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\-z "" EPA Risk Assessment 

Methodology 

Draft Environmental Impact 

Statement for Proposed 

ESHAPS for Radionuclides 

Volume 1 

Background Information 

Document 

40 CFR Part 61 

National Emission Standards 

for Hazardous Air Pollutants 

EPA 520/1-89-005 

Risk Assessment Methodology 

Draft Environmental Impact Statement 

for Proposed NESHAPS for Radionuclides 

VOLUME 1 

BACKGROUND INFORMATION DOCUMENT 

February 1989 

U.S. Environmental Protection Agency 

Office of Radiation Programs 

Washington, DC 20460 

The Environmental Protection Agency is proposing National 

Emission Standards for Hazardous Air Pollutants (NESHAPs) for 

Radionuclides. A Draft Environmental Impact Statement (DEIS) 

has been prepared in support of the proposed rulemaking. The 

DEIS consists of the following three volumes: 

Volume I - Risk Assessment Methodology 

This document contains chapters on hazard 

identification, movement of radionuclides through 

environmental pathways, radiation dosmetry, 

estimating the risk of health effects resulting 

from expose to low levels of ionizing radiation, 

and a summary of the uncertainties in calculations 

of dose and risks. 

Volume I1 - Risk Assessments 

This document contains a chapter on each 

radionuclide source category studied. The 

chapters include an introduction, category 

description, process description, control 

technology, health impact assessment, supplemental 

control technology, and cost. It has an appendix 

which contains the imputs to all the computer runs 

used to generate the risk assessment. 

Volume I11 - Economic Assessment 

This document has chapters on each radionuclide 

source category studied. Each chapter includes an 

introduction, industry profile, summary of 

emissions, risk levels, the benefits and costs of 

emission controls, and economic impact evaluations. 

Copies of the DEIS in whole or in part are avaiable to all 

interested persons; an announcement of the availability appears 

in the ~edeial Register. Comments on the DEIS should be sent in 

duplicate to: 

Central Docket Section (LE 131) 

Environmental Protection Agency 

Attention Docket No. A-79-11 

401 M Street, SW 

Washington, DC 20460 
For additional information, contact James Kardin at (202) 

475-9610 or write to: 

Director, Criteria and Standards Division 

Office of Radiation Programs (ANR-460) 

Environmental Protection Agency 

401 M Street, SW 

Washington, DC 20460 
LIST OF PREPARERS 

Various staff members from EPA- Office of Radiation 

Programs contributed in the development and preparation of the 

DEIS. 

Terrence McLaughlin Chief, Environmental 

Standards Branch 

James Hardin Health Physicist Project Officer 

Byron Bunger Economist Author/Reviewer 

Fran Cohen Attorney Advisor Reviewer 

Albert Colli 

Larry Gray 

Environmental 

Scientist 

Environmental 

Scientist 

W. Daniel Hendricks Environmental 

Scientist 

Paul Magno Environmental 

Scientist 

Christopher B. Nelson Environmental 

Scientist 

Reviewer 

Author 

Dr. Neal S. Nelson Radiobiologist Author 

Barry Parks Health Physicist Reviewer 

Dr. Jerome Puskin Chief BioeffecCs ~uthor/Reviewer 

Analysis Branch 

Jack L. Russell Engineer Author/Reviewer 

Dr. James. T. Walker Radiation 

Biophysicist 

Author 

Larry Weinstock Attorney Advisor Reviewer 

An EPA contractor, S. Cohen and Associates, Inc., McLean, 

VA, provided significant technical support in the preparation of 

the DEIS. 

TABLE OF CONTENTS 

VOLUME I: RISK ASSESSMENT METHODOLOGY 

LIST OF TABLES .................. vfii 

LIST OF FIGURES .................. xi 

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

2. CURRENT PROGRAMS AND STRATEGIES 

3. HAZARD IDENTIFICATION .............. 3-1 

4. MOVEMENT OF RADIONUCLIDES THROUGH ENVIRONMENTAL 

PATHWAYS .................. 4-1 

5. RADIATION DOSIMETRY ................ 5-1 

6. ESTIMATING THE RISK OF HEALTH EFFECTS 

RESULTING FROM EXPOSURE TO LOW LEVELS OF 

IONIZING RADIATION ................. 6-1 

7. SUMMARY OF UNCERTAINTIES IN DOSES AND RISKS . 7-1 

APPENDIX A ASSESSMENT METHODOLOGY .......... A-1 

APPENDIX B MECHANICS OF LIFE TABLE IMPLEMENTATION 

OF THE RISK ESTIMATES ........... B-1 

APPENDIX C OVERVIEW OF TECHNIQUES USED TO QUANTIFY 

UNCERTAINTY IN ENVIRONMENTAL RISK 

ASSESSMENTS ........... C-1 
LIST OF TABLES 

Quality Factor for Various Types of Radiation ... 

Weighting Factors Recommended by the ICRP 

for Stochastic Risks ............... 

Comparison of Customary and SI Special Units 

for Radiation Quantities ............ 

Target Organs and Tissues Used for Calculating the 

ICRP Effective Dose Equivalent and the EPA Cancer 

Risk, ...................... 

Comparative Estimate of Cancer Fatalities Induced 

by a Single 10-rad, Low-LET Radiation Exposure to 

the General Population .............. 

A Comparison of Estimates of the Risk of Fatal 

Cancer From Low-LET Radiation .......... 

Site Specific Incidence Risk Coefficients. .... 

Site Specific Mortality to Incidence Risk Ratios 

BEIR 111 L-L Model for Excess Fatal Cancers Other 

Than Leukemia and Bone Cancer ........... 

Mortality Risk Coefficients (lo-' per rad) for the 

Constrained Relative Risk Model .......... 

BEIR I11 L-L Model For Excess Incidence of (and 

Mortality from) Leukemia and Bone Cancer 

(absolute risk model) .............. 

Site Specific Mortality Risk Per Unit Dose 

per rad) for Combined Leukemia-Bone and 

Constrained Relative Risk Model .......... 

Site Specific Incidence Risk Per Unit Dose (lo-$ 

per rad) for Combined Leukemia-bone and 

Constrained Relative Risk Model .......... 

Site Specific Mortality Risk Per Unit Dose from 

Low-LET Radiation, Average for Both Sexes and All 

Ages ....................... 
Site-Specific Mortality Risk Per Unit Dose from 

Internally Deposited Alpha Emitters Averaged for 

Both Sexes and All Ages .............. 6-33 

Risk Estimate for Exposures to Radon Progeny .... 6-38 

Lifetime Risk From Radon Daughter Exposure Risk of 

Lung Cancer Death ................. 6-45 

Lifetime Risk From Excess Radon Daughter Exposure 

(adjusted for a Background Exposure of 0.25 WLM/yr) 6-47 

Lifetime Risk for Exposure of Varying Duration 

(Bkg = 0.25 WLM/yr) ................ 6-50 

Lifetime Risk for Varying Age at First Exposure and 

Duration of Exposure (Bkg = 0.25 WLM/yr) ...... 6-51 

BEIR I11 Estimates of Genetic Effects of An 

Average Population Exposure of 1 rem per 30-yr 

Generation (Chronic x-ray or gamma radiation 

exposure) ................... 6-62 

Summary of Genetic Risk Estimates Per 106 Liveborn 

of Low Dose Rate, Low-LET Radiation in a 30 Year 

Generation .................. 6-63 

Genetic Risk Estimates per 106 Liveborn for an 

Average Population Exposure of 1 Rad of High-LET 

Radiation in a 30-year Generation .... 6-64 

Radiation Induced Reciprocal Translocations in 

Several Species ............. 6-65 

Estimated Frequency of Genetic Disorders in a 

Birth Cohort Due to Exposure of the Parents to 1 

Rad per Generation ............. 6-68 

Increase in Background Level of Genetic Effects 

After 30 Generations or More .......... 6-70 

Causes of Uncertainty in the Genetic Risk Estimates 6-71 

Possible Effects of In Utero Radiation Exposure ... 6-79 

Summary of EPAfs Radiation Risk Factors ....... 6-81 

Presumed Sources of Food for Urban and 
AIRDOS-EPA Parameters Used Eor Generic Site 

Assessments .................... A-5 

Default Values Used for Element Dependent 

Factors ....................... A-7 

Cattle Densities and Vegetable Crop 

Distributions for use with AIRDOS-EPA . A-9 

Risk Factor for Selected Radionuclides . A-11 
LIST OF FIGURES 

Pathways of Airborne Radionuclides Into the 

Environment ............. 4-2 

Vertical Concentration Profiles For Plume vs 

Downwind Distance From Release .. 4-6 

Circular Grid System Used by AIRDOS-EPA ... 4-16 

Schematic Representation of Radioactivity 

Movement Among Respiratory Tract, 

Gastrointestinal Tract, and Blood ...... 5-13 

Typical Pattern of Decline of Activity of 

a Radionuclide in an Organ, Assuming an 

Initial Activity in the Organ and No 

Additional Uptake of Radionuclide by the 

Organ .................... 5-15 

The ICRP Task Group Lung Model for 

Particulates. ............. 5-20 

Example of the Output of a Risk Assessment 

Using Quantitative Uncertainty Analysis ... C-3 

1. INTRODUCTION 

1.1 HISTORY OF STANDARDS DEVELOPMENT 

In 1977, Congress amended the Clean Air Act (the Act) to 

address emissions of radioactive materials. Before 1977, these 

emissions were either regulated under the Atomic Energy Act or 

unregulated. Section 122 of the Act required the Administrator of 

the U.S. Environmental Protection Agency (EPA), after providing 

public notice and opportunity for public hearings (44 FR 21704, 

April 11, 1979), to determine whether emissions of radioactive 

pollutants cause or contribute to air pollution that may reasonably 

be expected to endanger public health. On December 27, 1979, EPA 

published a notice in the Federal Register listing radionuclides as 

hazardous air pollutants under Section 112 of the Act (44 FR 76738, 

December 27, 1979). To support this determination, EPA published a 

report entitled "Radiological Impact Caused by Emissions of 

Radionuclides into Air in the United States, Preliminary Report" 

(EPA 520/7-79-006, Office of Radiation Programs, U.S. EPA, 

Washington, D.C., August 1979) . 

On June 16, 1981, the Sierra Club filed suit in the U.S. 

District Court for the Northern District of California pursuant to 

the citizens' suit provision of the Act (Sierra Club v Gorsuch, No. 

81-2436 WTS). The suit alleged that EPA had a nondiscretionary duty 

to propose standards for radionuclides under Section 112 of the Act 

within 180 days after listing them. On September 30, 1982, the Court 

ordered EPA to publish proposed regulations establishing emissions 

standards for radionuclides, with a notice of hearing within 180 

days of the date of that order. 

On April 6, 1983, EPA published a notice in the Federal 

Register proposing standards for radionuclide emission sources in 

four categories: (1) DOE facilities, (2) Nuclear Regulatory 

Commission facilities, (3) underground uranium mines, and (4) 

elemental phosphorus plants. Several additional categories of 

sources that emit radionuclides were identified, but it was 

determined that there were good reasons for not proposing standards 

for them. These source categories were (1) coal-fired boilers; (2) 

the phosphate industry; (3) other mineral extraction industries; (4) 

uranium fuel cycle facilities, uranium tailings, and high-level 

waste management; and (5) low energy accelerators (48 FR 15077, 

April 6, 1983). To EPA's knowledge, these comprise the source 

categories that release potentially regulative amounts of 

radionuclides to the air. 

To support these proposed standards and determinations, EPA 

published a draft report entitled "Background Information 
* * * DRAFT * * * 

Document., Proposed Standards for Radionuclides" (EPA 520/1-83-001, 

Office of Radiation Programs, U.S. EPA, Washington, D.C., March 

1983). 

Following publication of the proposed standards, EPA held an 

informal public hearing in Washington, D.C., on April 28 and 29, 

1983. The comment period was held open an additional 30 days to 

receive written comments. Subsequently, EPA received a number of 

requests to extend the time for submission of public comments and to 

accommodate persons who were unable to attend the first public 

hearing. In response to these requests, EPA published a notice in 

the Federal Register that extended the comment period by an 

additional 45 days and held an additional informal public hearing in 

Denver, Colorado, on June 14, 1983 (48 FR 23655, May 26, 1983). 

On February 17, 1984, the Sierra Club again filed suit in the A U.S. District Court for the Northern District of California pursuant 

to the citizens' suit provision of the Act (Sierra Club v 

- Ruckelshaus, No. 84-0656 WHO). The suit alleged that EPA had a 

- nondiscretionary duty to issue final emissions standards for 

radionuclides or to find that they do not constitute a hazardous air 

pollutant (i.e., "de-list" the pollutant). In August 1984, the Court - granted the Sierra Club motion and ordered EPA to take final actions 

on radionuclides by October 23, 1984. 

. 

On October 22, 1984, the Agency issued its Background 

Information Document in support of the Agency's final action on 

radionuclides. The report contains an integrated risk assessment 

that provides the scientific basis for these actions (EPA 520/1-84- 

022-1). 

On February 6, 1985, National Emission Standards for Hazardous 

Air Pollutants (NESHAPS) were promulgated for radionuclide emissions 

from DOE facilities, NRC-licensed and non-DOE Federal facilities, 

and elemental phosphorus plants (50 FR 5190). Two additional 

radionuclide NESHAPS, covering radon-222 emissions from underground 

uranium mines and licensed uranium mill tailings, were promulgated 

on April 17, 1985 (50 FR 15386) and September 24, 1986 (51 FR 

34056), respectively. 

The EPA's basis for the radionuclide NESHAPS was challenged in 

lawsuits filed by the Sierra Club and the National Resources Defense 

Council (NRDC). While these suits were under adjudication, the U.S. 

Court of Appeals for the District of Columbia issued a decision 

finding that the EPA's NESHAP for vinyl chloride was defective in 

that costs had been improperly considered in setting the standard. 

Following the Court's order to review the potential effects of the 

vinyl chloride decision on other standards, the EPA determined that 

costs had been considered in many rulemakings on radionuclide 

emissions. On December 9, 1987, the Court accepted the EPA's 

proposal to leave the existing radionuclide NESHAPS in place while 
the Agency reconsidered the standards. in the interim, the suits 

filed by the Sierra Club and the NRDC have been placed in abeyance. 

1.2 PURPOSE OF THE FINAL BACKGROUND INFORMATION DOCUMENT 

Volume I contains background information on radiation 

protection programs and a detailed description of the Agency's 

procedures and methods for estimating radiation dose and risk due to 

radionuclide emissions to the air. This material is arranged as 

shown in the following descriptions of the chapters: 

o Chapter 2 - A summary of regulatory programs for 

radiation protection and the current positions of the 

various national and international advisory bodies and 

state and Federal agencies in regard to radiation. 

o Chapter 3 - A description of what makes radiation 

hazardous, the evidence that proves the hazard, and the 

evidence that relates the amount of radiation exposure to 

the amount of risk. 

o Chapter 4 - An explanation of how radionuclides, once 

released into the air, move through the environment and 

eventually cause radiation exposure of people. This 

chapter also contains a description of how EPA estimates 

the amounts of radionuclides in the environment, i.e., in 

the air, on surfaces, in the food chain, and in exposed 

humans. 

o Chapter 5 - A description of how radionuciides, once 

inhaled and ingested, move through the body to organs and 

expose these organs. This chapter also contains a 

description of how EPA estimates the amounts of radiation 

dose due to this radiation exposure of organs. It also 

describes how the amount of radiation dose is estimated 

when the source of radiation is gamma rays from a source 

outside of the body. 

o Chapter 6 - A description of how the risk of fatal 

cancers and genetic effects is estimated once the amount 

of radiation dose is known. 

o Chapter 7 - A summary of the uncertainties in the dose 

and risk estimates of source categories emitting 

significant amounts of radionuclides, which were made by 

using the procedures and information in the previous 

chapters. Associated uncertainties are discussed in the 

appropriate chapter, but overall uncertainties are 

discussed in this chapter. 
Volume I also contains three appendices. Appendix X describes 

the environmental transfer factors used in the dose assessment 

models. Appendix B describes the mechanics of the life table 

analysis used to estimate risk. Appendix C presents an overview of 

the quantitative uncertainty analysis techniques currently under 

review for use as a method for expanding the semiquantitative 

uncertainty analysis provided in Volume I. 

Volume I1 contains detailed risk estimates for each source of 

emissions, which were performed according to the procedures given in 

Volume I. Each chapter in Volume I1 addresses four topics: (1) the 

source category, the processes that result in releases of 

radionuclides to the environment, and existing controls, (2) the 

bases for the risk assessment, including reported emissions, source 

terms used, and other site parameters relevant to the dose 

assessment, (3) the results of the dose and risk calculation, along . with an extrapolation to the entire category, and (4) a description 

of supplementary emissions controls and their cost and effectiveness 

in reducing dose and risk. 

Two appendices are also provided in Volume 11. Appendix A 

presents the detailed AIRDOS input sheets used to calculate 

individual and population doses and risks associated with each 

category. Appendix B presents the methodology used to evaluate the 

. costs and effectiveness of earthen covers to control radon emissions 

from area sources of radon. 

1.3 UPDATE METHODOLOGY 

The categories of emissions addressed in this document are 

similar to those addressed in the 1984 Background Information 

Document. DOE and NRC-licensed facilities, elemental phosphorus 

plants, underground uranium mines, and licensed uranium mills are 

addressed because they are covered by NESHAPS. Uranium fuel cycle 

facilities, high-level waste disposal facilities, coal-fired 

boilers, and inactive uranium mill tailings sites are addressed 

because of challenges to previous determinations that they were 

adequately covered by other laws. Surface uranium mines, DOE radon, 

and phosphogypsum stacks are addressed because of challenges to the 

EPAfs lack of risk assessment for these facilities. In sum, this 

Background Information Document addresses the following categories 

of radiological emissions to air: 

DOE Facilities 

NRC-Licensed and Non-DOE Federal Facilities 

Uranium Fuel Cycle Facilities 

High-Level Waste 

Elemental Phosphorus Plants 

Coal-fired Boilers 

Inactive Uranium Mill Tailings 

Licensed Uranium Mill Tailings 
o DOE Itadon 

o Underground Uranium Mines 

o Surface Uranium Mines 

o Phosphogypsum Stacks 

For each category, Volume I1 presents updated information on 

the number of facilities, radionuclide emissions to air, and control 

technologies. Depending on the number of facilities in a category, 

risks are provided for individual facilities, or a set of reference 

facilities is defined that conservatively represents the category. 

Risks to the critical population group and the population within 80 

km are presented for each category. 

EPA recognizes that when it performed a risk assessment to 

determine the need for regulation of uranium mill tailings under the 

Uranium Mill Tailings Radiation Control Act (UMTRCA), the Agency 

considered the national health impact from the radon released from 

the tailings. In this assessment, EPA is considering only the 

health effects within 80 km of the source. EPA is using 80 km as 

the limit in order to be consistent with the other NESHAP 

rulemakings. This risk assessment in no way disputes the validity 

of the approach or the results used in the UMTRCA rulemaking. 

2. CURRENT PROGRAMS ARB STRATEGIES 

2.1 INTRODUCTION 

Awareness of radiation and radioactivity dates back only to the 

end of the last century--to the discovery of x-rays in 1895 and the 

discovery of radioactivity in 1896. These discoveries mark the 

beginning of radiation science and the deliberate use of radiation 

and radionuclides in science, medicine, and industry. 

The findings of radiation science rapidly led to the 

development of medical and industrial radiology, nuclear physics, 

and nuclear medicine. By the 192OPs, the use of x-rays in diagnostic 

medicine and industrial applications was widespread, and radium was 

being used by industry for luminescent dials and by doctors in 

therapeutic procedures. By the 19301s, biomedical and genetic 

researchers were studying the effects of radiation on living 

organisms, and physicists were beginning to understand the 

mechanisms of spontaneous fission and radioactive decay. By the 

19401s, a self-sustaining fission reaction was demonstrated, which 

led directly to the construction of the first nuclear reactors and 

atomic weapons. 

Developments since the end of World War I1 have been rapid. 

Today the use of x-rays and radioactive materials is widespread and 

includes : 

0 Nuclear reactors (and their supporting fuel-cycle 

facilities) generate electricity, power ships and 

submarines, produce radioisotopes for research, space, 

defense, and medical applications. They are also used as 

research tools for nuclear engineers and physicists. 

o Particle accelerators produce radioisotopes and are used 

as research tools for studying the structure of materials 

and atoms. 

o The radiopharmaceutical industry provides the 

radioisotopes needed for biomedical research and nuclear 

medicine. 

o Nuclear medicine has developed as a recognized medical 

specialty in which radioisotopes are used in the 

diagnosis and treatment of numerous diseases. 

o X-rays are widely used as a diagnostic tool in medicine 

and in such diverse industrial fields as oil exploration 

and nondestructive testing. 

o Radionuclides are used in such common consumer products 

as luminous-dial wristwatches and smoke detectors. 
The following sections of this chapter provide a brief history 

of the evolution of radiation protection philosophy and an outline 

of the current regulatory programs and strategies of the government 

agencies responsible for ensuring that radiation and radionuclides 

are used safely. 

2.2 THE INTERNATIONAL COMMISSION ON RADIOLOGICAL PROTECTION AND THE 

NATIONAL COUNCIL ON RADIATION PROTECTION AND MEASUREMENTS 

Initially, the dangers and risks posed by x-rays and 

radioactivity were little understood. By 1896, however, "x-ray 

burns" were being reported in the medical literature, and by 1910, 

it was understood that such "burns" could also be caused by 

radioactive materials. By the 1920rs, sufficient direct evidence 

(from experiences of radium dial painters, medical radiologists, and - miners) and indirect evidence (from biomedical and genetic 

experiments with animals) had been accumulated to persuade the 

scientific community that an official body should be established to 

make recommendations concerning human protection against exposure to 

x-rays and radium. 

At the Second International Congress of Radiology meeting in 

Stockholm, Sweden, in 1928, the first radiation protection 

- commission was created. Reflecting the use of radiation and 

radioactive materials at the time, the body was named the 

International X-ray and Radium Protection Commission and was charged 

with developing recommendations concerning protection from 

radiation. In 1950, to reflect better its role in a changing world, 

the Commission was reconstituted and renamed the International 

Commission on Radiation Protection (ICRP) . 

During the Second International Congress of Radiology, the 

newly created Commission suggested to the nations represented at the 

Congress that they appoint national advisory committees to represent 

their viewpoints before the ICRP, and to act in concert with the 

Commission in developing and disseminating recommendations on 

radiation protection. This suggestion led to the formation, in 1929, 

of the Advisory Group. After a series of reorganizations and name 

changes, this committee emerged in 1964 in its present form as the 

congressionally chartered National Council on Radiation Protection 

and Measurements (NCRP). The congressional charter provides for the 

NCRP to: 

o Collect, analyze, develop, and disseminate in the public 

interest information and recommendations about radiation 

protection and radiation quantities, units, and 

measurements. 

o Develop basic concepts about radiation protection and 

radiation quantities, units, and measurements, and the 

application of these concepts. 
o Provide a means by which organizations concerned with 

radiation protection and radiation quantities, units, and 

measurements may cooperate to use their combined 

resources effectively and to stimulate the work of such 

organizations. 

o Cooperate with the ICRP and other national and 

international organizations concerned with radiation 

protection and radiation quantities, units, and 

measurements. 

Throughout their existence, the ICRP and the NCRP have worked 

together closely to develop radiation protection recommendations 

that reflect the current understanding of the dangers associated 

with exposure to ionizing radiation. The ICRP and the NCRP function 

as non-government advisory bodies. Their recommendations are not 

binding on any government or user of radiation or radioactive 

materials. 

