

467

Transport 3, 4, 5, 6, 7, 8, 9, 13, 14, 15, 16, 17, 18, 26, 45, 60, 70, 75, 79, 89, 91, 92, 95, 103, 109, 112–118, 122, 123, 131, 140, 141, 142, 144, 148, 149, 151, 152, 153, 155, 161, 162, 166–173, 177, 179, 181, 183–196, 205–245, 262, 374, 384, 393 atmospheric transport 226–240 biotic transport 240–245 groundwater transport 205–220 migration 16, 89, 95, 166, 172, 181, 184, 185, 236, 245, 262, 374, 384, 393 surface water transport 206, 221–225 unsaturated zone transport 183–196 Transuranic waste 26, 29, 30, 31, 32, 47, 57, 60, 61, 65, 66, 74, 202, 316, 317, 336, 383, 393, 395 Treatment 9, 32, 48, 56, 59, 70, 71, 72, 86, 148, 158, 170, 353, 389, 394, 395 Tumulus 394

Uncertainty 278, 325, 328, 330, 332, 338, 340, 341, 343, 344, 346, 386, 387 perturbation analysis 343, 386 Undisturbed performance 55, 61, 288, 316, 394 Unsaturated zone 4, 7, 8, 18, 25, 76, 93, 116, 117, 121, 122, 132, 133, 143, 183–197, 206, 208, 209, 213, 214, 217, 219, 221, 235, 237, 328, 343, 359, 361, 362, 363, 364, 376, 392, 394 Usage factor 253–260, 272, 273, 279, 280, 281, 285, 395

Vadose zone (see unsaturated zone) Verification 79, 334, 395 Viscosity 133, 236, 383, 395

468 / INDEX Waste 1–10, 11–22, 24–34, 36, 37, 40, 46, 48, 50, 52, 53, 56, 59, 60, 64, 70, 72, 76, 77, 103, 106, 108, 110, 119, 134, 140, 144–153, 154, 157, 158, 159, 160, 161, 162, 163, 165–169, 171, 172, 173, 174, 175, 176, 177, 178, 180, 182, 183, 241, 242, 243, 288, 290, 292, 293, 294, 295, 296, 298, 299, 310, 311, 312, 314, 315, 319, 350, 360, 361, 370, 379, 383, 384, 386, 394, 395 classification system 5, 33, 48, 50, 52, 53, 56, 64, 145, 174, 288, 290, 292, 293, 294, 295, 296, 298, 299, 312, 370, 395 dilution factor 311, 314, 319, 395

form 4, 7, 8, 17, 20, 34, 52, 53, 60, 76, 108, 110, 119, 134, 140, 144–153, 154, 157, 158, 159, 160, 161, 162, 163, 165–169, 171, 172, 173, 174, 175, 176, 177, 178, 180, 182, 183, 241, 242, 243, 360, 361, 370, 379, 384, 386, 394, 395 form performance 160 management 28, 36, 37, 70, 72, 77, 176, 350, 395 package 7, 40, 46, 48, 56, 59, 60, 103, 106, 134, 147, 159, 177, 296, 310, 315, 383, 395 Weathering half-time 278, 395 Yield strength 137, 143, 395