The first exposure limits adopted by the ICRP and the NCRP 

(ICRP34, ICRP38, and NCRP36) established 0.2 roentgen/dayl as the 

"tolerance dose" for occupational exposure to x-rays and gamma 

radiation from radium. This limit, equivalent to an absorbed dose of 

approximately 25 rads/year as measured in air, was established to 

guard against the known effects of ionizing radiation on superficial 

tissue, changes in the blood, and "derangement" of internal organs, 

especially the reproductive organs. At the time the recommendations 

were made, high doses of radiation were known to cause observable 

effects, but the epidemiological evidence at the time was inadequate 

even to imply the carcinogenic induction effects of moderate or low 

doses. Therefore, the aim of radiation protection was to guard 

against known effects, and the "tolerance dose" limits that were 

adopted were believed to represent the level of radiation that a 

person in normal health could tolerate without suffering observable 

effects. The concept of a tolerance dose and the recommended 

occupational exposure limit of 0.2 R/day for x and gamma radiation 

remained in effect until the end of the 1940's. The recommendations 

of the ICRP and the NCRP made no mention of exposure of the general 

populace. 

BY the end of World War 11, the widespread use of radioactive 

materials and scientific evidence of genetic and somatic effects at 

lower doses and dose rates suggested that the radiation protection 

recommendations of the NCRP and the ICRP would have to be revised 

downward. 

The NCRPfs recommendation was 0.1 roentgen/day measured in 

air. This limit is roughly equivalent to the ICRP limit, which 

was conventionally measured at the point of exposure and included 

backscatter. 
By 1948, the NCRP had formulated its position on appropriate 

new limits. These limits were largely accepted by the ICRP in its 

recommendations of 1950 and formally issued by the NCRP in 1954 

(ICRP51, NCRP54). Whereas the immediate effect was to lower the 

basic whole body occupational dose limit to 0.3 rad/week 

(approximately 15 rads/year), the revised recommendations also 

embodied several new and important concepts in the formulation of 

radiation protection criteria. 

First, the recommendations recognized the difference in the 

effects of various types and energies of radiation; both ICRP and 

NCRP recommendations include discussions of the weighting factors 

that should be applied to radiations of differing types and 

energies. The NCRP advocated the use of the "rem" to express the 

equivalence in biological effect between radiations of differing 

types and energy.2 Although the ICRP noted the shift toward the 

acceptance of the rem, it continued to express its recommendations 

in terms of the rad, with the caveat that the limit for the absorbed 

dose due to neutron radiation should be one-tenth the limit for x, 

gamma, or beta radiation. 

Second, the recommendations of both organizations introduced 

the concept of critical organs and tissues. This concept was 

intended to ensure that no tissue or organ, with the exception of . 

the skin, would receive a dose in excess of that allowed for the 

whole body. At the time, scientific evidence was lacking on 

tissues and organs. Thus, all blood-forming organs were considered 

critical and were limited to the same exposure as the whole body. 

Third, the NCRP recommendations included the suggestion that 

individuals under the age of 18 receive no more than one-tenth the 

Defining the exact relationship between exposure, absorbed 

dose, and dose equivalent is beyond the scope of this document. 

In simple terms, the exposure is a measure of the charge induced 

by x and gamma radiation in air. Absorbed dose is a measure of 

the energy per unit mass imparted to matter by radiation. Dose 

equivalent is an indicator of the effect on an organ or tissue by 

weighting the absorbed dose with a quality factor, Q, dependent 

on the radiation type and energy. The customary units for 

exposure, absorbed dose, and dose equivalent are the roentgen, 

rad and rem, respectively. Over the range of energies typically 

encountered, the exposure, dose and dose equivalent from x and 

gamma radiation have essentially the same values in these units. 

For beta radiation, the absorbed dose and dose equivalent are 

generally equal also. At the time of these recommendations, a 

quality factor of 10 was recommended for alpha radiation. Since 

1977, a quality factor of 20 has primarily been used, i.e., for 

alpha radiation, the dose equivalent is 20 times the absorbed 

dose. 
exposure allowed for adults. The reasoning behind this particular 

recommendation is interesting, as it reflects clearly the lim'ted 

knowledge of the times. The scientific evidence indicated a clear 

relationship between accumulated dose and genetic effect. However, 

this evidence was obtained exclusively from animal studies that had 

been conducted with doses ranging from 25 to thousands of rads. 

There was no evidence from exposure less than 25 rads accumulated 

dose, and the interpretation of the animal data and the implications 

for humans were unclear and did not support a specific permissible 

dose. The data did suggest that genetic damage was more dependent on 

accumulated dose than previously believed, but experience showed 

that exposure for prolonged periods to the permissible exposure 

limit (1.0 R/week) did not result in any observable genetic effects. 

The NCRP decided that it was not necessary to change the 

occupational limit to provide additional protection beyond that 

provided by the reduction in the permissible exposure limit of 0.3 

R/week. At the same time, it recommended limiting the exposure of 

individuals under the age of 18 to assure that they did not 

accumulate a genetic dose that would later preclude their employment 

as radiation workers. The factor of ten was rather arbitrary but was 

believed to be sufficient to protect the future employability of all 

individuals (NCRP54) . 

Fourth, the concept of a tolerance dose was replaced by the 

concept of a maximum permissible dose. The change in terminology 

reflected the increasing awareness that any radiation exposure might 

involve some risk and that repair mechanisms might be less effective 

than previously believed. Therefore, the concept of a maximum 

permissible dose (expressed as dose per unit of time) was adopted 

because it better reflected the uncertainty in our knowledge than 

did the concept of tolerance dose. The maximum permissible dose was 

defined as the level of exposure that entailed a small risk compared 

with those posed by other hazards in life (ICRP51). 

Finally, in explicit recognition of the inadequacy of our 

knowledge regarding the effects of radiation and of the possibility 

that any exposure might have some potential for harm, the 

recommendations included an admonition that every effort should be 

made to reduce exposure to all kinds of ionizing radiation to the 

lowest possible level. This concept, known originally as ALAP (as 

low as practicable) and later as ALARA (as low as reasonably 

achievable), would become a cornerstone of radiation protection 

philosophy. 

During the 195Ors, a great deal of scientific evidence on the 

effects of radiation became available from studies of radium dial 

painters, radiologists, and survivors of the atomic bombs dropped on 

Japan. This evidence suggested that genetic effects and long-term 

somatic effects were more important than previously considered. 

Thus, by the late 195O1s, the ICRP and NCRP recommendations were 

again revised (ICRP59, NCRP59). These revisions include the 

following major changes: the maximum permissible occupational dose 
for who1.e body exposure and the most critical orgaris (blood forming 

organs, gonads, and the larger lens of the eye) was lowered to 5 

rems/year, with a quarterly limit of 3 rems; the limit for exposure 

of other organs was set at 30 rems/year; internal exposures were 

controlled by a comprehensive set of maximum permissible 

concentrations of radionuclides in air and water based on the most 

restrictive case of a young worker; and recommendations were 

included for some nonoccupational groups and for the general 

population (for the first time). 

The lowering of the maximum permissible whole-body dose from 

0.3 rad/week to 5 rems/year, with a quarterly limit of 3 rems, 

reflects both the new evidence and the uncertainties of the time. 

Although no adverse effects had been observed among workers who had 

received the maximum permissible dose of 0.3 rad/week, there was 

concern that the lifetime accumulation of as much as 750 rads (15 

rads/year times 50 years) was too much. Lowering the maximum 

permissible dose by a factor of three was believed to provide a 

greater margin of safety. At the same time, operational experience 

showed that a limit of 5 rems/year could be met in most instances, 

particularly with the additional operational flexibility provided by 

expressing the limit on an annual and quarterly basis. 

The recommendations given for nonoccupational exposures were 

based on concerns about genetic effects. The evidence available . suggested that genetic effects were primarily dependent on the total 

accumulated dose. Thus, having sought the opinions of respected 

geneticists, the ICRP and the NCRP adopted the recommendation that 

accumulated gonadal dose to age 30 be limited to 5 rems from sources 

other than natural background and medical exposure. As an 

operational guide, the NCRP recommended that the maximum dose to any 

individual be limited to 0.5 rem/year, with maximum permissible body 

burdens of radionuclides (to control internal exposures) set at onetenth that allowed for radiation workers. These values were derived 

from consideration of the genetically significant dose to the 

population and were established "primarily for the purpose of 

keeping the average dose to the whole population as low as 

reasonably possible, and not because of the likelihood of specific 

injury to the individual" (NCRP59). 

During the 1960fs, the ICRP and NCRP again lowered the maximum 

permissible dose limits (ICRP65, NCRP71). The considerable 

scientific data on the effects of exposure to ionizing radiation 

were still inconclusive with respect to the dose response 

relationship at low exposure levels; thus, both organizations 

continued to stress the need to keep all exposures to the lowest 

possible level. 

The NCRP and the ICRP made the following similar 

recommendations: 
o Limit the dose to the whole-body, red bone marrow, and 

gonads to 5 rems in any year, with a retrospective limit 

of 10 to 15 rems in any given year as long as total 

accumulated dose did not exceed 5X(N-l8), where N is the 

age in years. 

o Limit the dose to the skin, hands, and forearms to 15, 

75, and 30 rems per year, respectively. 

o Limit the dose to any other organ or tissue to 15 rems 

per year. 

o Limit the average dose to the population to 0.17 rem per 

year. 

The scientific evidence and the protection philosophy on which 

the above recommendations were based were set forth in detail in 

NCRP71. In the case of occupational exposure limits, the goal of 

protection was to ensure that the risks of genetic and somatic 

effects were small enough to be comparable to the risks experienced 

by workers in other safe industries. The numerical limits 

recommended were based on the linear, no-threshold, dose-response 

model and were believed to represent a level of risk that was 

readily acceptable to an average individual. For nonoccupational 

exposures, the goal of protection was to ensure that the risks of 

genetic or somatic effects were small compared with other risks 

encountered in everyday life. The derivation of specific limits was 

complicated by the unknown dose-response relationship at low 

exposure levels and the fact that the risks of radiation exposure 

did not necessarily accrue to the same individuals who benefited 

from the activity responsible for the exposure. Therefore, it was 

necessary to derive limits that adequately protected each member of 

the public and to the gene pool of the population as a whole, while 

still allowing the development of beneficial uses of radiation and 

radionuclides . 

In 1977, the ICRP made a fundamental change in its 

recommendations when it abandoned the critical organ concept in 

favor of the weighted whole-body effective dose equivalent concept 

for limiting occupational exposure (ICRP77). The change, made to 

reflect an increased understanding of the differing radiosensitivity 

of the various organs and tissues, did not affect the overall limit 

of 5 rems per year for workers, but included a recommendation that 

chronic exposures of the general public from all controllable 

sources be limited to 0.1 redyear. 

Also significant, ICRP's 1977 recommendations represent the 

first explicit attempt to relate and justify permissible radiation 

exposures with quantitative levels of acceptable risk. Thus, average 

occupational exposures (approximately 0.5 redyear) are equated with 

risks in safe industries, given as 1.0 E-4 annually. At the maximum 

limit of 5 rems/year, the risk is equated with that experienced by 
some workers in recognized hazardous occupations. Similarly, the 

risks implied by the nonoccupational limit of 0.5 rem/year are 

equated to levels of risk of less than 1.0 E-2 in a lifetime; the 

general populace's average exposure is equivalent to a lifetime risk 

on the order of 1.0 E-4 to 1.0 E-3. The ICRP believed these levels 

of risk were in the range that most individuals find acceptable. 

In June 1987, the NCRP revised its recommendations to be 

comparable with those of the ICRP (NCRP87). The NCRP adopted the 

effective dose equivalent concept and its related recommendations 

regarding occupational and nonoccupational exposures to acceptable 

levels of risk. However, the NCRP did not fully adopt a risk-based 

system because of the uncertainty in the risk estimates and because 

the details of such a system have yet to be elaborated. 

The NCRP recommendations in (NCRP87) for occupational exposures - correspond to the ICRP recommendations. In addition, the relevant 

nonoccupational exposure guidelines are: 

o 0.5 rem/year effective whole-body dose equivalent, not 

including background or medical radiation, for 

individuals in the population when the exposure is not 

continuous. 

- o 0.1 rem/year effective whole-body dose equivalent, not 

including background or medical radiation, for 

individuals in the population when the exposure is 

continuous. 

o Continuous use of a total dose limitation system based on 

justification of every exposure and application of the 

"as low as reasonably achievable" philosophy. 

The NCRP equates continuous exposure at a level of 0.1 rem/year 

to a lifetime risk of developing cancer of about one in a thousand. 

The NCRP has not formulated exposure limits for specific organs, but 

it notes that the permissible limits will necessarily be higher than 

the whole-body limit in inverse ratio for a particular organ to the 

total risk for whole-body exposure. 

In response to EPh's proposed national emission standards for 

radionuclides, the NCRE suggested that since the 0.1 rem/year limit 

is the limit for all exposures from all sources (excluding natural 

background and medical radiation), the operator of any site 

responsible for more than 25 percent of the annual limit be required 

to assure that the exposure of the maximally exposed individual is 

less than 0.1 rem/year from all sources (NCRP84, NCRP87). 

2.3 FEDERAL GUIDANCE 

The wealth of new scientific information on the effects of 

radiation that became available in the 1950's prompted the President 
to establish an official government entity with responsibility for 

formulating radiation protection criteria and coordinating radiation 

protection activities. Executive Order 10831 established the 

Federal Radiation Council (FRC) in 1959. The Council included 

representatives from all of the Federal agencies concerned with 

radiation protection and acted as a coordinating body for all of the 

radiation activities conducted by the Federal government. In 

addition to its coordinating function, the Council's major 

responsibility was to "...advise the President with respect to 

radiation matters, directly or indirectly affecting health, 

including guidance for all Federal Agencies in the formulation of 

radiation standards and in the establishment and execution of 

programs of cooperation with States ..." (FRC60). 

The Council's first recommendations concerning radiation 

protection standards for Federal agencies were approved by the 

President in 1960. Based largely on the work and recommendations of 

the ICRP and the NCRP, the guidance established the following limits 

for occupational exposures: 

o Whole-body head and trunk, active blood-forming organs, 

gonads, or lens of eye--not to exceed 3 rems in 13 weeks 

and total accumulated dose limited to 5 times the number 

of years beyond age 18. 

o Skin of whole body and thyroid--not to exceed 10 rems in 

13 weeks or 30 rems per year. 

o Hands, forearms, feet, and ankles--not to exceed 25 rems 

in 13 weeks or 75 rems per year. 

o Bone--not to exceed 0.1 microgram of Ra-226 or its 

biological equivalent. 

o Any other organ--not to exceed 5 rems per 13 weeks or 15 

rems per year. 

Although these levels differ slightly from those recommended by 

NCRP and ICRP at the time, the differences did not represent any 

greater or lesser protection. In fact, the FRC not only accepted the 

levels recommended by the NCRP for occupational exposure, it adopted 

the NCRP's philosophy of acceptable risk for determining 

occupational exposure limits. Although quantitative measures of risk 

were not given in the guidance, the prescribed levels were not 

expected to cause appreciable bodily injury to an individual during 

his or her lifetime. Thus, while the possibility of some injury was 

not zero, it was expected to be so low as to be acceptable if there 

was any significant benefit derived from the exposure. 

The guidance also established dose equivalent limits for 

members of the public. These were set at 0.5 rem per year (whole 

body) for an individual and an average of 5 rems in 30 years 
(gonadal) per capita. The guidance also provided for developing a 

suitable sample of the population as a basis for determining 

compliance with the limit when doses to all individuals are unknown. 

Exposure of this population sample was not to exceed 0.17 rem per 

capita per year. The population limit of 0.5 rem to any individual 

per year was derived from consideration of natural background 

exposure. Natural background radiation varies by a factor of two to 

four from location to location. 

In addition to the formal exposure limits, the guidance also 

established as Federal policy that there should be no radiation 

exposure without an expectation of benefit and that "every effort 

should be made to encourage the maintenance of radiation doses as 

far below this guide as practicable." The requirements to consider 

benefits and keep all exposure to a minimum were based on the 

possibility that there is no threshold dose for radiation. The 

linear non-threshold dose response was assumed to place an upper 

limit on the estimate of radiation risk. However, the FRC explicitly 

recognized that it might also represent the true level of risk. If 

so, then any radiation exposure carried some risk, and it was 

necessary to avoid all unproductive exposures and to keep all 

productive exposures as "far below this guide as practicable." 

In 1967, the Federal Radiation Council issued guidance for the 

control of radiation hazards in uranium mining (FRC67). The need for 

such guidance was clearly indicated by the epidemiological evidence 

that showed a higher incidence of lung cancer in adult males who 

worked in uranium mines compared with the incidence in adult males 

from the same locations who had not worked in the mines. The 

guidance established specific exposure limits and recommended that 

all exposures be kept as far below the guide limits as possible. The 

limits chosen represented a tradeoff between the risks incurred at 

various exposure levels, the technical feasibility of reducing the 

exposure, and the benefits of the activity responsible for the 

exposure. 

2.4 THE ENVIRONMENTAL PROTECTION AGENCY 

In 1970, the functions of the Federal Radiation Council were 

transferred to the Administrator of the U.S. Environmental 

Protection Agency. In 1971, the EPA revised the Federal guidance for 

the control of radiation hazards in uranium mining (EPA71). Based on 

the risk levels associated with the exposure limits established in 

1967, the upper limit of exposure was reduced by a factor of three. 

The EPA also provided guidance to Federal agencies in the diagnostic 

use of x-rays (EPA78). This guidance establishes maximum skin 

entrance doses for various types of routine x-ray examinations. It 

also establishes the requirement that all x-ray exposures be based 

on clinical indication and diagnostic need, and that all exposure of 

patients should be kept as low as reasonably achievable consistent 

with the diagnostic need. 
In 1981, the EPA proposed new Federal guidance for occupational 

exposures to supersede the 1960 guidance (EPA81). The 1981 

recommended guidance follows, and expands upon, the principles set 

forth by the ICRP in 1977. This guidance was adopted as Federal 

policy in 1987 (EPA87). 

The Environmental Protection Agency has various statutory 

authorities and responsibilities regarding regulation of exposure to 

radiation in addition to the statutory responsibility to provide 

Federal guidance on radiation protection. EPA's standards and 

regulations for controlling radiation exposures are summarized here. 

Reorganization Plan No. 3 transferred to the EPA the authority 

under the U.S. Atomic Energy Act of 1954, as amended, to establish 

generally applicable environmental standards for exposure to 

radionuclides. Pursuant to this authority, in 1977 the EPA issued 

standards limiting exposure from operations of the light-water 

reactor nuclear fuel cycle (EPA77). These standards cover normal 

operations of the uranium fuel cycle, excluding mining and spent 

fuel disposal. The standards limit the annual dose equivalent to any 

member of the public from all phases of the uranium fuel cycle 

(excluding radon and its daughters) to 25 mrems to the whole body, 

75 mrems to the thyroid, and 25 mrems to any other organ. To protect 

against the buildup of long-lived radionuclides in the environment, 

the standard also sets normalized emission limits for Kr-85, 1-129, 

and Pu-239 combined with other transuranics with a half-life 

exceeding one year. The dose limits imposed by the standard cover 

all exposures resulting from releases to air and water from 

operations of fuel cycle facilities. The development of this 

standard took into account both the maximum risk to an individual 

and the overall effect of releases from fuel cycle operations on the 

population and balanced these risks against the costs of effluent 

control. 

Under the authority of the Uranium Mill Tailings Radiation 

Control Act, the EPA has promulgated standards limiting public 

exposure to radiation from uranium tailings piles (EPA83a, (EPA83b). 

Whereas the standards for inactive and active tailings piles differ, 

a consistent basis is used for these standards. Again, the Agency 

sought to balance the radiation risks imposed on individuals and the 

population in the vicinity of the pile against the feasibility and 

costs of control. 

Under the authority of the U.S. Atomic Energy Act of 1954, as 

amended, the EPA has promulgated 40 CFR 191, which establishes 

standards for disposal of spent fuel, high-level wastes, and 

transuranic elements (EPA82). The standard establishes two different 

limits: (1) during the active waste disposal phase, operations must 

be conducted so that no member of the public receives a dose greater 

than that allowed for other phases of the uranium fuel cycle; and 

(2) once the repository is closed, exposure is to be controlled by 

limiting releases. The release limits were derived by summing, over 
long time periods, the estimated risks to all persons exposed to 

radioactive materials released into the environment. The 

uncertainties involved in estimating the performance of a 

theoretical repository led to this unusual approach, and the 

proposed standard admonishes the agencies responsible for 

constructing and operating such repositories to take steps to reduce 

releases below the upper bounds given in the standard to the extent 

reasonably achievable. 

Under the authority of the Atomic Energy Act of 1954, as 

amended, and the Toxic Substance Control Act, the EPA is developing 

proposed environmental standards for the land disposal of low-level 

radioactive wastes and certain naturally occurring and acceleratorproduced radioactive wastes. The proposed standards will establish 

(1) exposure limits for pre-disposal management and storage options, 

(2) criteria for other agencies to follow in specifying wastes that 

are Below Regulatory Concern (BRC), (3) post-disposal exposure 

limits, and (4) groundwater protection requirements. The proposed 

regulations are scheduled to be published in the Federal Register in 

late 1988 (Gr88). 

Under the authority of the Safe Drinking Water Act, the EPA has 

issued interim regulations covering the permissible levels of 

radium, gross alpha and man-made beta, and photon-emitting 

contaminants in community water systems (EPA76). The limits are 

expressed in picocuries/liter. The limits chosen for man-made beta 

and photon emitters equate to approximately 4 mrems/year whole-body 

or organ dose to the most exposed individual. 

Section 122 of the Clean Air Act amendments of 1977 (Public Law 

95-95) directed the Administrator of the EPA to review all relevant 

information and determine if emissions of hazardous pollutants into 

air will cause or contribute to air pollution that may reasonably be 

expected to endanger public health. In December 1979, EPA designated 

radionuclides as hazardous air pollutants under Section 112 of the 

Act. On April 6, 1983, EPA published proposed National Emission 

Standards for radionuclides for selected sources in the Federal 

Register (48 CFR 15076). Three National Emission Standards for 

Hazardous Air Pollutants (NESHAPS), promulgated on February 6, 1985, 

regulated emissions from Department of Energy (DOE) and non-DOE 

Federal facilities, Nuclear Regulatory Commission (NRC) licensed 

facilities, and elemental phosphorus plants (FR85a). Two additional 

NESHAPS, covering radon emission from underground uranium mines and 

licensed uranium mill tailings, were promulgated on April 17, 1985 

and September 24, 1986, respectively (FR85b, FR86). 

2.5 NUCLEAR REGULATORY COMMISSION 

Under the authority of the Atomic Energy Act of 1954, as 

amended, the NRC is responsible for licensing and regulating the use 

of byproduct, source, and special nuclear material, and for ensuring 

that all licensed activities are conducted in a manner that protects 
public health and safety. The Federal guidance on radiation 

protection applies to the NRC; therefore, the NRC must assure that 

none of the operations of its licensees exposes a member of the 

public to more than 0.5 rem/year. The dose limits imposed by the 

EPA's standard for uranium fuel cycle facilities also apply to the 

fuel cycle facilities licensed by the NRC. These facilities are 

prohibited from releasing radioactive effluents in amounts that 

would result in doses greater than the 25 mrems/year limit imposed 

by that standard. 

The NRC exercises its statutory authority by imposing a 

combination of design criteria, operating parameters, and license 

conditions at the time of construction and licensing. It assures 

that the license conditions are fulfilled through inspection and 

enforcement. The NRC licenses more than 7,000 users of 

radioactivity. The regulation of fuel cycle licensees is discussed 

separately from the regulation of byproduct material licensees. 

2.5.1 Fuel Cycle Licenses 

The NRC does not use the term "fuel cycle facilities" to define 

its classes of licensees. The term is used here to coincide with 

EPA's use of the term in its standard for uranium fuel cycle 

facilities. As a practical matter, this term includes the NRC's 

large source and special nuclear material and production and 

utilization facilities. The NRC's regulations require an analysis of 

probable radioactive effluents and their effects on the population 

near fuel cycle facilities. The NRC also ensures that all exposures 

are as low as reasonably achievable by imposing design criteria and 

specific equipment requirements on the licensees. After a license 

has been issued, fuel cycle licensees must monitor their emissions 

and take environmental measurements to ensure that they meet the 

design criteria and license conditions. For practical purposes, the 

NRC adopted the maximum permissible concentrations developed by the 

NCRP to relate effluent concentrations to exposure. 

In the 19701s, the NRC formalized the implementation of as low 

as reasonably achievable exposure levels by issuing a regulatory 

guide for as low as reasonably achievable design criteria. This 

coincided with a decision to adopt, as a design criterion, a maximum 

permissible dose of 5-mrems/year from a single nuclear electric 

generating station. The 5 mrem limit applies to the most exposed 

individual actually living in the vicinity of the reactor and refers 

to whole-body doses from external radiation by air pathway (NRC77). 

2.5.2 Byproduct Material Licenses 

The NRC's licensing and inspection procedure for byproduct 

material users is less uniform than that imposed on major fuel cycle 

licensees for two reasons: (1) the much larger number of byproduct 

material licensees, and (2) their much smaller potential for 

releasing significant quantities of radioactive materials into the 
environment. The prelicensing assurance procedures of imposing 

design reviews, operating practices, and license condLtbons prior to 

construction and operation are similar. 

The protection afforded the public from releases of radioactive 

materials from these facilities can vary considerably because of 

three factors. First, the requirements that the NRC imposes for 

monitoring effluents and environmental radioactivity are much less 

stringent for these licensees. If the quantity of materials handled 

is small enough, the NRC might not impose any monitoring 

requirements. Second, and more important, the level of protection 

can vary considerably because the exact point where the licensee 

must meet the effluent concentrations for an area of unrestricted 

access is not consistently defined. Depending on the particular 

licensee, this area has been defined as the nearest inhabited 

structure, as the boundary of the user's property line, as the roof 

of the building where the effluents are vented, or as the mouth of 

the stack of vent. Finally, not all users are allowed to reach 100 

percent of the maximum permissible concentration in their effluents. 

In fact, the NRC has placed as low as reasonably achievable 

requirements on many of their licensees by limiting them to 10 

percent of the maximum permissible concentration in their effluents. 

. 

2.6 DEPARTMENT OF ENERGY 

- The DOE operates a complex of national laboratories and weapons 

facilities. These facilities are not licensed by the NRC. The DOE 

is responsible, under the U.S. Atomic Energy Act of 1954, as 

amended, for ensuring that these facilities are operated in a manner 

that does not jeopardize public health and safety. 

The DOE is subject to the Federal guidance on radiation 

protection issued by EPA and its predecessor, the FRC. For 

practical purposes, the DOE has adopted the NCRP's maximum 

permissible concentrations in air and water as a workable way to 

ensure that the dose limits of 0.5 rem/year whole-body and 1.5 

rems/year to any organ are being observed. The DOE also has a 

requirement that all doses be kept as low as is reasonably 

achievable, but the contractors who operate the various DOE sites 

have a great deal of latitude in implementing policies and 

procedures to ensure that all doses are kept to the lowest possible 

level. 

The DOE ensures that its operations are within its operating 

guidelines by requiring its contractors to maintain radiation 

monitoring systems around each of its sites and to report the 

results in an annual summary report. New facilities and 

modifications to existing facilities are subject to extensive design 

criteria reviews (similar to those used by the NRC). During the mid19701s, the DOE initiated a systematic effluent reduction program 

that resulted in the upgrading of many facilities and effected a 
corresponding reductyion in the effluents (including airborne and 

liquid radioactive materials) released to the environment. 

As a continuation of this program, DOE has issued proposed 

Order 5400.3 "Draft Radiation Protection of the Public and the 

Environment'hnd has issued several internal guidance documents 

including procedures for the calculation of internal and.externa1 

doses to the public and guidance on environmental surveillance. 

2.7 OTHER FEDERAL AGENCIES 

2.7.1 Department Of Defense 

The Department of Defense operates several nuclear 

installations, including a fleet of nuclear-powered submarines and 

their shore support facilities. The DOD, like other Federal 

agencies, must comply with Federal radiation protection guidance. 

The DOD has not formally adopted any more stringent exposure limits 

for members of the public than the 0.5 rem/year allowed by the 

Federal guidance. 

2.7.2 Center for Medical Devices and Radiological Health 

Under the Radiation Control Act of 1968, the major 

responsibility of the Center for Medical Devices and Radiological 

Health in the area of radiation protection is the specification of 

performance criteria for electronic products, including x-ray 

equipment and other medical devices. This group also performs 

environmental sampling in support of other agencies, but no 

regulatory authority is involved. 

2.7.3 Mine Safety and Health Administration 

The Mine Safety and Health Administration (MSHA) has the 

regulatory authority to set standards for exposures of miners to 

radon and its decay products and other (nonradiological) pollutants 

in mines. The MSHA has adopted the Federal guidance for exposure of 

uranium miners (EPA71). It has no authority or responsibility for 

protecting members of the general public from the hazards associated 

with radiation. 

2.7.4 Occupational Safety and Health Administration 

The Occupational Safety and Health Administration (OSHA) is 

responsible for assuring a safe workplace for all workers. This 

authority, however, does not apply to radiation workers at 

government-owned or NRC-licensed facilities. This group does have 

the authority to set exposure limits for workers at unlicensed 

facilities, such as particle accelerators, but it does not have any 

authority to regulate public exposure to radiation. OSHA has adopted 

the occupational exposure limits of the NRC, except it has not 
imposed the requirement to keep aLL doses as ,low as is reasonably 

achievab1.e. 

2.7.5 Department of Transportation 

The Department of Transportation (DOT) has statutory 

responsibility for regulating the shipment and transportation of 

radioactive materials. This authority includes the responsibility to 

protect the public from exposure to radioactive materials while they 

are in transit. For practical purposes, the DOT has implemented its 

authority through the specification of performance standards for 

shipment containers and by setting maximum exposure rates at the 

surface of any package containing radioactive materials. These 

limits were set to assure compliance with the Federal guidance for 

occupational exposure, and they are believed to be sufficient to 

protect the public from exposure. The DOT also controls potential 

public exposure by managing the routing of radioactive shipments to 

avoid densely populated areas. 

2.8 STATE AGENCIES 

States have important authority for protecting the public from 

the hazards associated with ionizing radiation. In 26 states, the 

states have assumed NRC's inspection, enforcement, and licensing 

responsibilities for users of source and byproduct materials and 

users of small quantities of special nuclear material. These "NRC 

Agreement States," which license and regulate more than 11,500 users 

of radiation and radioactive materials, are bound by formal 

agreements to adopt requirements consistent with those imposed by 

the NRC. The NRC continues to perform this function for all 

licensable uses of the source, byproduct, and special nuclear 

material in the 24 states that are not Agreement States. 

Nonagreement states, as well as NRC Agreement States, regulate 

the exposures to workers from electronic sources of radiation. Also, 

all states retain the authority to regulate the use of naturally 

occurring (i.e., radium) and accelerator-produced radioactive 

materials. 
REFERENCES 

U.S. Environmental Protection Agency, "Radiation 

Protection Guidance for Federal Agencies: Underground 

Mining of Uranium Ore," Federal Register 36(132), July 9, 

1971. 

U.S. Environmental Protection Agency, National Interim 

Primary Drinking Water Regulations, EPA-570/9-76-003, 

1976. 

U.S. Environmental Protection Agency, "Environmental 

Radiation Protection Standards for Nuclear Power 

Operations,"40 CFR 190, Federal Register 42(9), January 

13, 1977. 

U.S. Environmental Protection Agency, "Radiation 

Protection Guidance to Federal Agencies for Diagnostic XRays," Federal Register 43(22), February 1, 1978. 

U.S. Environmental Protection Agency, "Federal Radiation 

Protection Guidance for Occupational Exposure," Federal 

Register 46(15), January 23, 1981. 

U.S. Environmental Protection Agency, "Environmental 

Standards for the Management and Disposal of Spent 

Nuclear Fuel, High-Level and Transuranic Radioactive 

Wastes," 40 CFR 191, -. Federal Register 47(250), December 

29, 1982. 

U.S. Environmental Protection Agency, "Standards for 

Remedial Actions at Inactive Uranium Processing Sites," 

Federal Register 48(590), January 5, 1983. 

U.S. Environmental Protection Agency, "Environmental 

Standards for Uranium Mill Tailings at Licensed 

Commercial Processing Sites; Final Rule," Federal 

Register 48(196), October 7, 1983. 

U.S. EPA, "Radiation Protection Guidance to Federal 

Agencies for Occupational ~xposure," Federal Register 52 

(2822), January 27, 1987. 

Federal Register 50, 5190-5200, February 6, 1985. 

- 

Federal Register 50, 15386-15394, April 17, 1985. 

Federal Register 51, 34056-34067, September 24, 1986 
FRCGO Federal Radiation Council, "Radiation Protection Guidance 

for Federal Agencies," Federal Register 44(02), May 18, 

1960. 

FRC67 Federal Radiation Council, Guidance for the Control of 

Radiation Hazards in Uranium Mininq, Report No. 8, 

September 1967. 

Gr88 Gruhlke, J.M., Galpin, F.L., and Holcomb, W.F., 

"Overview of EPA1s Environmental Standards for the Land 

Disposal of LLW and NARM Waste-1988," ORP/EPA, for 

presentation at DOE'S 10th Annual LLW Management 

Conference, Denver, Colorado, August 30 - September 1, 

1988. 

ICRP34 International X-Ray and Radium Protection Commission, - "International Recommendations for X-Ray and Radium 

Protection," British Journal of Radiology 7, 695-699, 

1934. 

ICRP38 International X-Ray and Radium Protection Commission, 

"International Recommendations for X-Ray and Radium 

Protection," Amer. J. of Roent and Radium 40, 134-138, 

1938. 

ICRP51 International Commission on Radiological Protection, 

"International Recommendations on Radiological Protection 

1950," British Journal of Radiology 24, 46-53, 1951. 

ICRP59 International Commission on Radiological Protection, 

Recommendations of the ICRP 1958, ICRP Publication 1, 

Pergamon Press, Oxford, 1958. 

ICRE'65 International Commission on Radiological Protection, 

Recommendations of the ICRP 1965, ICRP Publication 9, 

Pergarnon Press, Oxford, 1965. 

ICRP77 International Commission on Radiological Protection, 

Recommendations of the ICRP, ICRP Publication 26, 

Pergamon Press, Oxford, 1977. 

NCRP36 Advisory Committee on X-Ray and Radium Protection, X-Ray 

Protection, NCRP Report No. 3, 1936. 

NCRP 5 4 National Committee on Radiation Protection, Permissible 

Dose From External Sources of Ionizing Radiation, 

National Bureau of Standards Handbook 59, 1954. 

NCRP 5 9 National Committee on Radiation Protection, Maximum 

Permissible Body Burdens and Maximum Permissible 
Concentrations of Radionuclides in Air and in Water for 

Occupational Exposure, National Bureau of Standards 

Handbook 69, 1959. 

NCRP 7 1 National Council on Radiation Protection and 

Measurements, Basic Radiation Protection Criteria, NCRP 

Report No. 39, 1971. 

NCRP 8 4 National Council on Radiation Protection and 

Measurements, Control of Air Emissions of Radionuclides, 

September 18, 1984. 

NCRP8 7 National Council on Radiation Protection and 

Measurements, Recommendations on Limits for Exposure to 

Ionizing Radiation, NCRP Report 91, June 1, 1987. 

NRC7 7 U.S. Nuclear Regulatory Commission, 1977, Appendix I: 10 

CFR 50, Federal Register 44, September 28, 1979. 

3. HAZARD IDENTIFICATION 

The adverse biological reactions associated with ionizing 

radiations, and hence with radioactive materials, are 

carcinogenicity, mutagenicity, and teratogenicity. 

Carcinogenicity is the ability to produce cancer. Mutagenicity 

is the property of being able to induce genetic mutation, which 

may be in the nucleus of either somatic (body) or germ 

(reproductive) cells. Teratogenicity refers to the ability of an 

agent to induce or increase the incidence of congenital 

malformations as a result of permanent structural or functional 

deviations produced during the growth and development of an 

embryo (these are more commonly referred to as birth defects). 

Ionizing radiation causes injury by breaking constituent 

body molec~les into electrically charged fragments called "ions" 

and thereby producing chemical rearrangements that may lead to 

permanent cellular damage. The degree of biological damage 

caused by various types of radiation varies according to how 

close together the ionizations occur. Some ionizing radiations 

(e.g., alpha particles) produce intense regions of ionization. 

For this reason, they are called high-LET (linear energy 

transfer) particles. Other types of radiation (such as 

high-energy photons [x-rays]) that release electrons that cause 

ionization and beta particles are called low-LET radiations 

because of the sparse pattern of ionization they produce. In 

equal doses, the carcinogenicity and mutagenicity of high-LET 

radiations are generally an order of magnitude or more greater 

than those of low-LET radiations. 

Radium, radon, radon daughters, and several other naturally 

occurring radioactive materials emit alpha particles; thus, when 

these materials are ingested or inhaled, they are a source of 

high-LET particles within the body. Man-made radionuclides are 

usually beta and photon emitters of low-LET radiations. Notable 

exceptions to this generalization are plutonium and other 

transuranium radionuclides, most of which emit alpha radiation. 

3.1 EVIDENCE THAT RADIATION IS CARCINOGENIC 

The production and properties of x-rays were demonstrated 

within one month of the public reporting of Roentgen's discovery 

of x-rays. The first report of acute skin injury was made in 1896 

(Mo671. The first human cancer attributed to this radiation was 

reported in 1902 (Vo02). By 1911, 94 cases of radiation-related 

skin cancer and 5 cases of leukemia in man had been reported in 

the literature (Up15). Efforts to study this phenomenon through 
the use of experimental animals produced the first reported 

radiation-related cancers in experimental. anima1.s in 1910 and 

1912 (MalO, Ma12). Since that time, an extensive body of 

literature has evolved on radiation carcinogenesis in man and 

animals. This literature has been reviewed most recently by the 

United Nations Scientific Committee on the Effects of Atomic 

Radiation (UNSCEAR) and by the National Academy of Sciences 

Advisory Committee on the Biological Effects of Ionizing 

Radiations (NAS-BEIR Committee) (UNSCEAR82, NAS80) . 

Identification of the carcinogenicity of radioactive 

emissions followed a parallel course. In 1921, Uhlig 

first associated inhaled radioactive material and carcinogenesis 

in man in a study of lung cancer in underground miners in the Erz 

Mountains (Uh21). This association was reaffirmed by Ludewig and 

Lorenser in 1924 (Lu24). Ingestion of radioactive materials was 

also demonstrated to be a pathway for carcinogenesis in man. As 

early as 1925, ingested radium was known to cause bone necrosis 

(Ho25), and in 1929, the first report was published on the 

association of radium ingestion and osteogenic sarcoma (Ma29). 

. The expected levels of exposure to radioactive pollutants in 

the environment are too low to produce an acute (immediate) 

- response. Their effect is more likely to be a delayed response, 

in the form of an increased incidence of cancer long after 

exposure. An increase in cancer incidence or mortality with 

increasing radiation dose has been demonstrated for many types of 

cancer in both human populations and laboratory anima1.s 

(UNSCEAR77, 82). Studies of humans exposed to internal or 

external sources of ionizing radiation have shown that the 

incidence of cancer increases with increased radiation exposure. 

This increased incidence, however, is usually associated with 

appreciably greater doses and exposure frequencies than those 

encountered in the environment. Malignant tumors most often 

appear long after the radiation exposure, usually 10 to 35 years 

later (NAS80, UNSCEAR82). The tumors appear in various organs. 

In the case of internal sources of radiation due to radioactive 

materials, the metabolism of the materials generally leads to 

their deposition in specific organs, which results in a radiation 

dose and higher-than-normal risk of cancer in these organs. 

Whereas many, if not most, chemical carcinogens appear to be 

organ- or tissue-specific, ionizing radiation can be considered 

pancarcinogenic. According to Storer (St75): "Ionizing 

radiation in sufficiently high dosage acts as a complete 

carcinogen in that it serves as both initiator and promoter. 

Further, cancers can be induced in nearly any tissue or organ of 

man or experimental animals by the proper choice of radiation 
* * * DRAFT * * * 

dose and exposure schedule." Radiation-induced cancers in humans 

have been reported in the following tissues: thyroid, female 

breast, lung, bone marrow (leukemia), stomach, liver, large 

intestine, brain, salivary glands, bone, esophagus, small 

intestine, urinary bladder, pancreas, rectum, lymphatic tissues, 

skin, pharynx, uterus, ovary, mucosa of cranial sinuses, and 

kidney (UNSCEAR77, 82; NAS72, 80; Be77, Ka82, Wa83). 

Studies of populations exposed to high levels of radiation 

have identified the organs at greatest risk following radiation 

exposure. Brief discussions of these findings follow. 

1. Atomic Bomb Survivors - The survivors of the atomic bomb 

explosions at Hiroshima and Nagasaki, Japan, were exposed to 

whole-body external radiation doses of 0 to more than 200 

rads.' An international group has been observing the 

population since 1950. The most recent reports published by 

this group (Ka82, Wa83) indicate that an increase in cancer 

mortality has been shown for many cancers, including 

leukemia; thyroid, breast, and lung cancer; esophageal and 

stomach cancer; colon cancer; cancer of urinary organs; and 

multiple myeloma. 

2. Ankylosing Spondylitics - A large group of patients was 

given x-ray therapy for ankylosing spondylitis of the spine 

during the years 1934 to 1954. X-ray doses usually exceeded 

100 rad. British investigators have been following this 

group since about 1957. The most recent review of the data 

shows excess cancers in irradiated organs, including 

leukemia, lymphoma, lung and bone cancer, and cancer of the 

pharynx, esophagus, stomach, pancreas, and large 

intestine (UNSCEAR77, NAS80) . 

3. Mammary Exposure - Several groups of women who were 

exposed to x-rays during diagnostic radiation of the thorax 

or during radio-therapy for conditions involving the breast 

have been studied. Although most of the groups have been 

followed only a relatively short time (about 15 years), a 

significant increase in the incidence of breast cancer has 

been observed (UNSCEAR77). The dose that produced these 

effects averaged about 100 rads. 

4. Medical Treatment of Benign Conditions - Several groups 

of persons who were medically treated with x-rays to 

alleviate some benign conditions have been studied. Excess 

The rad is the unit of absorbed dose in common use; 1 rad 

equals 100 ergs of absorbed energy per gram of material. 
cancer has developed in many of th.e organs irradiated (e. g., 

breast, brain, thyroid, and probably salivary glands, skin, 

bone, and pelvic organs) following doses ranging from less 

than 10 to more than 100 rads (UNSCEAR77). Excess leukemia 

has also occurred in some groups. The Eollowup period for 

most groups has been short, often less than 20 years. 

5. Underground Miners - Studies of excess cancer mortality 

in U.S. underground miners exposed to elevated levels of 

radon started in the 1950's and 1960's. Groups that have 

worked in various types of mines, including uranium and 

fluorospar, are being studied in the United States, Canada, 

Great Britain, Sweden, China, and Czechoslovakia. Most of 

the miners studied have been subjected to high rates of 

exposure; however, a recent review indicates that increased 

incidence of lung cancer has been observed in some miners 

exposed at cumulative levels approximating those that can 

occur wherever high environmental concentrations of radon 

are present (NAS80). The response shown in all the study 

groups is nearly proportional to the dose (NAS80). 

6. Ingested or Injected Radium - Workers who ingested 

Ra-226 while painting watch and clock dials have been 

studied for 35 to 45 years, and patients who received 

injections of Ra-226 or Ra-224 for medical purposes have 

been studied for 20 to 30 years (NAS72, 80). Excess 

incidence of leukemia and osteosarcoma related to Ra-224 

exposure has been observed. Calculated cumulative average 

doses for these study groups ranged from 200 to 1,700 rads. 

A study now underway that deals with exposure levels under 

90 rads should provide additional data (NAS80). 

7. Injected Thorotrast - Medical use of Thorotrast 

(colloidalthorium dioxide) as an x-ray contrast medium 

introduced radioactive thorium and its daughters into a 

number of patients. Research studies have followed patients 

in Denmark, Portugal, Japan, and Germany for about 40 years 

and patients in the United States for about 10 years 

(UNSCEAR77, NAS80). An increased incidence of liver, bone, 

and lung cancer has been reported in addition to increased 

anemia, leukemia, and multiple myeloma (In79) . Calculated 

cumulative doses range from tens to hundreds of rads. 

8. Diagnostic X-ray Exposure During Pregnancy - Effects of 

x-ray exposure on the fetus during pregnancy have been 

studied in Great Britain since 1954, and several 

retrospective studies have been made in the United States 

since that time (NAS80, UNSCEAR77). Increased incidence of 
leukemia and other childhood cancers have been observed in 

populations exposed to absorbed doses of 0.2 to 20 rads in 

utero (NASBO, UNSCEAR77) . 

Not all of the cancers induced by radiation are fatal. The 

fraction of fatal cancers is different for each type of cancer. 

The BEIR-3 committee estimated the fraction of fatal cancers by 

site and sex (NASBO). Estimates of cancers by site ranged from 

about 20 percent fatal in the case of thyroid cancer to 100 

percent fatal in the case of liver cancer. They concluded that, 

on the average, females have 2 times as many total cancers as 

fatal cancers following radiation exposure, and males have 1.5 

times as many (NASBO). Although many of the radiation-induced 

cancers are not fatal, they still are costly and adversely affect 

the person's lifestyle for the remainder of his or her life. 

Just how these costs and years of impaired life should be 

weighed in evaluating the hazards of radiation exposure is not 

certain. This assessment addresses only the risk of fatal 

carcinogenesis. 

In addition to the evidence that radiation is a 

pancarcinogen, and as such can induce cancers in nearly any 

tissue or organ, it also appears that it can induce cancer by any 

route of exposure (dermal, inhalation, ingestion, and injection). 

Inhalation is likely to be the major route of environmental 

exposure to airborne radioactive pollutants, and the principal 

organ at risk is.likely to be the lung. Some radiation exposure 

to airborne pollutants Gy the ingestion route is possible, 

however, as these pollutants are deposited on soil, on plants, or 

in sources of water. Ingestion of inhaled particulates also 

occurs. Some radionuclides may also cause whole-body gamma 

radiation exposure while airborne or after their deposition on 

the ground. 

Estimates of cancer risk are based on the absorbed dose of 

radiation in an organ or tissue. Given the same type of 

radiation, the risk for a particular dosage would be the same, 

regardless of the source of the radiation. Numerical estimates 

of the cancer risk posed by a unit dose of radiation in various 

organs and tissues are presented in Chapter 6. The models used 

to calculate radiation doses from a specific source are described 

in Chapters 4 and 5. 

The overwhelming body of human epidemiological data makes it 

unnecessary to base major conclusions concerning the risk of 

radiation-induced cancers 011 evidence provided by animal tests; 

however, these data are relevant to the interpretation of human 
data (NAS80) and contribute additional evidence to the 

epidemiological database for humans. Radiation-induced cancers 

have been demonstrated in several animal species, including rats, 

mice, hamsters, guinea pigs, cats, dogs, sheep, cattle, pigs, and 

monkeys. Induced through multiple routes of administration and 

at multiple dose levels, these cancers have occurred in several 

organs or tissues. These animal studies have provided 

information on the significance of dose rate compared with the 

age of the animals at exposure, the sex of the animals, and the 

genetic characteristics of the test strain. They have shown that 

radiation-induced cancers become detectable after varying latent 

periods, sometimes several years after exposure. The studies 

further show that the total number of cancers that eventually 

develop varies consistently with the dose each animal receives. 

Experimental studies in animals have also established that the 

carcinogenic effect of high-LET radiation (alpha radiations or 

neutrons) is greater than that of low-LET radiation (x-rays or 

gamma rays). 

A number of researchers have induced transformations in 

mammalian tissue culture, including the embryo cells of mice and 

hamsters (Bo84, Ke84, Ha84, Gu84). Chromosome aberrations in 

cultured human peripheral lymphocytes have been demonstrated at 

Rn-222 alpha doses of about 48 mrads/y with an external gamma 

dose of about 100 mrads/y (Po77). Another major finding of recent 

research (Gu84) is that DNA from radiation-induced mouse tumors 

contains an activated oncogene that can transform specific types 

of cultured cells when introduced into these cells. The 

researchers also found that a difference in only one base in the 

oncogene was responsible for the transformation. Thus, radiation 

can induce tumors even when only a small change in the DNA occurs 

as a result of irradiation. 

3.2 EVIDENCE THAT RADIATION IS MUTAGENIC 

Radiation can change the structure, number, or genetic 

content of the chromosomes in a cell nucleus. These genetic 

radiation effects are classified as either gene mutations or 

chromosomal aberrations. Gene mutations refer to alterations of 

the basic units of heredity, the genes. Chromosomal aberrations 

refer to changes in the normal number or structure of 

chromosomes. Both gene mutation and chromosomal aberrations are 

heritable; therefore, they are considered together as genetic 

effects. Mutations and chromosomal aberrations can occur in 

somatic (body) or germ (reproductive) cells. In the case of germ 

cells, the mutagenic effect of radiation is not seen in those 

persons exposed to the radiation, but in their descendents. 
Mutations often result in miscarriages or produce such 

undesirable changes in a population as congenital malformations 

that result in mental or physical defects. Mutations occur in 

many types of cells; no tendency toward any specific locus or 

chromosome has been identified. For this reason, they can affect 

any characteristic of a species. A relatively wide array of 

chromosome aberrations occurs in both humans and animals. 

Early experimental studies showed that x-radiation is 

mutagenic. In 1927, H.J. Muller reported radiation-induced 

genetic changes in animals, and in 1928, L.J. Stadler reported 

such changes in plants (Ki62). Although genetic studies were 

carried out in the 193O1s, mostly in plants and fruit flies 

(Drosophila), the bulk of the studies on mammals started after 

the use of nuclear weapons in World War I1 (UNSCEAR58). 

Very few quantitative data are available on radiogenic 

mutations in humans, particularly from low-dose exposures, for 

the following reasons: these mutations are interspersed over 

many generations, some are so mild they are not noticeable, and 

some mutagenic defects that do occur are similar to nonmutagenic 

effects and are therefore not necessarily recorded as mutations. 

The bulk of data supporting the mutagenic character of ionizing 

radiation comes from extensive studies of experimental animals, 

mostly mice (UNSCEAR77, 82; NAS72, 80). These studies have 

demonstrated all forms of radiation mutagenesis--lethal 

mutations, translocations, inversions, nondisjunction, point 

mutations, etc. Mutation rates calculated from these studies are 

extrapolated to humans (because the basic mechanisms of mutations 

are believed to be the same in all cells) and form the basis for 

estimating the genetic impact of ionizing radiation on humans 

(NAS80, UNSCEAR82). The vast majority of the demonstrated 

mutations in human germ cells contribute to both increased 

mortality and illness (NAS80, UNSCEAR82). Moreover, the 

radiation protection community is generally in agreement that the 

probability of inducing genetic changes increases linearly with 

dose and that no "threshold" dose is required to initiate 

heritable damage to germ cells. 

Considerable evidence has been documented concerning the 

production of mutations in cultured cells exposed to radiation. 

Such mutations have been produced in Chinese hamster ovary cells, 

mouse lymphoma cells, human diploid fibroblasts, and human blood 

lymphocytes. Many of the radiation-induced specific types of 

mutations produced in human and Chinese hamster cultured cells 

are associated with structural changes in the X chromosome. 

Evidence suggests that these mutations may be largely due to 

deletions in the chromosomes. 
Mutagenicity in human somatic cells has been demonstrated on 

the basis of chromosome aberrations detected in cultured 

lymphocytes. Chromosome aberrations in humans have been 

demonstrated in lymphocytes cultured from persons exposed to 

ingested Sr-90 and Ra-226 (Tu63); inhaled/ingested Rn-222, 

natural uranium, or Pu-239 (Br77); or inhaled Rn-222 (Po78); and 

in atomic bomb survivors (Aw78). Although no direct evidence of 

health impact currently exists, these chromosome aberrations 

demonstrate that mutagenesis is occurring in somatic cells of 

humans exposed to ionizing radiation. 

Evidence of mutagenesis in human germ cells (cells of the 

ovary or testis) is less conclusive. Studies have been made of 

several populations exposed to medical radiation, atomic bomb 

- survivors, and a population in an area of high background 

radiation in India (UNSCEAR77). Although these studies suggest 

an increased incidence of chromosomal aberrations in germ cells 

following exposure to ionizing radiation, the data are not 

convincing (UNSCEAR77). Investigators who analyzed the data on 

children born to survivors of the atomic bombings of Hiroshima 

and Nagasaki found no statistically significant genetic effects 

due to parental exposure (Sc81). They did find, however, that 

the observed effects are in the direction of genetic damage from 

the bomb radiation exposure. 

The incidence of serious genetic disease due to mutations 

and chromosome aberrations induced by radiation is referred to as 

genetic detriment. Serious genetic disease includes inherited 

ill health, handicaps, or disabilities. Genetic disease may be 

manifest at birth or may not become evident until some time in 

adulthood. Radiation-induced genetic detriment includes 

impairment of life, shortened life span, and increased 

hospitalization. Estimates of the frequency of radiation-induced 

genetic impairment are presented in Chapter 6 of this document. 

Although the numbers represent rough approximations, they are 

relatively small in comparison with the magnitude of detriment 

associated with spontaneously arising genetic diseases 

(UNSCEAR82) . 

3.3 EVIDENCE THAT RADIATION IS TERATOGENIC 

Teratogenicity is the malformation of tissues or organs of a 

fetus resulting from physiologic and biochemical changes. 

Radiation is a well-known teratogenic agent. Case reports of 

radiation-induced teratology were made as early as 1921 (St21). 

By 1929, an extensive review of a series of pregnancies yielded 

data indicating that 18 of the children born to 76 irradiated 
mothers had abliormally small heads (microcephaly) (Mu30). 

Although the radiation dose in these cases is not known, it was 

high. 

Early experimental studies (primarily in the 1940's and 

1950's) demonstrated the teratogenic properties of x-rays in 

fish, amphibia, chick, mouse, and rat embryos (Ru53). These 

experiments showed that the developing fetus is much more 

sensitive to radiation than the mother and provided data on 

periods of special sensitivity and dose-response. The 

malformations produced in the embryo depend on which cells, 

tissues, or organs in the fetus are most actively differentiating 

at the time of radiation. Embryos are relatively resistant to 

radiation-induced teratogenic effects during the earliest stages 

of their development and are most sensitive during development of 

the neuroblast (these cells eventually become the nerve cells). 

These experiments showed that different malformations could be 

elicited by irradiating the fetus at specific times during its 

development. 

Substantial evidence points to the ability of radiation to 

induce teratogenic effects in human embryos as well. In a study 

of mental retardation in children exposed in utero to atomic bomb 

radiation in Hiroshima and Nagasaki, researchers found that 

damage to the child appears to be related linearly to the 

radiation dose that the fetus receives (Ot84). The greatest risk 

of damage occurs at 8 to 15 weeks, which is the time the nervous 

system is undergoing the most rapid differentiation and 

proliferation of cells. They concluded that the age of the fetus 

at the time of exposure is the most important factor in determining the extent and type of damage from radiation. A numerical 

estimate of mental retardation risk due to radiation is given in 

Chapter 6. 

3.4 UNCERTAINTIES 

Although much is known about radiation dose-effect 

relationships at high-level doses, uncertainty exists when 

dose-effect relationships based on direct observations are 

extrapolated to lower doses, particularly when the dose rates are 

low. As described in Chapter 6, the range of extrapolation 

varies depending on the sensitivity of the organ system. For 

breast cancer, this may be as small as a factor of four. 

Uncertainties in the dose-effect relationships are recognized to 

relate to such factors as differences in quality and type of 

radiation, total dose, dose distribution, dose rate, and 

radiosensitivity (including repair mechanisms, sex, variations in 

age, organ, and state of health). The range of uncertainty in 
the estimates of radiation risk is examined in some detail in 

Chapters 5, 6 and 7. 

The uncertainties in the details of the mechanisms of 

carcinogenesis, mutagenesis, and teratogenesis make it necessary 

to rely on the considered judgments of experts on the biological 

effects of ionizing radiation. These findings, which are well 

documented in publications by the National Academy of Sciences 

and the United Nations Scientific Committee on the Effects of 

Atomic Radiation, are used by advisory bodies such as the 

International Commission on Radiological Protection (ICRP) in 

developing their recommendations. The EPA has considered all 

such findings in formulating its estimate of the relationship 

between radiation dose and response. 

- Estimates of the risk from ionizing radiation are often 

limited to fatal cancers and genetic effects. Quantitative data 

on the incidence of nonfatal radiogenic cancers are sparse, and 

the current practice is to assume that the total cancer incidence 

resulting from whole-body exposure is 1.5 to 2.0 times the 

mortality. In 1980, the NAS-BEIR Committee estimated the effects 

of ionizing radiation directly from epidemiology studies on the 

basis of both cancer incidence and the number of fatal cancers 

induced per unit dose (NAS80). The lifetime risk from chronic 

exposure can be estimated from these data, either on the basis of 

(1) relative risk (i.e., the percentage of increase in fatal 

cancer), or (2) absolute risk (i.e., the number of excess cancers 

per year at risk following exposure). The latter method results 

in numerically smaller estimated risks for common cancers, but a 

larger estimated risk for rare cancers. 

3.5 SUWY OF EVIDENCE THAT RADIATION IS A CARCINOGEN, 

MUTAGEN, AND TERATOGEN 

Radiation has been shown to be a carcinogen, a mutagen, and 

a teratogen. At sufficiently high doses, radiation acts as a 

complete carcinogen, serving as both initiator and promoter. 

With proper choice of radiation dose and exposure schedule, 

cancers can be induced in nearly any tissue or organ in both 

humans and animals. At lower doses, radiation produces a delayed 

response in the form of increased incidence of cancer long after 

the exposure period. This has been documented extensively in 

both humans and animals. Human data are extensive and include 

atomic bomb survivors, many types of radiation-treated patients, 

underground miners, and radium dial workers. Animal data include 

demonstrations in many mammalian species and in mammalian tissue 

cultures. 
Evidence of mutageni-c properties of radiation comes mostly 

from animal data, in which all forms of radiation-induced 

mutations have been demonstrated, mostly in mice. Tissue 

cultures of human lymphocytes have also shown radiation-induced 

mutations. 

Evidence that radiation is a teratogen has been demonstrated 

in animals and in humans. A fetus is most sensitive to radiation 

during the early stages of organ development (between 8 and 15 

weeks for the human fetus). The radiation-induced malformations 

produced depend on which cells are most actively differentiating. 

In conclusion, evidence of the mutagenic and teratogenic 

properties of radiation in man is strong, and for carcinogenesis, 

the evidence is overwhelming. 
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Wa83 Wakabayashi, T., et al., "Studies of the Mortality of 

A-Bomb Survivors, Report 7, Part 111, Incidence of 

Cancer in 1959-1978, based on the Tumor Registry, 

Nagasaki, " Radiat. Res. 93, 112-146, 1983. 

4. MOVEMENT OF RADIONUCLIDES THROUGH 

ENVIRONMENTAL PATHWAYS 

4.1 INTRODUCTION 

When radionuclides are released to the air, they can enter a 

number of pathways leading to human exposure. These environmental 

pathways are shown in Figure 4-1. 

Radionuclides, released in the form of particulates or gases, 

form a plume that disperses down wind (Section 4.2). These 

radionuclides in the air can directly affect people in two ways: 

through external dose caused by photon exposure from the plume, or 

through internal dose resulting from radionuclide inhalation. As 

the airborne radionuclides move from the point of release, they 

(especially those in particulate form) deposit on ground surfaces 

and vegetation as a result of dry deposition and precipitation 

scavenging (Section 4.3). Photon radiation from the radionuclides 

deposited on the ground contributes to the external doses. Finally, 

small fractions of the radionuclides deposited on plant surfaces and 

agricultural land enter the food chains, concentrating in produce 

and in animal products such as milk and meat (Section 4.4). 

Consumption of contaminated foodstuff then contributes to the 

internal doses of radiation to individuals. 

The concentrations of radionuclides in air, on soil surfaces, 

and in food products are calculated using the computer code 

AIRDOS-EPA. A description of the code and some examples of its 

applications, with an overview of the uncertainties, are provided in 

Section 4.5. (See references Ha82, Ti83, and NCRP84 for a more 

detailed description of the processes, modeling techniques, and 

uncertainty estimates. ) 

4.2 DISPERSION OF RADIONUCLIDES THROUGH THE AIR 

4.2.1 Introduction 

Radionuclides entering the atmosphere are transported away from 

their point of release and are diluted by atmospheric processes. To 

perform a radiological assessment, it is necessary to model the 

long-term average dispersion resulting from these processes. This 

is because the sources under consideration release radionuclides at 

rates that are substantially uniform when considered over long 

periods of time, and because the somatic and genetic effects on 

human health are generally treated as being the result of chronic 

exposure over long periods of time. 

As large-scale winds move over the earth's surface, a turbulent 

boundary layer, or mixed layer, is created that controls the 
I AIR DISPERSION 

EXTERNAL DOSE INTERNAL DOSE 

Figure 4-1. Pathways of airborne radionuclides into the 

environment. 
dispersion of the released radionuclides. The depth and dispersion 

properties of the mixed layer, which are highly variable over short 

periods of time, are controlled by two sources of turbulent effects: 

mechanical drag of the ground surface and heat transfer into or from 

the boundary layer. The mechanical drag of the ground surface on 

the atmosphere creates a shear zone that can produce significant 

mechanical mixing. The mechanical mixing is stronger when the wind 

is stronger and the roughness elements (water, grains of dirt, 

grass, crops, shrubs and trees, buildings, etc.) are larger. The 

vertical scale (dimension or thickness) of the mechanical mixing 

zone is related to the size of these roughness elements. Heat 

transfer into or from the boundary layer, the second source of 

turbulent effects, also strongly affects the mixed layer's turbulent 

structure and thickness. Solar heating creates huge rising bubbles 

or thermals near the ground. These large bubbles produce turbulent 

eddies of a much larger scale than those from the mechanical drag of 

the ground surface. With strong solar heating on a clear day, the 

mixing layer may be a few thousand meters deep. On a clear, calm 

night, the boundary layer virtually disappears, so that 

radionuclides (and other pollutants) are dispersed with very little 

turbulent diffusion. 

The objective of the atmospheric transport models used by EPA 

is to incorporate the essential physical data necessary to 

characterize an extremely complex turbulent flow process into a 

simplified model that is adequate to predict the long-term 

dispersion of radionuclide releases. In general, the data necessary 

to implement a detailed theoretical model of atmospheric dispersion 

are not available and would be impractical to obtain. Apart from 

the data problem, the mathematical complexities and difficulties of 

a direct solution to the turbulent dispersion problem are profound 

and beyond the practical scope of routine EPA regulatory 

assessments. The widely accepted alternative has been to 

incorporate experimental observations into a semi-empirical model, 

such as outlined below, that is practicable to implement. 

Three basic meteorological quantities govern dispersion: wind 

direction, wind speed, and stability. Wind direction determines 

which way a plume will be carried by the wind: a wind from the 

northwest moves the plume toward the southeast. Although wind 

direction is a continuous variable, wind directions are commonly 

divided into 16 sectors, each centered on one of the cardinal 

compass directions (e.g., north, north-northeast, northeast, etc.). 

Since there are 16 sectors, each one covers a 22-1/2-degree angle. 

Wind speed directly influences the dilution of radionuclides in the 

atmosphere. If other properties are equal, concentration is 

inversely proportional to wind speed. Customary wind speed 

categories include 0 to 3 knots (lowest speed) to greater than 21 

knots (highest speed). 
Atmospheric stabili,ty, the third meteorological. quantity, 

categorizes the behavicr of a parcel of air when it is adiabaticall-y 

(without heat transfer) displaced in a vertical direction. If the 

displaced parcel would be expected to return toward its original 

position, the category is stable; if it would continue to move away 

from its original position, the category is unstable. Under 

conditions of neutral stability, the parcel would be expected to 

remain at its new elevation without moving toward or away from its 

old one. 

Typically, the unstable classes are associated with conditions 

of very little cloud cover, low wind speeds, and a sun high in the 

sky. The atmosphere is neutral on a windy, cloudy day or night and 

is stable at the surface at night when the sky is clear and wind 

speeds are low. Dilution due to vertical mixing occurs more rapidly 

with increasing distance under unstable conditions than under stable 

ones. Stability categories range from A (very unstable) to D 

(neutral) to G (very stable) . 

A table of joint frequencies (fractions of time) for each 

combination of stability, wind direction, and wind speed is the 

starting point for any assessment of long-term atmospheric 

dispersion. These data are usually obtained by the analysis of 

long-term observations from weather stations or from site- specific 

meteorological facilities. 

4.2.2 Air Dispersion Models 

EPA uses an empirical Gaussian model for most radionuclide 

dispersion calculations. The model also considers such processes as 

plume rise, depletion due to deposition, and radionuclide ingrowth 

and decay. 

Gaussian Plume Model 

The basic workhorse of EPA dispersion calculations is the 

Gaussian model. Several reasons why the Gaussian model is one of 

the most commonly used are quoted below (Ha82): 

"(1) It produces results that agree with experimental data 

as well as any model. 

"(2) It is fairly easy to perform mathematical operations 

on this equation. 

"(3) It is appealing conceptually. 

"(4) It is consistent with the random nature of turbulence. 
"(5) It is a solution to the Fickian diffusion equation for 

constants K and u. 

"(6) Other so-called theoretical formulas contain large 

amounts of empiricism in their final stages. 

"(7) As a result of the above, it has found its way into 

most government guidebooks, thus acquiring a 'blessed' 

(sic) status." 

The long-term Gaussian plume model gets its name from the shape 

presumed for the vertical concentration distribution. For a ground 

level source, the concentration is maximum at ground level and 

decreases with elevation like half of a normal or Gaussian 

distribution. For an elevated release, the concentration is 

symmetrically distributed about the effective height of the plume, 

characteristic of a full Gaussian distribution. Actually, the 

vertical dispersion is limited by the ground surface below and any 

inversion lid above the release (see Figure 4-2). An inversion lid 

is defined by the altitude in the atmosphere where the potential 

temperature begins to increase with increasing height, thus limiting 

the volume of air available for diluting releases. 

At large distances from the point of the release, the 

radionuclide concentration becomes uniformly distributed between the 

ground and the lid. Within each of the 16 direction sectors, the 

concentration is considered to be uniform at any given distance from 

the release. For a ground-level release, the ground-level 

concentration decreases monotonically with distance from the release 

point. For an elevated release, the ground-level concentration 

increases, reaches a maximum value, and then decreases with 

increasing distance from the release point. 

Mathematically, the long-term average dispersion calculation 

used by EPA can be expressed as 

where x/Q (s/m3) is the concentration for a unit release rate at a 

distance x(m) from the release point, h,(m) is the effective height 

of the release, o,(m) is the vertical dispersion parameter 

appropriate to the stability category and distance x, and p(m/s) is 

the wind speed. At distances where the release is uniformly mixed 

between the ground and lid, the expression becomes 
----- 

~~MIDIATI omm ~nos mnAycs 

IORWNO AND YlXlWO IUMI~ORYLV YlXRDI 

L10 AIIICT PIOCILTI I 

I 

t 

Figure 4-2. Vertical concentration profiles for plume vs 

downwind distance from release 
where h (m) is the lid height (meters), and the other quantities are L the same as before. 

Plume Rise Model 

Vertical momentum or buoyancy can cause a plume to rise to an 

effective height that is several times the physical height of the 

release. The momentum flux of a release is proportional to the 

product of the volume flow rate and the vertical exit velocity, 

while the buoyancy flux is proportional to the product of the volume 

flow rate and the difference between the temperatures of the release 

gases and the ambient air. Momentum rise is initially dominant for 

most plumes, even though buoyant rise may become the more important 

process at larger distances. In any case, plume rise increases with 

distance from the release point; the effective height of the plume 

may not reach a limiting value until the plume is several kilometers 

from the point of release. 

Plume Depletion Model 

As radionuclides in the plume are dispersed, their activity is 

depleted by dry deposition and precipitation scavenging. The rate 

of plume depletion due to dry deposition and precipitation 

scavenging is proportional to the deposition rate (see Section 4.3). 

EPA's Office of Radiation Programs uses a source depletion model 

which considers the shape of the vertical concentration profile to 

be unchanged by depletion. Depletion due to deposition generally 

does not cause more than half of the released activity to be removed 

at a distance of 80 km. Depletion by precipitation scavenging 

occurs only during periods of precipitation. 

Radiological Decay and Ingrowth 

Radiological decay can also reduce the radionuclide 

concentration in the plume. A typical elapsed time for traverse 

between the point of release and a receptor located 80 km away is 

about 5 hours. Thus, only nuclides with short half-lives would be 

appreciably depleted by radiological decay. For example, Ar-41, 

which has a 1.8 hour half-life, decays to about 15 percent of its 

original activity in 5 hours. When a released radionuclide is a 

parent for other radionuclides in a chain, those decay products will 

become part of the plume's activity even though they were not 

released by the source. For example, Cs-137 is the parent of 

Ba-137m, which has a half-life of about 2.6 minutes. The Ba-137m 

activity would reach 90 percent of that of the Cs-137 in about 8.5 

minutes, the time required at a typical wind speed of 5 m/s for the 

release to travel about 2.5 km. For many nuclides, the radiological 

effects associated with exposure to decay products are at least as 

important as those from exposure to the parent. For example, the 
external photon dose from a release of Cs-137 is entirely due t.o 

photons from its decay product Ba-13711. 

4.2.3 Uncertainties in Atmospheric Dispersion Modeling 

EPA must deal with several uncertainties in its modeling of 

atmospheric dispersion. Two basic considerations contribute to 

these uncertainties. The first involves the parameters that enter 

into the model and how well they are known or can be determined for 

a particular situation. The presumption is that the basic 

assumptions for which the model was developed are satisfied and that 

the uncertainty of predicted concentrations depends primarily on the 

uncertainty of the data used in the calculations. The second 

consideration involves the use of a modeling technique under 

conditions that do not satisfy the basic assumptions for which the 

model was developed. Such use may be the only practicable 

alternative available for assessing atmospheric dispersion, but the 

principal uncertainties are now related to evaluating the 

significance of these effects that are not considered in the model. 

An example of this would be the use of the Gaussian plume model, 

whlch was developed for short distances over an open, flat terrain, 

to assess dispersion over large distances or in a complex terrain 

dominated by hills and valleys. 

In regard to the first consideration, the authors of NCRP84 

concluded that the appropriate basic parameters, such as wind speed 

and direction,, can be determined accurately enough so that they are 

not major contributors to model uncertainty. However, the 

uncertainties associated with derived parameters (such as stability 

class) or lumped parameters (such as those used to characterize 

deposition, resuspension, or building wake effects) can dominate the 

model uncertainties. 

The effect of the uncertainty of an input variable can strongly 

or weakly influence the model output depending upon circumstances. 

For example, the effective height of a release, h,, can be estimated 

using a plume rise model to within a factor of about 1.4 (NCRP84). 

From equations 4-1 and 4-2, it is clear that when 0, is much smaller 

than h,, the effect of this uncertainty on equation 4-1 is strong; 

whereas at large distances where equation 4-2 is appropriate, the 

value of he has little effect on the calculated concentration. 

Little and Miller (Li79 and Mi82) have surveyed a number of 

validation studies of atmospheric dispersion models. Although 

these studies provide limited data, they indicate an uncertainty of 

approximately a factor of 2 for annual average concentrations for 

locations within 10 km of the release and approximately a factor of 

4 (77 percent of their samples) to 10 (92 percent of their samples) 

for locations between 30 and 140 km of the release. The validation 

studies were for fairly complex terrain, i.e., substantial hills and 
valleys, but not extreme conditions of either terrain or 

meteorology. 

4.3 DEPOSITION OF ATMOSPHERIC RADIONUCLIDES 

4.3.1 Introduction 

Atmospheric deposition includes a complex set of processes that 

result in the transfer of radionuclides from the plume to the ground 

surface and vegetation. Processes are categorized as "dry" when 

they result in the direct transfer from the plume to the surfaces in 

contact with it and "wet" when the transfer is first from the plume 

to precipitation and then from the precipitation to the ground or 

vegetation surfaces. 

4.3.2 Dry Deposition Model 

Dry deposition models generally relate the surface deposition 

flux to the air concentration at some reference height, typically 

1 meter above the ground. The resulting equation is 

where W is the deposition flux to the surface (Ci/m2s), X, is the 

reference height air concentration (ci/m3), and v, is the deposition 

velocity (m/s). Although v, has the units of a velocity (hence its 

name), it is a lumped variable relating the deposition flux to the 

air concentration. The value of the deposition velocity depends on 

a complex interaction of effects--atmospheric, aerosol, and surface 

(canopy). Thus, while the deposition velocity is often assigned a 

simple fixed value, it actually represents the result of a diverse 

combination of effects. 

4.3.3 Wet Deposition Model 

Wet deposition models relate the flux due to precipitation 

scavenging to the concentration in the plume. Since the activity 

scavenged from the plume by an element of precipitation is presumed 

to remain with the precipitation element until reaching the ground 

surface, the deposition flux is proportional to the total. wetted 

activity in a vertical segment of the plume (ci/m2). The resulting 

equation can be expressed as 

- 

w = h,, x L (4-4) 

- where W is the surface flux (Ci/mzs), Xis the average wetted air 

concentration (Ci/m3), L is the depth of the wetted layer (m), and 

h., is the scavenging rate (s-'). h,, is a variable that lumps 

together the complex interactions between precipitation and the 
plume. Because the deposition flux is proportional to the 

vertically integrated concentration (i.e., the total activity in a 

column of unit ground surface area), it is independent of the 

effective height of the release. Raising the effective height of a 

release may lower the dry deposition flux but leaves the flux 

resulting from precipitation scavenging unchanged. 

4.3.4 Soil Concentration Model 

The deposited radionuclides accumulate in the surface soil 

until they are removed either by radiological decay or by processes 

such as leaching. The areal concentration can be expressed as 

where C, is the areal concentration (Ci/m2), W is the radionuclide 

flux to the ground surface (Ci/mzs) , t, (s) is the time for 

radionuclide buildup in soils, and h, is the effective removal rate 

from soil (s-'). When the deposited radionuclide is the parent of 

other radionuclides, their soil concentrations at time t, due to 

ingrowth from the parent must also be calculated. For calculating 

root transfer to crops, the radionuclide concentration in the 

surface soil layer can be expressed as 

where C, is the soil concentration (Ci/kg) and P is the areal 

density of dry soil (kg/m2) for the plowed or mixed soil layer. 

The value of t,, the deposition accumulation time, is typically 

in the range of 20 to 100 years. For nearby individual assessments, 

t, is chosen to correspond to the expected operational life of the 

facility. If EPA considers it likely that the facility would be 

replaced by another similar one at that time, then t, is increased 

accordingly up to a maximum value of 100 years. Of course, only 

those environmental concentrations that depend on soil deposition 

are affected by the choice of t,. For collective (population) 

assessments, a value of 100 years is used for t,. This value 

corresponds to establishing a 100-year cutoff for the time following 

a release when any significant intake or external exposure 

associated with deposition on soil might take place. Since 

radionuclide inhalation is generally the dominant risk pathway, 

total risk is not sensitive to the choice of t,. 

The value of h, is the sum of the radiological decay constant, 

h, and an environmental removal rate for deposited radionuclides 
from soil, h,. Hoffman and Baes (No79) considered a simplified 

leaching-loss model appropriate to agricultural soil for calculating 

h.. Their range of values for the parameter K, (the equilibrium 

distribution coefficient relating the ratio of the radionuclide 

concentration in soil water to that on soil particles) for Cs is 

from 36.5 to 30,000 ml/g. The corresponding ratio of & is 820:l. 

The uncertainty in h, is also significantly affected by the 

uncertainty in the other parameters. Although their model is a 

reasonable one, adequate studies for its validation do not exist. 

Since the choice of appropriate values for h, is so uncertain, EPA 

has used 0.2 y-' as a genetal nominal value (the geometric mean of h, 

for Put, I-, Csi, and Sr2+ ions is 1.2x10-' y-' using Hoffman and Baes 

median data values) and a value of 0.1 yT1 for urban settings where 

strong surface runoff would be expected to increase the effective 

removal rate. 

4.3.5 Uncertainties 

Uncertainties in v, and h,, are substantial; NCRP84 lists 

measured values of v, which vary over three orders of magnitude. 

Hanna et al. note that "The use of scavenging coefficient for wet 

removal modeling is probably best regarded as an order of magnitude 

estimation procedure" (Ha82). Actually, much of the wide range of 

values reflects measurement uncertainties as well as actual 

variations. Furthermore, most field deposition measurements reflect 

short-term or episodic studies rather than long-term observations. 

Miller and Little (Mi82) concluded that the data necessary to 

quantify the accuracy of calculated ground concentrations are not 

currently available. 

4.4 TRANSPORT THROUGH THE FOOD CHAIN 

4.4.1 Introduction 

Deposited radionuclides may become associated with vegetation 

by two principal routes: (1) direct interception of a fraction of 

the deposited activity by plant surfaces, and (2) transfer of 

deposited activity from the soil through the plant's root system. 

Radionuclides in animal feed crops such as pasture grass or stored 

feeds can be transferred to foods such as milk and meat. 

4.4.2 Concentration in Vegetation 

The radionuclide concentrations in plants due to interception of the 

deposition flux can be calculated as (Ba76) 
where C: is the crop concentration (Cilkg) at harvest, W is the 

deposition flux (ci/m2s), f, is the fraction of the deposition flux 

which the vegetation intercepts, Y, is the vegetation yield (kg/m2), 

T, is a translocation factor, h, is the effective removal rate of 

the intercepted radionuclide from the vegetation (s-I), and t, is the 

exposure time of the vegetation to the radionuclide flux (s). 

Miller (Mi79) has observed that data for f, and Y, are well 

represented by the expression 

where y was found to range between 2.3 and 3.3 m2/kg when Y, is 

expressed in kg/m2, dry. Since the product 'yY, is generally less 

than 1.0, for many practical purposes equation 4-8 can be 

- approximated as 

In this case, the quantity f,/Y, (4-7) can be replaced by y 

which shows much less environmental variation than f, and Y, do 

separately. Note that Y, is the total vegetative yield which can be 

several times the edible portion yield for a crop. T,, the - 

- translocation factor, relates the radionuclide concentration in the 

edible portion to that in the entire plant. Baker et al. (Ba76) 

suggest a value of 1.0 for leafy vegetables and fresh forage, and 

0.1 for all other produce. (A value of 1.0 is used for all crops in 

AIRDOS-EPA. ) 

The value for h, is the sum of h, the radionuclide decay 

constant and h,, the weathering rate factor. For a typical 

weathering half-life of 14 days, has a value of 5.7~10.' s-I. In 

general, the product h, te >1 and equation 4-9 can be simplified to 

Radionuclides also transfer directly from the soil to 

vegetation through the plant's root system. The plant concentration 

due to this process can be calculated as 

C: = C, B,, (4-11) 

where C: is the plant concentration at harvest (Ci/kg), C, is the 

soil concentration (Ci/kg), and B,, is the element-specific soil to 

plant transfer factor. The total concentration from both processes 

is 
Generally, the contribution of C:' to C, is greater than that of 

C: for atmospherically dispersed radionuclides. 

4.4.3 Concentration in Meat and Milk 

For a concentration C, (Ci/kg) in animal feed, the concentration 

in meat C, (Ci/kg) can be calculated as 

where Q, is the animal's feed consumption (kg/d) and F, is the feed 

to meat transfer factor (d/kg). F, is element dependent and 

represents the average mean concentration at slaughter for a unit 

ingestion rate over the animal's lifetime. Most systematic studies 

of F, have been made for cattle or other ruminants, although a few 

measurements for other species also exist (NCRP84). In practice, 

even the F, values for beef are often based on collateral data 

(Ba84) . 

Similarly for milk, the concentration C, (Ci/L) can be 

calculated as 

where F, (d/L) is the equilibrium transfer factor to milk and the 

other parameters are as for equation 4-13. Although more 

statistical data are available for F, than for F,, the estimation of 

transfer coefficients to animal products is a subject needing both 

integration and better documentation (NCRP84). 

4.4.4 Summary 

Radionuclide intake through the food chain depends upon both 

the concentration in food and human usage. The concentration in 

food depends upon the food source use of foods grown in proximity to 

the release location, the fraction of an individual's food that is 

home produced and other factors that can strongly influence the 

significance of the food pathway. Unfortunately, generally useful 

validation studies to quantify the substantial uncertainties in the 

food chain have not been made. References such as NCRP84, Ti83, 

Mi82, and Li79 cite ranges for some parameters and make limited 

model uncertainty estimates but do not make quantitative evaluations 

of the uncertainties for the ingestion pathway taken as a whole. 

EPA has chosen a factor of 10 as a reasonable upper bound for 

the uncertainty in both the deposition rate model and the calculated 

intake from eating food containing deposited radionuclides. 

Assuming that the two factors are independent, uncorrelated, and 

correspond to the 2 sigma values for a log normal distribution, the 
combined uncertainty for the pathway (deposition and intake of 

radionuclides from food) is a factor of 26.' EPA has rounded this 

value to 30 as an estimate of the overall food pathway uncertainty 

factor. 

4.5 CALCULATING THE ENVIRONMENTAL CONCENTRATION OF 

RADIONUCLIDES: THE AIRDOS-EPA CODE 

4.5.1 Introduction 

Environmental concentrations of radionuclides calculated by EPA 

may be site specific, meaning that available data relevant to the 

site are incorporated into the assessment. Or an assessment may be 

generic; that is, an assessment of a hypothetical facility at a 

location considered an appropriate possibility for such a facility 

class. Frequently, EPA performs site-specific assessments for 

existing facilities, e.g., a national laboratory. In addition, EPA 

often employs generic assessments in evaluating alternative sitings 

for a proposed facility or assessing a widespread class of 

facilities, e.g., industrial coal-burning boilers. 

In any case, EPA makes both individual and collective 

(population) assessments. The purpose of the individual assessment 

is to assess doses and lifetime risk to individuals living near a 

facility. EPAfs assumption is that these individuals reside at the 

same location much of their lives and that their exposures extend 

from infancy on through adulthood. The doses and risks calculated 

are expectation values, i.e., the estimates are intended to be 

typical for a person living a long period of time under the assessed 

conditions. EPA's collective (or population) assessments evaluate 

doses and risks to a population that may be regional (typically up 

to 80 km distant), long-range (e.g., the coterminous United States), 

or worldwide as appropriate. The risk is usually expressed as the 

expected number of premature deaths in the population per year of 

facility operation. 

4.5.2 AIRDOS-EPA 

EPA has used the AIRDOS-EPA code (Mo79) to calculate 

environmental concentrations resulting from radionuclide emissions 

into air. The results of this analysis are estimates of air and 

ground surface radionuclide concentrations; intake rates via 

inhalation of air; and ingestion of radioactivity via meat, milk, 

and fresh vegetables. The atmospheric and terrestrial transport 

models used in the code, their implementation, and the applicability 

of the code to different types of emissions are described in detail 
* * * DRAFT * * * 

in Mo79. Input to AIRDOS-EPA is extensive, but its preparation can 

be facilitated by using the preprocessor PREPAR (Sj84). Appendix A 

of this document summarizes many of the default values and 

assumptions used in EPA's assessments. 

AIRDOS-EPA calculates atmospheric dispersion for radionuclides 

released from one to six stacks or area sources. Radionuclide 

concentrations in meat, milk, and fresh produce are estimated by 

coupling the deposition rate output of the atmospheric dispersion 

models with the Regulatory Guide 1.109 (NRC77) terrestrial food 

chain models. Radionuclide concentrations for specified distances 

and directions are calculated for the following exposure pathways: 

(1) immersion in air containing radionuclides, (2) exposure to 

ground surfaces contaminated by deposited radionuclides, (3) 

inhalation of radionuclides in air, and (4) ingestion of food in the 

area. The code may be used to calculate either annual individual 

exposures or annual population exposures at each grid location. For 

either option, AIRDOS-EPA output tables summarize air concentrations 

and surface deposition rates as well as the intakes and exposures 

for each location. In addition, working level exposures are 

calculated and tabulated for evaluating the inhalation of 

short-lived progeny of Rn-222. 

Assessment Grid 

AIRDOS-EPA has provision for either a rectangular or a circular 

calculational grid. The customarily used circular grid (see Figure 

4-3) has 16 directions proceeding counterclockwise from north to 

north-northeast. The user chooses the grid distances. Generally, 

successive distances are chosen with increasing spacing. It is 

important to realize that the calculational grid distances and the 

set of distances associated with population and food production data 

are one and the same. Hence, the concentration calculated for each 

grid distance must be the appropriate average value for the 

corresponding range of distances covered by the population and 

agricultural data. Choosing a suitable set of grid distances may 

require different compromises of convenience for different 

assessments and may be different for individual and collective 

assessments of the same facility. 
X - Assessment grid locations at up to 20 distances 

(2 shown) and 16 directions (5 shown) 

Figure 4-3. Circular grid system used by AIRDOS-EPA. 
Environmental P,ccumulation -- Time 

An AIRDOS-EPA assessment is based on what can be viewed as a 

snapshot of environmental concentrations after the assessed facility 

has been operating for some period of time. The choice of an 

environmental accumulation time affects only those pathways 

dependent on terrestrial concentrations, i.e., ground surface 

exposure and food intakes. Usually, the accumulation time for an 

individual assessment is chosen to be consistent with the expected 

life of the facility (or 100 years when a similar facility might be 

expected to replace the present one at the end of its useful life). 

For collective assessments, 100 years is customarily used. 

Source Considerations 

Point sources are characterized by their physical height and, 

when desired, the parameters to calculate buoyant or momentum plume 

rise using Brigg's (Br69) or Rupp's (Ru48) formulations 

respectively. Alternatively, a fixed plume rise may be specified 

for each Pasquill-Gifford atmospheric stability class A through G. 

The area source model is similar to that of Culkowski and 

Patterson (Cu76) and transforms the original source into an annular 

segment with the same area. At large distances, the transformed 

source approaches a point source at the origin, while at distances 

close to the origin, it approaches a circle with the receptor at its 

center. 

Building wake effects and downwash are not included in the 

AIRDOS-EPA models. The same type of rise calculation (buoyant, 

momentum, or fixed) is used for all sources. As many as six sources 

may be assessed, but for calculational purposes, they are all 

considered to be co-located at the origin of the assessment grid. 

Radionuclide Releases 

Releases for up to 36 radionuclides may be specified for 

AIRDOS-EPA. Each release is characterized by the radionuclide name, 

effective decay constant during dispersion, precipitation scavenging 

coefficient, deposition velocity, and settling velocity, as well as 

the annual activity release for each source. Decay products that 

are significant for the assessment of a radionuclide must be 

included in the list of releases. There is no explicit method for 

calculating radionuclide ingrowth during atmospheric dispersion in 

AIRDOS-EPA. 

Parameters such as particle size, respiratory clearance class, 

and gastrointestinal absorption factor (f,) are passed on for use in 

the DARTAB (Be81) dose and risk assessments as described in Chapters 

5 and 6. 
The approach ORP has used for calculating a precipitation 

scavenging coefficient is based on Slinn's (S177) equation 32: 

h,, = Po c J E(a,%) 

where h., is the scavenging coefficient, c is a constant (Slinn uses 

0.5), J, is the rainfall rate, and E is the collection efficiency 

for a particle of radius a by drops of characteristic radius %. 

Slinn (S177, p. 23) considers the effects of dry deposition and 

interprets Dana and Wolf's (Da68, Wo69, Da70) data as supporting a 

value for E of 0.2, essentially independent of particle size. 

Adopting Slinn's typical value of q, for a frontal rain (0.3 mm) and 

selecting a long-term average value of 1,000 mm/yr (3.16~10-~ mm/s) 

for J,, we obtain: 

h,, = 0.5 (3.16~10-~~ 

0.3 

- This value has been rounded to s-' as a working value for 

the precipitation scavenging coefficient and then scaled according - to the annual precipitation at the assessment location for use in 

AIRDOS-EPA. There is substantial uncertainty in interpreting 

environmental scavenging data, and this estimate is accurate to 

within an order of magnitude. The EPA scaling procedure reflects 

the premise that the variation of rainfall from one location to 

another depends more on rain frequency than on intensity during 

rainfall episodes. 

Dispersion 

Wind and stability class frequencies for each direction are the 

primary data for calculating atmospheric dispersion. The required 

data for AIRDOS-EPA are calculated from a joint frequency 

distribution of wind speed and atmospheric stability class for each 

direction. Inasmuch as the assessments require long-term average 

dispersion values, the sector-averaged Gaussian plume option is 

used. The vertical dispersion parameter (0,) is calculated using 

Brigg's formulas (Gi76). Vertical dispersion is limited to the 

region between the ground and a mixing depth lid. The harmonic mean 

of Holzworth's (Ho72) morning and afternoon mixing depths is 

customarily employed for this value; that is, 
where E, andl, are respectively the morning and afternoon mixing 

depths and h is their harmonic mean. At large distances, the 

concentration is uniform between the ground and the lid. 

Deposition Rate 

AIRDOS-EPA models both dry and wet deposition processes. 

Resuspension, the reintroduction of deposited material into the 

atmosphere, is not modeled in AIRDOS-EPA. The dry deposition rate 

is the product of the deposition velocity and the near ground-level 

air concentration, while the wet deposition rate is the product of 

the precipitation scavenging coefficient and the vertically 

integrated air concentration. Wet deposition decreases 

monotonically with distance and is independent of the effective 

release height of the source, while the effect of source height can - 

. be significant for dry deposition. For locations close to an 

elevated source, wet deposition can provide the principal source of . radionuclide exposure. Concentrations are adjusted for depletion - due to deposition at each downwind distance. 

Ground Surface Concentration 

- AIRDOS-EPA calculates the ground surface concentration from the - total (dry plus wet) deposition rate. The soil concentration is 

calculated by dividing this value by the effective agricultural soil 

surface density (kg/m2). Both concentrations are calculated for the 

end of the environmental accumulation time t, and can include the 

ingrowth from deposited parent radionuclides as well as removal due 

to radiological decay and environmental processes such as leaching. 

Ingrowth from a parent radionuclide is calculated using a decay 

product ingrowth factor. The ingrowth factor is the equivalent 

deposition rate for a unit deposition rate of the parent 

radionuclide. For example, the ingrowth factor for Pb-210 as a 

parent of Po-210 would be calculated by determining the 

concentration of Po-210 at time t, due to a unit deposition rate of 

Pb-210 and dividing it by the corresponding concentration for a unit 

deposition rate of Po-210. These ingrowth factors must be calculated 

in advance of running AIRDOS-EPA and are dependent on both the 

accumulation time t, and the soil removal constants for the nuclides 

in the radionuclide chain (Pb-210, Bi-210, and Po-210 in this case). 

Concentrations in Food 

Radionuclide concentrations in food are calculated using 

essentially the same model as in NRC Regulatory Guide 1.109 (NRC77). 

Changes from that model include consideration of environmental 

removal from the root zone, and separate values for food and pasture 
crops of the interception fraction, areal yield, and soil-to-plant 

transfer values. Concentration calculations for meat and milk use 

the same models as the Regulatory Guide model. There are numerous 

parameters in the terrestrial pathways model. Appendix A of this 

volume of the BID contains tables of values used in these 

assessments. 

Population and Agricultural Data 

For a collective (population) assessment, population and 

agricultural data for each grid location must be provided. EPA uses 

the 1970 census enumeration district data to calculate population 

distributions. AIRDOS-EPA calculates the collective assessment for 

agricultural products based on consumption by the assessment area 

population. The assessment can be based on agricultural production 

by choosing utilization factors large enough to ensure that all 

items produced are consumed. 

Food Utilization Factors 

In addition to the consumption rate for different food 

categories (leafy vegetables, other produce, meat, and milk), the 

user may specify the fraction of vegetables, meat, and milk that are 

(1) home grown, (2) produced in the assessment area, or (3) imported 

from outside the assessment area. Those in the third category are 

considered to contain no radionuclides. Those from the second 

category have the average concentration for that category produced 

within the assessment area, while concentrations for the first 

category are those that would occur at each grid location. Appendix 

A of this volume provides some typical food source fractions for 

urban and rural assessment areas. Note that if the assessment 

considers food to be only home grown or imported from outside the 

assessment area, then the actual quantity of food produced at each 

location is not relevant to the assessment. Experience has shown 

that the ingestion doses and risks for the nearby individual are 

usually dominated by the radionuclide intake from home-grown food, 

and hence there is generally no significant difference between 

assuming that food that is not home grown is obtained from the 

assessment area or is imported from outside the assessment area. 

Special Radionuclides 

Special consideration is given to the radionuclides H-3 

(tritium), C-14, and Rn-222. The specific activity of tritium in 

air (pCi/g of H,O) is calculated for an absolute humidity of 8 mg/m3 

(NRC77). Etnier (Et8O) has calculated average absolute humidities 

for over 200 U.S. locations. The 8 mg/m3 value would be within a 

factor of 2 for most of them. The specific activity of atmospheric 

C-14 (pCi/g of carbon) is calculated for a CO, concentration of 330 

ppm by volume (Ki78). Concentrations of these nuclides in 
vegetation are cal..cul.ateci on the assuntptioil thaL the water and 

carbon content in vegetation are front the atmosphere and have the 

same specific activity as in the atmosphere. The Rn-222 

concentration in air is replaced by its short-lived decay product 

concentration in working level units using a fixed equilibrium 

fraction (typically 0.5 for calcul.ating populatiori health risks). 
REFERENCES 

Baker, D.A., Hoenes, G.R., and Soldat, J.K., "FOOD-An 

interactive code to calculate internal radiation doses 

from contaminated food products," in Proceedings of the 

Conference on Environmental Modeling and Simulation, Ott, 

W.R., editor, EPA 600/9-76-016, p. 204, Office of 

~esearch ~evklo~ment and Office of Planning and 

Management, U.S. Environmental Protection Agency, 

Washington, D.C., July 1976. 

Baes, C.F. 111, Sharp, R.D., Sjoreen, A.L., and Shor, 

Begovich, C.L., Eckerman, K.F., Schlatter, E.C., Ohr, 

S.Y., and Chester, R.O., DARTAB: A program to combine 

airborne radionuclide environmental exposure data with 

dosimetric and health effects data to generate tabulation 

of predicted impacts, ORNL/5692, Oak Ridge National 

Laboratory, Oak Ridge, Tenn., August 1981. 

Briggs, G.A., Plume Rise, TID-25075, U.S. Atomic Energy 

Commission Critical Review Series, National Technical 

Information Service, Springfield, Va., November 1969. 

Culkowski, W.M. and Patterson, M.R., A Comprehensive 

Atmospheric Transport and Diffusion Model, 

ORNL/NSF/EATC-17, National Oceanic and Atmospheric 

Administration, Atmospheric Turbulence and Diffusion 

Laboratory, Oak Ridge, Tenn., 1976. 

Dana, M.T. and Wolf, M.A., "Experimental Studies in 

Precipitation Scavenging," in Pacific Northwest 

Laboratory Annual Report for 1967 to the USAEC Division 

of Biology and Medicine, Vol. 11, Physical Sciences, Part 

3, Atmospheric Sciences, Simpson C.L. et al., US.AEC 

~e~ort BNwL-~I~-~, pp. 128-140, Battelle Pacific 

Northwest Laboratories, Richland, Wa., October 1968. 

Dana, M.T., Wolf, M.A., and duplessis, L.A., "Field 

Experiments in Precipitation Scavenging," in Pacific 

Northwest Laboratory Annual Report for 1969 to the USAEC 

Division of Biology and Medicine, Vol. 11, Physical 

Sciences, Part 1, Atmospheric Sciences, Simpson, C.L. et 

al., USAEC Report BNWL-1307 (Pt. I), pp. 77-81, Battelle 

Pacific Northwest Laboratories, June 1970. 
Etnier, E.L., "Regional and site-specific absolute 

humidity data for use in tritium dose calculations," 

Health Phys. 39, 318-320, 1980. 

Gifford, F.S. Jr., "Turbulent Diffusion-Typing Schemes: A 

Review," Nucl. Saf. 17(1), 68-86, 1976. 

Hanna, S.R., Briggs, G.A., and Hosker, R.P. Jr., 

Handbook on Atmospheric Diffusion, DOE/TIC-11223, 

Technical Information Center, U.S. Department of'~nergy, 

Washington, D.C., January 1982. 

Holzworth, G.C., Mixing Heights, Wind Speeds an Potential 

for Urban Air Pollution Throughout the Contiguous United 

States, Publication No. AP-101, U.S. Environmental 

Protection Agency, Office of Air Programs, Research 

Triangle Park, N.C., 1972. 

Hoffman, F.O. and Baes, C.F. 111, A Statistical 

Analysis of Selected Parameters for Predicting Food Chain 

Transport and Internal Dose of Radionuclides, 

NUREG/CR-1004, Oak Ridge National Laboratory, Oak Ridge, 

Tenn., 1979. 

Killough, G.C. and Rohwer, P.S., "A new look at the 

dosimetry of 14c released into the atmosphere as carbon 

dioxide," Health Phys., 34, 141-159, 1978. 

Little, C.A. and Miller, C.W., The Uncertainty 

Associated with Selected Environmental Transport Models, 

ORNL-5528, Oak Ridge National Laboratory, Oak Ridge, 

Tenn., November 1979. 

Miller, C.W. and Little, C.A., A Review of Uncertainty 

Estimates Associated with Models for Assessing the Impact 

of Breeder Radioactivity Releases, ORNL-5832, Oak Ridge 

National Laboratory, Oak Ridge, Tenn., August 1982. 

Moore, R.E., Baes, C.F. 111, McDowell-Boyer, L.M., 

Watson, A.P., Hoffman, F.O., Pleasant, J.C., and Miller, 

C.W., AIRDOS-EPA: A Computerized Methodology for 

Estimating Environmental Concentrations and Dose to Man 

from Airborne Releases of Radionuclides, EPA 520/1-79-009 

(reprint of ORNL-5532), U.S. Environmental Protection 

Agency, Office of Radiation Programs, Washington, D.C., 

December 1979. 
NCRP 8 4 Nat:l.onal Council on Radiation Protection and 

Measurements, Radiol.ogical . Assessment : Predictin~the A 

- 

Tr-ort. ~ioaccumulation, and Uptake by -- Man OF 

Radionuclides Releascd - to the Environment, NCRP Report 

No. 76. National Council on Radiation Protection and 

Measurement, Bethesda, Md., March 1984. 

U.S. Nuclear Regul-atory Commission, "Calculation of 

Annual Doses to Man from Routine Releases of Reactor 

Effluents for the Purpose of Evaluating Compliance with 

10 CFR Part 50 Appendix I (Revisio~ I)," Regulatory Guide 

1.109, Office of Standards Development, Washington, D.C., 

October 1977. 

Ru48 Rupp, A.F., Beall, S.E., Bornwasser, L.P., and Johnson, 

D.H., Dilution of Stack Gases in Cross Winds, USAEC 

Report AECD-1811 (CE-l620), Clinton Laboratories, 1948. 

5177 Slinn, W.G.N., "Precipitation Scavenging: Some Problems, 

Approximate Solution, and Suggestions for Future 

Research," in Precipitation Scavenging (1974), 

CONF-741003, Technical Information Center, Energy 

Research and Development Administration, Washington, 

D.C., June 1977. 

S j84 S joreen, A. L., and Miller, C.W., PREPAR-A User Friendly 

Preprocessor to Create AIRDOS-EPA Input Data Sheets, 

ORNL-5952, August 1984. 

Ti83 Till, J.E. and Meyer, K.R., Radiological Assessment, 

NUREG/CR-3332, ORNL-5968, Division of Systems 

Integration, Office of Nuclear Reactor Regulation, U.S. 

Nuclear Regul.atory Commission, Washington, D.C., 

September 1983. 

Wolf, M.A. and Dana, M.?'., '"Experimental Studies in 

Precipitation Scavenging," in Pacific Northwest 

Laboratory Annual Report for 1968 to the USAEC Division 

of Biology and Medicine, Vol. 
5. RADIATION DOSIMETRY 

5.1 INTRODUCTION 

The setting of standards for radionuclides requires an 

assessment of the doses received by individuals who are exposed 

by coming into contact with radiation sources. Two forms of 

potential radiation exposures can occur from these sources -- 

internal and external. Internal exposures can result from the 

inhalation of contaminated air or the ingestion of contaminated 

food or water. External exposures can occur when individuals are 

immersed in contaminated air or water or are standing on 

contaminated ground surfaces. Internal or external doses can 

result from either direct contact with the radiation from 

radionuclides at the site area or from radionuclides that have 

been transported from these sites to other locations in the 

environment. The quantification of the doses received by 

individuals from these radiation exposures is called radiation 

dosimetry. This chapter highlights the internal and external 

dosimetric models used by EPA to assess the dose to individuals 

exposed to radionuclides. 

The models for internal dosimetry consider the quantity of 

radionuclides entering the body, the factors affecting their 

movement or transport through the body, and the energy deposited 

in organs and tissues from the radiation that is emitted during 

spontaneous decay processes. The models for external dosimetry 

consider only the photon doses to organs of individuals who are 

immersed in air or are exposed to a contaminated ground surface. 

In addition, the uncertainties associated with each model will be 

discussed. 

5.2 BASIC CONCEPTS 

Radioactive materials produce radiation as their constituent 

radioactive nuclides undergo spontaneous radioactive decay. The 

forms of emitted energy are characteristic of the decay process 

and include energetic charged particles (alpha and beta 

particles) and photons (gamma rays and x-rays). Alpha particles 

are nuclei of helium atoms and carry a positive charge two times 

that of an electron. These particles can produce dense ionization 

tracks in the biological material that they traverse. Beta 

particles are electrons or positrons emitted in radioactive 

decay. Their penetration power in material is greater than that 

of alpha particles. Gamma and x-rays are electromagnetic 

radiation and are distinguishable from alpha and beta particles 

by their greater penetrating power in material. 
This section introduces some terminology used in Chapters 

5 and 6 to describe internal and external dosimetry. For a more 

detailed explanation, the reader is referred to reports published 

in this area by the International Commission on Radiation Units 

and Measurements (ICRU80), International Commission on 

Radiological Protection (ICRP84), and National Council on 

Radiation Protection and Measurements (NCRP71). 

Activity 

The activity of a sample of any radionuclide of species, i, 

is the rate at which the unstable nuclei spontaneously decay. If 

N is the number of unstable nuclei present at a certain time, t, 

its activity, A,(t) , is given by 

where h: is the radioactive decay constant. The customary unit 

of activity is the curie (Ci); its submultiples, the millicurie 

(mCi), the microcurie (pCi), and the picocurie (pCi), are also 

often used. The curie, which is defined as 3.7x10L0 

disintegrations per second, is the approximate activity of 1 gm 

of Ra-226. 

The time variation of the activity can be expressed in the 

form: 

Ai(t) =Aoi exp(- h: t). (5-2) 

A,, is the activity of nuclide i at time t=O. For a sample 

of radioactive material containing more than one radionuclide, 

the total activity is determined by summing the activities for 

each radionuclide: 

Radioactive Half-Life 

From the above equations, it is apparent that the activity 

exponentially decays with time. The time when the activity of a 

sample of radioactive material containing species i becomes onehalf its original value (i.e., the time t that A,(t) = A,,/2) is 

called its radioactive half-life, T:, and is defined as: 
The unit for the radioactive half-life is any suitable unit 

of time such as seconds, days, or years. The specific activity 

of a radionuclide (the activity per unit mass) is inversely 

proportional to the half-life and can vary over many orders of 

magnitude. 

Radionuclide Chains 

Radionuclides decay either to stable atoms or to other 

radioactive species called daughters. For some species, a decay 

chain of daughter products may be produced until stable atoms are 

formed. For example, Sr-90 decays by emitting a beta-particle, 

producing the daughter Yr-90, which also decays by beta emission 

to form the stable atom Zr-90: 

90 Sr (28.6 yr) ~?OY (64.0 h) -!, "Zr (stable) (5-5) 

Biological Half-Life 

The biological half-life of radionuclides is the time 

required for biological tissues to eliminate one-half of the 

activity by elimination processes. This time is the same for 

both stable and radioactive isotopes of any given element. 

Internal and External Exposures to Radionuclides 

The term "exposure"', in the context of this report, denotes 

physical interaction of the radiation emitted from the 

radioactive material with cells and tissues of the human body. 

An exposure can be "acute" or "chronic" depending on how long an 

individual or organ is exposed to the radiation. Internal 

exposures occur when radionuclides, which have entered the body 

through the inhalation or ingestion pathway, deposit energy to 

organ tissues from the emitted gamma, beta, and alpha radiation. 

External exposures occur when radiation enters the body directly 

from sources located outside the body, such as radiation from 

material on ground surfaces, dissolved in water, or dispersed in 

the air. 

In general, for sources of concern in this report, external 

exposures are from material emitting gamma radiation. Gamma rays 

are the most penetrating of the emitted radiations, and external 

gamma ray exposure may contribute heavily to radiation doses to 

the internal organs. Beta and alpha particles are far less 
penetrating and deposit their energy primarily on the skinrs 

outer layer. Consequently, their contribution to the absorbed 

dose to the total body, compared to that deposited by gamma rays, 

is negligible and will not be considered in this report. 

Absorbed Dose and Absorbed Dose Rate 

The radiological - quantity absorbed dose, D, denotes the mean 

energy imparted A&, by ionizing radiation to a small finite mass 

of organ tissue with a mass, Am, and is expressed as 

- - D = de/dm = lirn (A EJ A m) . A m4 

Internal and external exposures from radiation sources are 

not usually instantaneous but are distributed over extended 

periods of time. The resulting time rate of change of the 

absorbed dose to a small volume of mass is referred to as the 

absorbed dose rate, D: 

D = dD/dt = lirn (A D/A 1). 

A L-40 

The customary unit of absorbed dose rate is any quotient of 

the rad (or its multiple or submultiple) and a suitable unit of 

time. In this report, absorbed dose rates are generally given in 

mrad/yr. 

Linear Energy Transfer (LET) 

The linear energy transfer, L,, is a quantity that 

represents the energy lost, by collision, per unit length by 

charged particles in an absorbing medium. It represents the 

increment of the mean energy lost, AE, to tissue by a charged 

particle of specified energy in traversing a distance, AX: 

L, = dE/dX = lim (A EE/ X) (keV pm-l) (5-8) Ax4 

For photons, L-represents the energy imparted by the 

secondary electrons (electrons that are knocked out of their 

orbitals by primary radiation) resulting from secondary 

interactions between the photons and tissue material. High-LET 
radiation (alpha particles) imparts more energy per unit length 

of organ tissue than does low-LET radiation (x-rays, gamma rays, 

and beta particles). Consequently, the former are more effective 

per unit dose in causing bioLogica1 damage. 

Dose Equivalent and Dose Equivalent Rate 

Dose equivalent is a special radiation protection quantity 

that is used to express the absorbed dose in a manner that 

considers the difference in biological effectiveness of various 

kinds of ionizing radiation. The ICRU has defined the dose 

equivalent, HI as the product of the absorbed dose, D, the 

quality factor, Q, and all other modifying factors, N, at the 

point of interest in biological tissue (ICRUBO). This 

relationship can be expressed in the following manner: 

The quality factor is a dimensionless quantity that depends 

on the collision stopping power for charged particles, and it 

accounts for the differences in biological effectiveness found 

among varying types of radiation. By definition, it is 

independent of tissue and biological endpoint. The generally 

accepted values for quality factors for high- and low-LET 

radiation, which are used by EPA, are given in Table 5-1. The 

product of all other modifying factors, N, such as dose rate, 

fractionation, etc., is taken as 1. 

The dose equivalent rate, H, is the time rate of change of 

the dose equivalent to organs and tissues and is expressed as: 

H = dH/dt = lim (A H/A t) . (mrem/yr) (5-10) 

AI+O 

Table 5-1. Quality factor for various types of radiation 

(ICRP77). 

Radiation Type Quality Factors (Q) 

x-rays, gamma rays, and electrons 1 

alpha particles 2 0 
Effective Dose Equivalent and Effective Dose Equivalent Rate 

The ICRP has defined the effective dose equivalent, HE, as: 

where H, is the dose equivalent in tissue and w, is the weighting 

factor, which represents the estimated proportion of the 

stochastic risk resulting from tissue, T, to the stochastic risk 

when the whole body is uniformly irradiated (ICRP77). The 

weighting factors recommended by the ICRP are listed in 

Table 5-2. 

The effective dose equivalent rate is the time derivative of 

the dose equivalent and is expressed as %, where: 

Relationship of the Dose Equivalent and the Effective Dose 

Equivalent to Risk 

The dose equivalent was introduced by the ICRP to allow one 

to combine and compare - on a risk basis - absorbed doses of 

different types of radiation. Subsequently, the effective dose 

equivalent was introduced to provide a single-valued indicator of 

risk for dose equivalents distributed nonuniformly in the body. 

By convention, these concepts, in combination with the ICRPrecommended quality factors and organ-weighting factors, are 

widely used in radiation protection. These recommended factors, 

however, are based on dose response models that differ 

significantly from those used by EPA to estimate risk (see 

Chapter 6) . 

To calculate risk, EPA first calculates age-specific, 

high- and low-LET absorbed dose rates, by organ, for a uniform 

intake or external exposure rate. The risk from each year's dose 

is then calculated using a life table procedure in conjunction 

with age- and organ-specific risk models adapted from the BEIR 

I11 report (NAS80). 

These models (see Chapter 6) assume a linear dose-response 

relationship and a lifetime relative risk projection for cancers 
Table 5-2. Weighting factors recommended by the ICRP 

for stochastic risks (ZCRP77). 

Organ or Tissue WT 

Gonads 

Breast 

Red Bone Marrow 

Lung 

Thyroid 

Bone Surfaces 

Remainder 

other than bone cancer and leukemia, for which absolute risk 

projection is employed. Finally, the risks from each year's dose 

are summed to arrive at the risk from lifetime exposure. 

In calculating dose equivalents and effective dose 

equivalents, the ICRP Publication 30 convention was employed, 

including the same quality factors and organ-weighting factors. 

Nevertheless, in calculating the - risk from a given absorbed dose 

of alpha particle irradiation, RBEs of 8 and 2.7 were used for 

the induction of cancers and genetic effects, respectively, (see 

Chapter 6). Since these RBEs are lower than the assumed alpha 

quality factor (Q=20), the estimated risks per unit dose 

equivalent (mrem) will generally be lower for alpha particles 

than for either x-rays or gamma rays. Likewise, the ICRP organweighting factors shown in Table 5-2 do not stand in the same 

proportion as the organ risks calculated using the EPA models for 

cancer induction or genetic mutations. Furthermore, EPA 

considers somatic and genetic risks separately. Thus, even if 

attention was restricted to low-LET radiation, the estimated risk 

from a given effective dose equivalent will vary, depending on 

how the absorbed dose is distributed within the body. 

To summarize, because EPA risk models differ from those 

underlying the ICRP recommendations, the risks calculated 

directly by EPA are not strictly proportional to the effective 

dose equivalents derived using ICRP quality factors and organ 

weighting factors. 
Working Levels and Working Level Months 

The working level is a unit that has been used as a measure the 

radon decay-product activity in air. It is defined as any 

combination of short-lived radon daughters (through Po-214) per 

liter of air that will result in the ultimate emission of 

1.3 x lo5 MeV of alpha energy. An activity concentration of 

100 pCi/l of Rn-222 in equilibrium with its short-lived daughters 

gives rise to a potential alpha-energy concentration of 

approximately 1 WL. The WL unit could also be used for thoron 

daughters. The potential alpha energy exposure is commonly 

expressed in units of working level months (WLM). One WLM 

corresponds to an exposure to a concentration of 1 WL for the 

commonly used reference period of 170 hours. 

Customary and SI Units 

The relationship between the customary units used in this text 

and the international system of units (SI) for radiological 

quantities is shown in Table 5-3. While the SI radiological 

units are almost universally used in other countries for radiation 

protection regulation, the United States has not yet officially 

adopted their use for such purposes. 

5.3 EPA DOSIMETRIC MODELS 

The EPA dosimetric models, to be discussed in the following 

sections, have been described in detail in previous publications 

(Du80, Su81). Information on the elements treated in these sections 

was taken directly from those documents or reports. 

With the permission of the authors, many edited passages have been 

adopted to explain EPArs dosimetric models. In most cases, the EPA 

models are similar or identical to those recommended by the ICRP 

(ICRP79, ICRP80, ICRP81). However, differences in model parameters 

do exist for some radionuclides (Su81). The basic physiological 

and metabolic data used by EPA in calculating radiation doses are 

taken from ICRP reports (ICRP75, ICRP79) . 
Table 5-3. Comparison of customary a.nd SI special units for radiation 

quantities. 

Quantity Customary Unit Definition SI Unit Definition 

Activity (A) Curie (Ci) 3.7~10'~ s-' becquerel (Bq) s-' 

Absorbed Dose (D) rad (rad) J kg-1 gray (Gy) J kg" 

Absorbed . rad per second J kg-' s-I gray per J kg-' s-' i 

Dose Rate (D) (rad s-' ) second 

(GY s-') 

Dose rem (rem) 

Equivalent (H) 

J kg-' sievert (Sv) J kg-' 

Dose Equivalent rem per second J kg-' s-' sievert per J kg-' s-' 

- Rate (H) (rem s-') second 

(Sv s-') 

Linear Energy kiloelectron 1.602~10-'~ ~m-' 

Transfer (L,) volts per 

micrometer 

(keV m-') 
5.3.1 Internal Dose Models 

EPA implements contemporary models to estimate absorbed dose 

rates as a function of time to specified organs in the body. 

Estimates of the doses resulting from the deposition and retention 

of inhaled particulates in the lung and their subsequent absorption 

into the blood and clearance into the gastrointestinal (GI) tract 

are made using the ICRP Task Group Lung Model (ICRP66). 

5.3.1.1 Generalized Scheme for Estimating Organ Absorbed Dose 

Rates 

Distribution of Activity of Radionuclides in the Body 

The complex behavior of radionuclides is simplified 

conceptually by considering the body as a set of compartments. A 

compartment may be any anatomical, physiological, or physical 

subdivision of the body throughout which the concentration of a 

radionuclide is assumed to be uniform at any given time. The terms 

"compartment" and "organ" are often used interchangeably, although 

some of the compartments considered in this report may represent 

only portions of a structure usually considered to be an organ, 

while some compartments may represent portions of the body usually 

not associated with organs. Examples of compartments used in this 

report are the stomach, the pulmonary lung, the blood, or the bone. 

Within a compartment, there may be more than one "pool" of activity. 

A pool is defined to be any fraction of the activity within a 

compartment that has a biological half-life which is distinguishable 

from the half-time(s) oE the remainder of activity within the 

compartment. 

Activity entering the body by ingestion is assumed to originate 

in the stomach compartment; activity entering through inhalation is 

assumed to originate in a compartment within the lung (the tracheobronchial, pulmonary, or naso-pharyngeal region). From the stomach, 

the activity is viewed as passing in series through the small 

intestine, the upper large intestine, and the lower large intestine, 

from which it may be excreted. Also, activity reaching the small 

intestine may be absorbed through the wall into the bloodstream, 

from which it may be taken in parallel into any of several 

compartments within the skeleton, liver, kidney, thyroid, and other 

organs and tissues. 

The list of organs or regions for which dose rates are 

calculated is found in Table 5-4. Activity in the lung may reach 

the bloodstream either directly or indirectly through the stomach or 

lymphatic system. The respiratory system and gastrointestinal tract 

models are discussed further in later sections. Figure 5-1 
Table 5-4. Target organs and tissues used for calculating the 

ICRP effective dose equivalent and the EPA cancer 

risk. 

ICRP effective 

dose equivalent 

EPA cancer risk 

Ovaries 

Testes 

Muscle" 

Red marrow 

~ungs~ 

Thyroid 

Bone surface 

Stomach wall 

Small intestine wall 

Upper large intestine wall 

Lower large intestine wall 

Kidneys 

Liver 

Pancreas 

Brain 

Breast 

Red Marrow 

Pulmonary lung 

Thyroid 

Bone surface (endosteum) 

Stomach wall 

Intestinec 

Kidneys 

Liver 

pancreasd 

Spleen 

Thymus 

Uterus 

Adrenals 

Bladder wall 

a) Dose to breast is assumed to equal dose to muscle. 

b) The ICRP estimate is a mass-weighted mean of the values 

for the tracheo-bronchial region, the pulmonary region, 

and the respiratory lymph nodes. 

C) The EPA averages the values for the small, upper large, 

and lower large intestine using weights of 0.2, 0.4, and 

0.4 respectively for calculating the risk of bowel cancer. 

d) The pancreas is also used as a surrogate organ for 

calculating the cancer risk for all other organs and tissues. 
i.ll.ustrates the EPA modei used to represent the movement of 

radioactivity in the body. 

EPA models separately consider the intake and subsequent behavior 

of each radionuclide in the body. The models also allow for the 

formation of radioactive decay products within the body, and it is 

assumed that the movement of internally produced radioactive 

daughters is governed by their own metabolic properties rather than 

those of the parent. This is in contrast to the ICRP assumption 

that daughters behave exactly as the parent. 

If A,,(t) denotes the activity of the ith species of the chain in 

organ k and if that activity is divided among several "pools" or 

"compartments" indexed by subscript 1, then the time rate of change 

of activity can be modeled by a system of differential equations of 

the following form: 

where compartment 1 is assumed to have L, separate pools of 

activity, and where: 

Ailk = the activity of species i in compartment 1 of organ k; 

h p = (ln 2) T:, where T: = radioactive half of species i; 

h p, = rate coefficient (time.') for biological removal of 

species i from compartment 1 of organ k; 

Lik = number of exponential terms in the retention function for 

species i in organ k; 

Bij = branching ratio of nuclide j to species i; 

PI, = inflow rate of the ith species onto the organ k; and 

ci, = the fractional coefficient for nuclide i in the lth 

compartment of organ k. 

The subsystem described by these L,, equations can be 

interpreted as a biological compartment in which the fractional 

retention of radioactive species is governed by exponential 
Figure 5-1. A schematic representation of radioactivity 

movement among respiratory tract, gastrointestinal 

tract, and blood. 

S = stomach 

SI = small intestine 

ULI = upper large intestine 

LLI = lower large intestine 

h = elimination rate constant 
decay. Radioactivity that enters an organ may be lost by both 

radioactive decay and biological removal processes. For each 

source organ, the fraction of the initial activity remaining at 

any time after uptake at time t = 0 is described by a retention 

function consisting of one or more exponentially decaying terms: 

The subscript 1 in the above equation represents the lth 

term of the retention function, and the coefficients c,, can be 

considered as "pathway fractions." Figure 5-2 illustrates an 

example of the decline of activity in an organ as a function of 

time t. 

Dose Rates to Target Organs 

The activity of a radionuclide in a compartment is a measure 

of the rate of energy being emitted in that compartment, at any 

time, t, and can be related to the dose rate to a specific organ 

at that time. This requires estimating the fraction of the 

energy emitted by the decay of the radionuclide in each 

compartment that is absorbed by the specific organ. 

The absorbed dose rate, D,(X;t) to target organ X at time t 

due to radionuclide species i in source organs Y,,Y,, ...., Y, is 

estimated by the following equation: 

where: Dl (X+Y,;t) = S,(X+Y ,) A,,(t) ; and A,,(t) is the 

activity, at time t of species i in source organ Y,. 

S,(XtY,), called the S-factor, represents the average dose 

rate to target organ X from one unit of activity of the 

radionuclide uniformly distributed in source organ or compartment 

Y,. It is expressed in the following manner: 

where : 

C - - a constant that depends on the units of 

dose, energy, and time being used; 
Figure 5-2. Typical pattern of decline of activity of a 

radionuclide in an organ, assuming an initial 

activity in the organ and no additional uptake of 

radionuclide by the organ. 
f, = intensity of decay event (number per 

disintegration); 

Em = average energy of decay everit (Mevj; and 

@,(XcY ,) = specific absorbed fraction, i .e., the 

fraction emitted energy from source organ Y, 

absorbed by target organ X per gram of X, 

where the summation is taken over all events of type m. The 

units for S-factors depend on the units used for activity and 

time; thus, the S-factor units may be rad/Ci-day. The S-factor 

is similar in concept to the SEE factor (specific effective 

energy) used by the ICRP Committee 2 in Publication 30. However, 

the SEE factor includes a quality factor for the type of 

radiation emitted during the transformation. 

The above equations are combined to produce the following 

expressions for the absorbed dose rates to target organs at any 

time due to one unit of activity of radionuclide species, i, 

uniformly distributed in source organs Y, ... Y,: 

The corresponding dose equivalent rate, H,(X;t), can be estimated 

by inclusion of the quality factor, Q,, and the modifying factor, 

N,(Y,) : 

Hi (X;t) = C C Aik(t) Q Si,(X+Yk) (5-19) k m 

Implicit in the above equations is the assumption that the 

absorbed dose rate to an organ is determined by averaging 

absorbed dose distributions over its entire mass. 

Alpha and beta particles are usually not sufficiently 

energetic to contribute a significant cross-irradiation dose to 

targets separate from the source organ. Thus, the absorbed 

fraction for these radiations is generally assumed to be just the 

inverse of the mass of organ X, or if the source and target are 

separated, then Q,(XtY) = 0. Exceptions occur when the source 

and target are in very close proximity, as is the case with 

various skeletal tissues. Absorbed fractions for crossirradiations by beta particles among skeletal tissues were taken 

from ICRP Publication 3 (ICRP80). The energy of alpha particles 

and their associated recoil nuclei is generally assumed to be 
absorbed in the source organ. Therefore, $,(X+X) is taken to be 

the inverse of the organ mass, and @,(X+Y) = 0 if X and Y are 

separated. Special calculations are performed for active marrow 

and endosteal cells in bone, based on the method of Thorne 

(Th77). 

Monte Carlo Methodology to Estimate Photon Doses to Organs 

The Monte Carlo method uses a computerized approach to 

estimate the probability of photons interacting within target 

organ X after emission from source organ Y. The method is 

carried out for all combinations of source and target organs and 

for several photon energies. The body is represented by an 

idealized phantom in which the internal organs are assigned 

masses, shapes, positions, and attenuation coefficients based 

on their chemical composition. A mass attenuation coefficient, 

p,, is chosen, where p, is greater than or equal to the mass 

attenuation coefficients for any region of the body. Photon 

courses are simulated in randomly chosen directions, and 

potential sites of interactions are selected by taking distances 

traversed by them as -In r/p,, where r is a random number 

distributed between 0 and 1. The process is terminated when 

either the total energy of photons has been deposited or the 

photon escapes from the body. The energy deposition for an 

interaction is determined according to standard equations 

(ORNL74). 

Effects of Decay Products 

In calculating doses from internal and external exposures, 

the in-growth of radioactive decay products (or daughters) must 

be considered for some radionuclides. When an atom undergoes 

radioactive decay, the new atom created in the process, which may 

also be radioactive, can contribute to the radiation dose to 

organs or tissues in the body. Although these decay products may 

be treated as independent radionuclides in external exposure, the 

decay products of each parent must be followed through the body 

in internal exposure situations. The decay product contributions 

to the absorbed dose rates, which are included in EPA 

calculations, are based on the metabolic properties of the 

individual daughters and the organ in which they occur. 

5.3.1.2 Inhalation Dosimetry - ICRP Respiratory Tract Model 

As stated earlier, individuals immersed in contaminated air 

will breathe radioactive aerosols or particulates, which can lead 

to doses to the lung and other organs in the body. The total 

internal dose caused by inhalation of these aerosols can depend 
on a varie'ty of factors, such as breathing rates, particle sizes, 

and physical activity. Estimating the total dose to individuals 

over a specific time period requires specifying the distribution 

of particle depositions in the respiratory tract and the 

mathematical characteristics of the clearance parameters. The 

EPA currently uses assumptions established by the ICRP Task Group 

on Lung Dynamics (TGLM) (ICW66). This section will summarize the 

essential features of that model. For a more comprehensive 

treatment, the reader is referred to the actual report. 

The basic features of the ICRP lung compartmental model are 

shown in Figure 5-3. According to this model, the respiratory 

tract is divided into four regions: naso-pharyngeal (N-P), 

tracheo-bronchial (T-B), pulmonary (P), and lymphatic tissues. 

- In the model, the regions N-P, T-B, and P are assumed to 

receive fractions D,, D,, and D, of the inhaled particulates, 

where the sum of these is less than 1 (some particles are removed 

by prompt exhalation). The values D,, D,, and D, depend on the 

activity median aerodynamic diameter (AMAD) of the inspired 

particles. For purposes of risk calculations, EPA uses AMADs of 

1 micron. The lung model employs three clearance classes, D, W, 

and Y, corresponding to rapid, intermediate, and low clearance, 

respectively, of material deposited in the respiratory passages. 

The clearance class depends on chemical properties of the inhaled 

particles. 

Like the ICW, EPA assumes that the absorbed dose rate to 

the N-P region can be neglected. Unlike the ICW, however, EPA 

averages the dose over the pulmonary region of the lung 

(compartments e through h), to which is assigned a mass of 570 g, 

including capillary blood (ICRP75). In addition, it is assumed 

that the total volume of air breathed in one day by a typical 

member of the general population is 22,000 liters. This value 

was determined by averaging the 23 ICW adult male and female 

values based on 8 hours of working "light activity," 8 hours of 

nonoccupational activity, and 8 hours of resting. 

5.3.1.3 Ingestion Dosimetry - ICW GI Tract Model 

According to the ICRP 30 GI tract model, the 

gastrointestinal tract consists of four compartments: the stomach 

(S), small intestine (SI), upper large intestine (ULI), and lower 

large intestine (LLI). The fundamental features of the model are 

shown in Figure 5-1. It is assumed that absorption into the 

blood occurs only from the small intestine (SI). 
This model postulates that radioactive material entering the 

compartments of the GI tract is exponentially removed by both 

radioactive decay and biological removal processes, and that 

there is no feedback. Absorption of a particular nuclide from 

the GI tract is characterized by f,, which represents that 

fraction of the nuclide ingested which is absorbed inta body 

fluids if no radiological decay occurs: 

where 

h,": = the absorption coefficient (s-') 

h,, = the transfer coefficient fron the small intestine 

to the large intestine (s-l) 

Figure 5-1 graphically presents the role of these coefficients in 

the gastrointestinal model. The kinetic model, as formulated by 

the ICRP, does not permit total absorption of a nuclide (f, = 1). 

5.3.1.4 Dose Rate Coversion Factors 

EPA uses the computer code RADRISK (Du80) for calculating 

radiation doses and risks to individuals resulting from a unit 

intake of a radionuclide, at a constant rate, for a lifetime 

exposure (50-yr dose commitment). These calculations are done 

for the inhalation and ingestion pathways to individuals who are 

exposed by immersion in contaminated air or by contaminated 

ground surfaces. 

RADRISK computes doses for both chronic and acute exposures. 

Following an acute intake, it is assumed the activity in the body 

decreases monotonically, particularly for radionuclides with 

rapid radiological decay rates or rapid biological clearance. In 

the case of chronic exposure, the activity in each organ of the 

body increases monotonically until a steady state is achieved, at 

which time the activity remains constant. The instantaneous dose 

rates at various times after the start of chronic exposure 

provide a reasonably accurate (and conservative) estimate of the 

total annual dose for chronic exposure conditions. However, the 

instantaneous dose rates may err substantially in the estimation 

of annual dose from an acute exposure, particularly if the 

activity levels decrease rapidly. 

Since the rate of change in activity levels in various 

organs is more rapid at early times after exposure, doses are 

computed annually for the first several years and for 

progressively longer periods thereafter, dividing by the length 
COMPARTMENT 

N-P a 

(D3 = 0.30) b 

Figure 5-3. The ICRP Task Group lung Model for particulates. 

The colmns labeled D, W, and Y correspond, respectively, to rapid, intewediate, and slaw 

clearnnce of the inspired material (in days, weeks, or years). The symbols T and P denote the 

biological half-tirne (dayel and coefficient, respectively, of a term in the appropriate retention 

function. The values shown for Dj, Dq, and Dg correspond to activity median aerodpamic diameter 

AHAD 1 pn and represent the fraction of the inspired material depositing in the lung regione, 
of the interval to estimate the average annual dose. This method 

produces estimates of risk that are similar to those computed by 

the original RADRISK methodology for chronic exposures and 

provides a more accurate estimate of the risks from acute 

intakes. 

5.3.1.5 Special Radionuclides 

The following paragraphs briefly summarize some of the 

special considerations for particular elements and radionuclides. 

Tritium and Carbon-14 

Most radionuclides are nuclides of elements found only in 

trace quantities in the body. Others like tritium (hydrogen-3) 

or C-14 must be treated differently since they are long-lived 

nuclides of elements that are ubiquitous in tissue. An intake of 

tritium is assumed to be completely absorbed and to be rapidly 

mixed with the water content of the body (Ki78a). 

The estimates for inhalation include consideration of 

absorption through the skin. Organ dose estimates are based on 

the steady-state specific-activity model described by Killough 

et al. (Ki78a) . 

C-14 is assumed to be inhaled as CO, or ingested in a 

biologically bound form. Inhaled C-14 is assumed to be diluted 

by stable carbon from ingestion (Ki78b). This approach allows 

separate consideration of the ingestion and inhalation pathways. 

The specific-activity model used for organ dose estimates is also 

that of Killough et al. (Ki78a). Short-lived carbon 

radionuclides (e.g., C-11 or C-15) are treated as trace elements, 

and the organ doses are calculated accordingly. 

Noble Gases 

Retention of noble gases in the lungs is treated according 

to the approach described by Dunning et al. (Du79). The inhaled 

gas is assumed to remain in the lungs until lost by radiological 

decay or respiratory exchange. Translocation of the noble gas to 

systemic organs is not considered, but doses due to translocated 

decay products produced in the lungs are calculated. The 

inhalation of the short-lived decay products of radon is assessed 

using a potential alpha energy exposure model (see Chapter 6) 

rather than by calculating the doses to lung tissues from these 

radionuclides. 
For some essential elements, such as potassium and chemically 

similar radioelements, such as rubidium and cesium, absorption 

into the bloodstream is nearly complete at all ages, so that 

changes with age and possible homeostatic adaptations in 

absorption are not discernible. The fraction of a radioelement 

that is transferred to the blood depends on its chemical form, 

and wide ranges of values are found in the literature for 

individuals who ingest the material under different conditions. 

For example, f, values for uranium were found to range from 0.005 

to 0.05 for industrial workers, but a higher average value of 0.2 

(0.12 to 0.31) is indicated by dietary data from persons not 

ocoupationally exposed (ICFS79). EPA has used the 0.2 value for 

uranium ingestion by the general population. 

It appears that all iodine entering the small intestine is 

absorbed into the blood; hence the f, value is taken as 1 for all 

ages, which is the value used in this analysis. 

Organ Masses, m 

To a large extent, the variability in organ masses among 

individuals in the general population is related to age. For 

most of the target organs listed in Table 5-2, the mass increases 

during childhood and continues to increase until adulthood, at 

which time the net growth of the organ ceases; there may be a 

gradual decrease in mass (for some organs) in later years. 

Based on data reviewed by Dunning and Schwarz (Du81), the 

mass of an adult thyroid ranges from 2 to 62 g. It is expected 

that this parameter variability would be reflected in large 

dosimetric variability among adults. Children in the age group 

from .5 to 2 yr were found to have a mean thyroid mass of 2.1 g, 

while the adult group had a mean mass of 18.3 g. For this 

illustration, the same values are used as employed by the ICFS 

(20 g for the adult thyroid mass and 1.8 g for that of a 

6-month-old child), which are also consistent with the 

recommendation of Bryant (Br69) . 

The fraction of a radionuclide taken up from the blood in an 

organ is strongly correlated with the size of the organ, its 

metabolic activity, and the amount of material ingested. Iodine 

introduced into the bloodstream is rapidly deposited in the 

thyroid, usually reaching a peak slightly after 24 hours. The 

uptake of 1-131 by the thyroid is similar to that of stable 

iodine in the diet and can be influenced by sex and dietary 

differences. There can be considerable variation among 

populations. 
Dunning and Schwarz (Du81) found a mean f,'value of 0.47 

for newborns, 0.39 for infants, 0.47 for adolescents, and 0.19 

for adults. This analysis uses f; values of .35 and .15 

for a child and adult, respectively. 

Effective Half-Life, TI,, 

Some data suggest a strong correlation between biological 

half-lives of radionuclides in organs in the body and the age of 

the individual. Children are expected to exhibit faster 

elimination rates and greater uptakes (Ro58). For iodine, a 

range of biological half-lives of 21 to 200 days for adults has 

been observed, and a similarly wide range would be expected for 

other age groups (Dual). Rosenberg (Ro58) found a significant 

correlation between the biological half-life and the age of the 

individual and an inverse relationship between uptake and age in 

subjects from 22 to 50 yr of age. Dunning and Schwarz (Dual) 

concluded that for adults the observed range was from 21 to 372 

days; for children in the age group from .5 to 2 yr, the range 

was 4 to 39 days. 

In light of the possible inverse relation between the 

biological half-life and the f; value, this analysis uses 

biological half-lives of 24 and 129 days, respectively, for 

children and adults, based on the paper by Bryant (Br69). 

Including the effect of radioactive decay, these values imply an 

effective half-life of 6 days in adults and 8 days in children. 

Effective Energy per Disintegration, E 

The effective energy per disintegration (MeV/dis) of a 

radionuclide within an organ depends on the decay energy of the 

radionuclide and the effective radius of the organ containing the 

radionuclide (ICRP59). It is expected, therefore, that E is an 

age-dependent parameter which could vary as the size of the organ 

changes. While very little work has been done in determining E 

for most radionuclides, some information has been published for 

1-131 and Cs-137. Considering the differences between the child 

and the adult thyroid, Bryant (Br69) estimates E to be 0.18 

MeV/dis for the child and 0.19 MeV/dis for the adult. The above 

values correspond to a 6-month-old child with a mass of 1.8 g and 

an f: value of 0.35. The corresponding E value for the adult was 

calculated for a 20-9 thyroid with an f; value of 0.3. 

Taking into account all the age-dependent factors discussed 

above, this analysis indicates that, for a given concentration of 

1-131 in milk, the estimated dose rate to the thyroid of a 

6-month-old child would be approximately 13 times that to an 
adult thyroid. In other words, use of adult parameters would 

underestimate the thyroid dose to the child by about a factor 

of 13. 

5.3.1.6.8 Significance of Parameter Variabililty to EPA Dose 

and Risk Assessments 

In its radiological risk assessments, EPA is generally 

interested in estimating the risk to an average individual due to 

chronic lifetime exposures. Variation in dosimetric parameters 

between people and between age groups is of reduced importance in 

this context because such variation gets averaged over a 

population and/or over a lifetime. Nevertheless, it should be 

kept in mind that some individuals in a population are going to 

be at higher risk from a given exposure. Furthermore, despite 

such averaging, parameter variability can contribute 

substantially to the uncertainty in the dose and risk estimates. 

Parameter variation among individuals contributes uncertainty 

to the models by causing random errors in any measured human data 

upon which the dosimetric models are based. To the extent that 

the subjects from whom such data are collected are atypical of 

the U.S. population (e.g., with respect to health status), 

parameter variation may also be a source of bias. In this 

respect, since the parameters contained in the dosimetric models 

were estimated for adult males, primarily, they may not provide 

an adequate basis for calculating the average dose or risk in 

cases where age- and sex-related variations in these parameters 

are large. This problem becomes more significant in light of the 

generally higher risks associated with a given dose for childhood 

exposures (see Chapter 6); if doses are also higher in childhood, 

the enhanced effect on risk will be compounded. 

5.3.1.6.9 Past Approaches Used in Estimating Uncertainties 

in Calculated Organ Dose 

As in any predictive exercise, it is useful to question the 

reliability of the predictions. Variations in environmental 

levels, dietary and life style preferences, and the variability 

of controlling physiological and metabolic processes contribute 

to the distribution of dose among members of the exposed 

population. Superimposed on this variability is a component of 

uncertainty arising from limitations in the predictive ability of 

the dosimetric models themselves. Various approaches have been 

taken to understand and quantify these uncertainties. 

It has recently become popular to estimate the uncertainty by 

computing the distribution of dose among exposed individuals. 
This approach consists of repeated solution of the dosimetric 

model using parameter values selected at random from a frequency 

distribution of potential values suggested in the literature. It 

is assumed that the dosimetric model has been properly 

formulated, although these models were developed to yield point 

estimates. Despite these and other difficulties, propagation of 

parameter uncertainty through the dosimetric equation can provide 

a measure of the model uncertainty. Application of these methods 

to the estimation of dose from 1-131 and Cs-137 ingestion can be 

found in the literature (Du81, Sc82). 

An alternative approach to assessing the potential 

variability is to consider that the observed frequency 

distribution of a measurable quantity is closely related to dose. 

Cuddihy and co-workers (Cu79) have investigated the variability 

of selected target organ deposition among test animals and some 

individuals exposed. However, they did not address differences 

in age, gender, magnitude or duration of exposure. 

5.3.1.6.10 Uncertainty Classification of Radionuclides 

In this section, radionuclides of interest are classified in 

terms of the uncertainties in estimated dose per unit intake. 

Nuclides are placed in broad groups, largely reflecting the 

general status of information on their biokinetic behavior in the 

body. It is assumed that the uncertainty associated with the 

calculation of the energy deposition in the target tissues is a 

minor contributor to the overall uncertainty. 

Classification of Uncertainty in Radionuclide Dose 

Establishing numerical values of uncertainty for model dose 

estimates of each of the many radionuclides, for each route of 

exposure, is a formidable task. Even if there is agreement on 

the definition of uncertainty, any quantification will be 

arbitrary to a degree. No model has been verified in man for any 

long-term exposure scenario; some of the models may be 

fundamentally wrong in their formulation. In addition, the data 

selected to establish the parameters used in the model may not be 

representative of the population being evaluated. Most risk 

assessors use some informed scientific judgment in estimating the 

level of uncertainty in a dose model. 

A broad categorization of radionuclides reflecting the 

estimated magnitude of the dosimetric uncertainties is presented. 

Because of the problems cited above with respect to the 

development of models and model parameters, it is quite possible 

that the error in model estimates may be larger than indicated in 

some cases. Nevertheless, this exercise is useful since it 
provides some perspective on the magnitude of the uncertainties 

in light of current evidence and focuses attention on the largest 

gaps in knowledge. Ultimately, however, better quantification of 

dose estimates and their associated uncertainties can be obtained 

only through the development and verification of improved 

dosimetric models. 

Radioisotopes behave biologically like their stable elements. 

The elements, in turn, can be broadly grouped as: (1) essential 

elements and their analogs, (2) inert gases, (3) well-studied 

toxic metals and (4) others. Uncertainties for each of these 

categories will be expressed as multiplicative factors, which 

roughly estimate the 95% upper and lower confidence interval 

limits. [Since the interval is based on judgment, a preferable 

term would be "credibility interval" (NIH85).] 

Group I - Essential Elements and Their Analogs 

Essential elements are controlled by homeostatic mechanisms 

to within narrow tolerances. Usually, analogs of essential 

elements have distribution and deposition patterns similar to 

those of the essential element. The uncertainty expected in 

calculated dose for essential elements is a factor of two or less 

in major critical organs, perhaps 3 or less in other significant 

tissues and organs. The expected dose uncertainty for analogs of 

essential elements is perhaps a little greater, a factor of 3 or 

less in major organs and up to 5 or more in less significant 

tissues. Important radionuclides of essential elements include 

H-3, C-14, P-32, K-40, Ca-45, Co-60, 1-129, and 1-131; important 

analogs include Sr-89, Sr-90, Cs-134, Cs-137, Ra-226, and Ra-228. 

Group I1 - Inert Gases 

Uptake and retention of inhaled inert gases has been fairly 

well studied. The uncertainty in dose, particularly average 

whole body dose, is not expected to be large. However, the gases 

do not distribute uniformly in body tissues, and the effect of 

distribution on organ dose estimates has not been carefully 

addressed. The uncertainty in the calculated dose is expected to 

be about a factor of 2. This group includes, but is not limited 

to Ar-41, Kr-85, Xe-133, and Rn-222. 

Group 111 - Well-Studied Toxic Metals 

A number of elements have been extensively studied in animals 

with limited information available for man. Examples here 

include toxic elements encountered in industrial activities, 

e.g., mercury, cadmium, lead, and uranium, for which studies were 

carried out to help establish safe working conditions. Often the 
available information is not sufficiently complete to identify 

the dominant processes governing the biokinetic behavior or is 

simply fragmentary. For example, while much information exists 

on the biokinetics of uranium, considerable uncertainty remains 

associated with the absorption to blood from the small intestine. 

Uncertainties for dose estimates in this group of elements would 

be variable, ranging from 2 or less for lead up to about 5 or 

more for polonium, thorium, uranium, and the transuranics. 

Nuclides in this group include, but are not limited to Pb-210, 

Po-210, U-235, U-238, Th-230, Th-232, Pu-239, Pu-241, and Am-241. 

Group IV - Other Elements 

For a number of radionuclides information is largely limited 

to data from animal studies. While animal studies often are the 

major source of detailed information on the processes governing 

the biokinetics, the lack of a general framework for 

extrapolations to man and the limited information upon which to 

judge the reasonableness of the extrapolations suggest that the 

estimates must be considered to be potentially in error by at 

least an order of magnitude. Nuclides in this group include, but 

are not limited to Ce-144 and other rare earth elements, Tc-99, 

Cm-244, Cf-252, etc. 

The groupings listed above represent the Agency's best 

judgment on the uncertainty of internal radionuclide dose 

estimates. The primary source of uncertainty is in the 

biokinetic modeling with little uncertainty in the physics. 

The magnitudes of the uncertainties posited for each group of 

radionuclides should be regarded as only rough estimates; 

however, the qualitative breakdown between groups is fairly 

reliable. 

Specific Problems 

Certain radioisotopes and aspects of dosimetry pose unique 

problems. While the effect of these problems may be to increase 

the uncertainty in dose estimates, the extent of such an increase 

has yet to be evaluated. 

Long-Lived Bone Seekers 

Radioisotopes with effective half-lives that are short 

compared to the average life span are expected to be in dynamic 

equilibrium. However, some bone seekers have long effective 

half-lives; therefore, they do not reach dynamic equilibrium 

during a life span. Since the relevant human biokinetic data are 
qui.te iimi.ted, dose estimates for such radionuclides are more 

uncertain. 

Nonuniformity of Distribution 

The distribution of an element within an organ may not be 

uniform; in particular, the distribution may be nonuniform with 

respect to biological targets of interest. This can be a serious 

problem with respect to the estimation of relevant doses from 

internally deposited alpha emitters, given the short range of 

alpha particles in matter. For example, where an alpha emitter 

is distributed nonuniformly in bone, the calculation of doses to 

sensitive cells in the bone and the bone marrow will be 

difficult. Another example is the uncertainty in estimating 

doses to cells lining the GI tract from ingested alpha emitters 

passing through the tract. In some cases, the mucus lining may 

effectively shield the target cells from irradiation. 

5.3.2 External Dose Models 

This section is concerned with the calculation of dose rates 

for external exposure to photons from radionuclides dispersed in 

the environment. Two exposure models are discussed: (1) 

immersion in contaminated air and (2) irradiation from material 

deposited on the ground surface. The immersion source is 

considered to be a uniform semi-infinite radionuclide 

concentration in air, while the ground surface irradiation source 

is viewed as a uniform radionuclide concentration on an infinite 

plane. In both exposure modes, the dose rates to organs are 

calculated from the dose rate in air. 

Dose rates are calculated as the product of a dose rate 

factor, which is specific for each radionuclide, tissue, and 

exposure mode, and the corresponding air or surface 

concentration. The dose rate factors used were calculated with 

the DOSE FACTOR code (Ko8la,b). Note that the dose rate factors 

for each radionuclide do not include any contribution for decay 

products. For example, the ground surface dose factors for 

cesium-137 are all zero, since no photons are emitted in its 

decay. To assess surface deposition of Cs-137, the ingrowth of 

its decay product, metastable Ba-137, which is a photon emitter, 

must first be calculated. 

5.3.2.1 Immersion 

For immersion exposure to the photons from radionuclides in 

air, EPA assumes that an individual is standing at the base of a 

semi-infinite cloud of uniform radionuclide concentration. 

First, the dose rate factor (the dose rate for a unit 
concentration) in air is calculated for a source of photons with 

energy E,. At all points in an infinite uniform source, 

conservation of energy considerations require that the rates of 

absorbed and emitted energy per unit mass be equal. The absorbed 

energy rate per unit mass at the boundary of a semi-infinite 

cloud is just half that value. Hence 

where : 

Dm; = the immersion dose rate per unit air concentration 

(rad m3/Ci s); 

E, = emitted photon energy (MeV); 

k = units conversion factor 

= 1.623-13 (J/MeV) x 3.7E+10 (dis/s-Ci) x 1.OE+3 (g/kg) 

x 100 (rad kg/J) 

= 5.93Ei2 (g rad/MeV Ci s); and 

P = density of air (g/m3). 

The above equation presumes that for each nuclide 

transformation, one photon with energy E, is emitted. The dose 

rate factor for a nuclide is obtained by adding together the 

contributions from each photon associated with the transformation 

process for that radionuclide. 

5.3.2.2 Ground Surface Irradiation 

In the case of air immersion, the radiation field was the 

same throughout the source region. This allows the dose rate 

factor to be calculated on the basis of energy conservation 

without having to consider explicitly the scattering processes 

taking place. For ground surface irradiation, the radiation 

field depends on the height of the receptor above the surface, 

and the dose rate factor calculat-ion is more complicated. The 

radiation flux per unit solid angle is strongly dependent on the 

angle of incidence. It increases from the value for photons 

incident from immediately below the receptor to a maximum close 

to the horizon. Attenuation and buildup due to scattering must 

be considered to calculate the dose rate factor. Secondary 

scattering provides a distribution of photon energies at the 

receptor, which increases the radiation flux above that 

calculated on the basis of attenuation. Trubey (Tr66) has 

provided a useful and reasonably accurate expression to 

approximate this buildup: 
but more precise estimates of these distributions are not likely 

to change the organ dose rate factors substantially. 

Kocher (Ko81) has noted that the idealized photon dose rate 

factors are "likely to be used quite extensively even for 

exposure conditions for which they are not strictly applicable ... because more realistic estimates are considerably more difficult 

and expensive [to make] ." 
EPA77 U.S. Environmental Protection Agency, Proposed Guidance on 

Dose Limits for Persons Exposed to Transuranium Elements 

in the General Environment, EPA 520/4-77-016, 1977. 

Fi35 Findeisen, W., Uber das Absetzen Kleiner in der Luft 

Suspendierten Teilchen in der Menschlichen Lunge bei der 

Atmung, Pflugers Arch. f d ges. Physiol., 236, 367, 1935. 

FRC67 Federal Radiation Council, Guidance for the Control of 

Radiation Hazards in Uranium Mining, FRC Report No. 8, 

Revised, U.S. Government Printing Office, Washington, 

D.C., 1967. 

H075 Holden, W.S., and Marshal, R., "Variations in 

Bronchial Movement", Clin. Radiol., - 26:439-454, 1975. 

Hu72 Hughes, J.M.B., Hoppin, F.G., Jr. and Mead, J., "Effect of 

Lung Inflation on Bronchial Length and Diameter in Excised 

Lungs", J. Appl. Physiol., - 32:25-35, 1972. 

Hu73 Hursh, J.B., and Spoor, N.L., "Data on Man", Chapter 4 in 

Uranium, Plutonium and the Transplutonic Elements, 

Springer, New York, 1973. 

ICRP59International Commission on Radiological Protection, 

Report of Committee I1 on Permissible Dose for Internal 

Radiation, ICRP Publication 2, Pergamon Press, Oxford, 

1959. 

ICRP66ICRP Task Group on Lung Dynamics, "Depositions and 

Retention Models for Internal Dosimetry of the Human 

Respiratory Tract", Health Phys., g(2) : 173-207, 1966. 

ICRP75International Commission on Radiological Protection, 

Report on the Task Group on Reference Man, ICRP 

Publication No. 23, Pergamon Press, Oxford, 1975. 

ICRP77International Commission on Radiological Protection, 

"Recommendations of the International Commission on 

Radiological Protection", ICRP Publication 26, Annals of 

the ICRP, Vol. 1, No. 3, Pergamon Press, Oxford, 1977. 
ICRP79International Commission on Radiological Protection, 

Limits for Intakes of Radionuclides by Workers, LCRP 

Publication No. 30, Pergamon Press, Oxford, 1979. 

ICRP80International Commission on Radiological Protection, 

"Limits for Intakes of Radionuclides by Workers", ICRP 

Publication 30, Part 2, Annals of the ICRP, Vol. 4 (3/4), 

Pergamon Press, Oxford, 1980. 

ICRP81International Commission on Radiological Protection, 

"Limits for Intakes of Radionuclides by Workers", ICRP 

Publication 30, Part 3, Annals of the ICRP, Vol. 6 (2/3), 

Pergamon Press, Oxford, 1981. 

ICRP84International Commission on Radiological Protection, "A 

Compilation of the Major Concepts and Quantities in Use by 

ICRP", ICRF Publication No. 42, Pergamon Press, Oxford, 

1984. 

ICRU80International Commission on Radiation Units and 

Measurements, ICRU Report No 33, Washington, D.C., 1980 

Ke80 Kerr, G.D., and Eckerman, K.F., Oak Ridge National 

Laboratory, private communication; see also Abstract P/192 

presented at the Annual Meeting of the Health Physics 

Society, Seattle, Washington, July 20-25, 1980. 

Ke80a Kerr., G.D., "A Review of Organ Doses from Isotropic 

Fields of X-Rays", Health Phys., - 39 (1) :3, 1980. 

Ki78a Killough, G.C., Dunning, D.E Jr., Bernard, S.R. and 

Pleasant, J.C., Estimates of Internal Dose Equivalent to 

22 Target Organs for Radionuclides Occurring in Routine 

Releases from Nuclear Fuel Cycle Facilities, Vol. 1, 

Report No. ORNL/NUREG/TM-190, Oak Ridge National 

Laboratory, Tennessee, June 1978. 

Ki78b Killough, G.C., and Rohwer, P. S., "A New Look at the 

Dosimetry of 14c Released to the Atmosphere as Carbon 

Dioxide", Health Phys., z(2) :141, 1978. 
Ko81a Kocher, D.C., and Eckerrnan, K.F., "Electron Dose-Rate 

Conversion Factors for External Exposure of the Skin", 

Health Phys., - 40 (1) : 67, 1981. 

Ko8lb Kocher, D.C., "Dose-Rate Conversion Factors for External 

Exposure to Photon and Electron Radiation from 

Radionuclides Occurring in Routine Releases from Nuclear 

Fuel-Cycle Facilities", Health Phys., - 38(4):543-621, 1981 

NAS80 National Academy of Sciences - National Research Council, 

The Effects on Populations of Exposure to Low Levels of 

Ionizing Radiation, Report of the Committee on the 

Biological Effects of Ionizing Radiation (BEIR 111), 

Washington, D.C., 1980. 

NCRP71National Council on Radiation Protection and Measurements, 

Basic Radiation Protection Criteria, NCRP Report No. 39, 

Washington, D.C., 1971. 

Ne84 Nelson, C.B., and Yang, Y., An Estimation of the Daily 

Average Food Intake by Age and Sex for Use in Assessing 

the Radionuclide Intake of Individuals in the General 

Population, EPA 520/1-84-015, 1984. 

NIH85 National Institutes of Health, Report of the National 

Institutes of Health Ad Hoc Working Group to Develop 

Radioepidemiological Tables, NIH Publication No. 85-2748, 

U.S. Government Printing Office, Washington, DC 20402, p 

92, 1985. 

ORNLElOak Ridge National Laboratory, Estimates of Health Risk 

from Exposure to Radioactive Pollutants, ORNL/RM-7745, Oak 

Ridge, Tenn., 1981. 

ORNL850ak Ridge National Laboratory, "Report of Current Work of 

the Metabolism and Dosimetry Research Group", ORNL/TM9690, Oak Ridge, Tennessee, 1985. 

Ph85 Phalen, R.F., Oldham, M.J., Beaucage, C.B., Crocker, T.T., 

and Mortensen, J.D., Postnatal Enlargement of Human 

Tracheobronchial Airways and Implications for Particle 

Deposition, Anat. Rec. 212: 368, 1985. 
Ro58 Rosenberg, E., '%iologic Half-life of 1311 in the Thyroid of 

Healthy Males", J. Clin. Endocrinol. Metab., E, 516-521, 

1958. 

Sc82 Schwarz, G., and Dunning, Jr., D.E., Imprecision in 

Estimates of Dose from Ingested Cs-137 due to Variability 

in Human Biological Characteristics, Health Phys. -8 43 

631-645, 1982. 

Sn74 Snyder W.S., Ford, M.R., Warner, G.G., and Watson, S.B., A 

Tabulation of Dose Equivalent per Microcurie-Day for 

Source and Target Organs of an Adult for Various 

Radionuclides, Oak Ridge National Laboratory, ORNL-5000, 

1974. 

. Sp73 Spoor, N.L., and Hursh, J.B., "Protection Criteria", Chapter 

5 in Uranium, Plutonium and the Transplutonic Elements, 

Springer, New York, 1973. 

Su81 Sullivan, R.E., Nelson, N.S., Ellett, W.H., Dunning, D.E. 

Jr., Leggett, R.W., Yalcintas, M.G. and Eckerman, K.F., 

Estimates of Health Risk from Exposure to Radioactive 

Pollutants, Report No. ORNL/TM-7745, Oak Ridge National 

Laboratory, Oak Ridge, Tennessee, 1981. 

Th77 Thorne, M.D., "Aspects of the Dosimetry of Alpha-Emitting 

Radionuclides in Bone with Particular Emphasis on '"~a and 

239~~~~ , Phys. Med. Biol., 22:36-46, 1977. 

- 

Th78 Thurlbeck, W.M. "Miscellany", 287-315 in The Lung: Structure 

Function and Disease, Thurlbeck, W.M. and Abell, M.R., 

editors, The Williams and Wilkins Co., Baltimore, Maryland, 

1978. 

Tr66 Trubey, D.K., A Survey of Empirical Functions Used to Fit 

Gamma-Ray Buildup Factors, Oak Ridge National Laboratory 

Rep., ORNL-RSIC-10, 1966. 
6. ESTIMATING THE RISK OF HEALTH EFFECTS RESULTING FROM 

EXPOSURE TO LOW LEVELS OF IONIZING RADIATION 

6.1 INTRODUCTION 

This chapter describes how EPA estimates the risk of fatal 

cancer, serious genetic effects, and other detrimental health 

effects caused by exposure to low levels of ionizing radiation. 

Ionizing radiation refers to radiation that strips electrons 

from atoms in a medium through which it passes. The highly reactive 

electrons and ions created by this process in a living cell can 

produce, through a series of chemical reactions, permanent changes 

(mutations) in the cell's genetic material, the DNA. These may 

result in cell death or in an abnormally functioning cell. A 

mutation in a germ cell (sperm or ovum) may be transmitted to an 

offspring and be expressed as a genetic defect in that offspring or 

in an individual of a subsequent generation; such a defect is 

commonly referred to as a genetic effect. There is also strong 

evidence that the induction of a mutation by ionizing radiation in a 

non-germ (somatic) cell can serve as a step in the development of a 

cancer. Finally, mutational or other events, including possible 

cell killing, produced by ionizing radiation in rapidly growing and 

differentiating tissues of an embryo or fetus can give rise to birth 

defects; these are referred to as teratological effects. At acute 

doses above about 25 rads, radiation induces other deleterious 

effects in man; however, for the low doses and dose rates of 

interest i n t h i s document, only those three kinds of effects 

referred to above are thought to be significant. 

Most important from the standpoint of the total societal risk 

from exposures to low-level ionizing radiation are the risks of 

cancer and genetic mutations. Consistent with our current 

understanding of their origins in terms of DNA damage, these are 

believed to be stochastic effects; i.e., the probability (risk) of 

these effects increases with the absorbed dose of radiation, but the 

severity of the effects is independent of dose. For neither 

induction of cancer nor genetic effects, moreover, is there any 

convincing evidence for a "threshold," i.e., some dose level below 

which the risk is zero. Hence, so far as is known, any dose of 

ionizing radiation, no matter how small, might give rise to a cancer 

or to a genetic effect in future generations. Conversely, there is 

no way to be certain that a given dose of radiation, no matter how 

large, has caused an observed cancer in an individual or will cause 

one in the future. 

Beginning nearly with the discovery of x-rays in 1895 but 

especially since World War 11, an enormous amount of research has 

been conducted into the biological effects of ionizing radiation. 

This research continues at the level of the molecule, the cell, the 

tissue, the whole laboratory animal, and man. There are two 

fundamental aspects to most of this work: 
1. Estimating the radiation dose to a target (cell, tissue, 

etc.) . This aspect (dosimetry), which may involve 

consideration of physiological, metabolic, and other 

factors, is discussed more fully in Chapter 5. 

2. Measuring the number of effects of a given type 

associated with a certain dose (or exposure). 

For the purpose of assessing the risk to man from exposures to 

ionizing radiation, the most important information comes from human 

epidemiological studies in which the number of health effects 

observed in an irradiated population is compared to that in an 

unirradiated control population. The human epidemiological data 

regarding radiation-induced cancer are extensive. As a result, the 

risk can be estimated to within an order of magnitude with a high 

degree of confidence. Perhaps for only one other carcinogen - 

tobacco smoke - is it possible to estimate risks more reliably. 

Nevertheless, there are gaps in the human data on radiation 

risks. No clear-cut evidence of excess genetic effects has been 

found in irradiated human populations, for example, probably due to 

the limited numbers in the exposed cohort providing inadequate power 

to detect a dose-response. Likewise, no statistically significant 

excess of cancers has been demonstrated below about 5 rads, the dose - 

range of interest from the standpoint of environmental exposures. 

Since the epidemiological data are incomplete in many respects, risk 

assessors must rely on mathematical models to estimate the risk from 

exposures to low-level ionizing radiation. The choice of models, of 

necessity, involves subjective judgments but should be based on all 

relevant sources of data collected by both laboratory scientists and 

epidemiologists. Thus, radiation risk assessment is a process that 

continues to evolve as new scientific information becomes available. 

The EPA estimates of cancer and genetic risks used here are 

based largely on the results of a National Academy of Sciences (NAS) 

study as given in the BEIR I11 report (NAS80). The study assessed 

radiation risks at low exposure levels. As phrased by the President 

of the Academy, "We believe that the report will be helpful to the 

EPA and other agencies as they reassess radiation protection 

standards. It provides the scientific bases upon which standards 

may be decided after nonscientific social values have been taken 

into account." 

In this discussion, the various assumptions made in calculating 

radiation risks based on the 1980 NAS report are outlined, and these 

risk estimates are compared with those prepared by other scientific 

groups, such as the 1972 NAS BEIR Committee (NAS72), the United 

Nations Scientific Committee on the Effects of Atomic Radiation - (UNSCEAR77, 82), and the ICRP (ICRP77). Because information on 

radiation risks is incomplete, estimates of risk based on the 

various models may not be highly accurate. This discussion 

identifies some of the deficiencies in the available data base and 

points out possible sources of bias in current risk estimates. 
Nevertheless, the risk estimates made by EPA are believed to be 

reasonable in light of current evidence. 

Sections 6.2 to 6.2.8 consider the cancer risk resulting from 

whole-body exposure to low-LET (see Chapter 5) radiation, i.e., 

sparsely ionizing radiation like the energetic electrons produced by 

x-rays or gamma rays. Environmental contamination by radioactive 

materials also leads to the ingestion or inhalation of the material 

and subsequent concentration of the radioactivity in selected body 

organs. Therefore, the cancer risk resulting from low-LET 

irradiation of specific organs is examined in Sections 6.2.9 to 

6.2.11. 

Organ doses can also result from high-LET radiation, such as 

that associated with alpha particles. The cancer risks when highLET radiation is distributed more or less uniformly within a body 

organ is the third situation considered (Section 6.3). Because 

densely ionizing alpha particles have a very short range in tissue, 

there are exposure situations where the dose distribution to 

particular organs is extremely nonuniform. An example is the case 

of inhaled radon progeny, Po-218, Pb-214, and Po-214. For these 

radionuclides, cancer risk estimates are based on the amount of 

radon progeny inhaled rather than the estimated dose, which is 

highly nonuniform and cannot be well quantified. Therefore, risk 

estimates of radon exposure are examined separately (Section 6.4). 

In Section 6.5, the causes of uncertainty in the cancer risk 

estimates and the magnitude of this uncertainty are reviewed so that 

both the public and EPA decision makers have a proper understanding 

of the degree of confidence to place in them. In Section 6.6 

reviews and quantifies the risk of deleterious genetic effects from 

radiation and the effects of exposure in utero on the developing 

fetus. Finally, in Section 6.7, cancer and genetic risks from 

background radiation are calculated using the models described in 

this chapter. 

6.2 CANCER RISK ESTIMATES FOR LOW-LET RADIATION 

Most of the observations of radiation-induced carcinogenesis in 

humans are of groups exposed to low-LET radiations. These groups 

include the Japanese A-bomb survivors and medical patients treated 

with diagnostic or therapeutic radiation, most notably for 

ankylosing spondylitis in England from 1935 to 1954 (Sm78). 

Comprehensive reviews of these and other data on the carcinogenic 

effects of human exposures are available (UNSCEAR77, NAS80). 

The most important source of epidemiological data on radiogenic 

cancer is the population of Japanese A-bomb survivors. The A-bomb 

survivors have been studied for more than 38 years, and most of them 

(the Life Span Study Sample) have been followed since 1950 in a 

carefully planned and monitored epidemiological survey (Ka82, Wa83). 

They are the largest group that has been studied, and they provide 

the most detailed information on the response pattern for organs by 
age and sex over a wide range of doses of low-LET radiation. 

Unfortunately, the 1980 BEIR Committee's analysis of the A-bomb 

survivor data collected up to 1974 was prepared before bias in the 

dose estimates for the survivors (the tentative 1965 dose estimates, 

T65) became widely recognized (Lo81). It is now clear that the T65 

dose equivalents to organs tended, on average, to be overestimated 

(Bo82, RERF83,84) so that the BEIR Committee's estimates of the risk 

per unit dose are likely to be too low. A new dosimetry system, 

termed the Dosimetry System 1986 (DS86), is now nearly complete, and 

preliminary analyses of the risk based on DS86 have been published 

(Pr87,88; Sh87) . 

At present, the "BEIR V Committee" of the National Academy of 

Sciences is preparing a report on radiation risks in light of DS86 

and other new information. A detailed reevaluation of EPA's current 

risk estimates is indicated when this report is issued, probably in 

early 1989. A brief discussion of the new dosimetry and its likely 

effect on risk estimates is included in Section 6.5.2. 

Uncertainties in radiation risk estimates do not result just 

from the uncertainties about the Japanese and other epidemiological 

studies. As discussed below, risk projections based on these 

studies require certain assumptions (e.g., with regard to low-dose 

extrapolation). The degree of uncertainty associated with these 

assumptions is probably greater than the uncertainty of the 

estimated risk per unit dose among the A-bomb survivors or other 

sources of risk estimates for radiogenic cancer in humans. 

6.2.1 Assumptions Needed to Make Risk Estimates 

A number of assumptions must be made on how to extrapolate 

observations made at high doses to estimate effects from low doses 

and low-dose rates. Excess cancers have been observed, for the most 

part, only following doses of ionizing radiation that are relatively 

high when compared to those likely to occur as a result of the 

combination of background radiation and environmental contamination 

from controllable sources of radiation. Therefore, a dose response 

model must be chosen to allow extrapolation from the number of 

radiogenic cancers observed at high doses to the number of cancers 

at low doses resulting from all causes including background 

radiation. 

The range of extrapolation is not the same for all kinds of 

cancer because it depends upon the radiosensitivity of a particular 

tissue. For example, the most probable radiogenic cancer for women 

is breast cancer. As described below, the incidence of radiogenic 

breast cancer does not seem to diminish when the dose is protracted 

over a long period of time. For example, the number of excess 

cancers per unit dose among Japanese women, who received acute 

doses, is about the same per unit dose as women exposed to small 

periodic doses of x-rays over many years. If this is actually the 

case, background radiation is as carcinogenic per unit dose for 

breast tissue as the acute exposures from A-bomb gamma radiation. 
Moreover, the female A-bomb survivors show an excess of breast 

cancer at doses below 20 rads which is linearly proportional to that 

observed at several hundred rads (To84). (Evidence of a nonlinear 

dose response relationship for induction of breast cancer has been 

obtained in a study of Canadian fluoroscopy patients, but only at 

doses above about 500 rads [Hoe41 .) Women in their 401s, the 

youngest age group in which breast cancer is common, have received 

about 4 rads of whole-body low-LET background radiation and usually 

some additional dose incurred for diagnostic medical purposes. 

Therefore, for this cancer, the difference between the lowest dose 

at which radiogenic cancers are observed, less than 20 rads, and the 

dose resulting from background radiation is less than a factor of 5, 

not several orders of magnitude as is sometimes claimed. Based on 

data from irradiated tinea capitis patients, induction of thyroid 

cancer also seems to be linear with doses down to 10 rads or lower 

(NCRP85). However, for most other cancers, a statistically 

significant excess has not been observed at doses below 50 rads of 

low-LET radiation. Therefore, the range of dose and dose rate 

extrapolation is often large. 

6.2.2 Dose Response Functions 

The 1980 NAS report (NAS80) examined only three dose response 

functions in detail: (1) linear, in which the number of effects 

(risk) is directly proportional to dose at all doses; (2) linearquadratic, in which risk is very nearly proportional to dose at very 

low doses and proportional to the square of the dose at high doses; 

and (3) quadratic, where the risk varies as the square of the dose 

at all dose levels. 

The first two of these functions are compatible with most of 

the data on human cancer. Information that became available only 

after the BEIR I11 report was published indicates that a quadratic 

response function is inconsistent with the observed excess risk of 

solid cancers at Nagasaki, where the estimated gamma-ray doses are 

not seriously confounded by an assumed neutron dose component. The 

chance that a quadratic response function underlies the excess 

cancer observed in the Nagasaki incidence data has been reported as 

only 1 in 10,000 (Wa83) . 

Although a quadratic response function is not incompatible with 

the Life Span Study Sample data on leukemia incidence at Nagasaki, 

Beebe and others (Be78, E177) have pointed out how unrepresentative 

these data are of the total observed dose response for leukemia in 

that city. There is no evidence that a quadratic response function 

provides a better fit to the observed leukemia excess among all Abomb survivors in the Life Span Study Sample than a simple linear 

model (NAS80). Based on these considerations, for the purposes of 

this rulemaking, a quadratic response cannot be used to estimate 

cancer risks due to ionizing radiation. 
The 1980 NAS HEIR Committee considered only the Japanese 

mortality data in its analysis of possible dose response functions 

(NAS80). Based on the T65 dose estimates, this Committee concluded 

that the excess mortality from solid cancers and leukemia among the 

A-bomb survivors is compatible with either a linear or linearquadratic dose response to the low-LET radiation component and a 

linear response to the high-LET neutron component (NAS80). Although 

the 1980 BEIR report indicated risk estimates for low-LET radiation 

based on a linear-quadratic response were "preferred" by most of the 

scientists who prepared that report, opinion was not unanimous, and 

the subsequent reassessment of the A-bomb dose seriously weakens the 

Committee's conclusion. The Committee's analysis of dose response 

functions was based on the assumption that most of the observed 

excess leukemia and solid cancers among survivors in Hiroshima 

resulted from neutrons (see Tables V-13, A-7, Equations V-10, V-11 

in NAS80). Current evidence, however, is conclusive that neutrons 

were only a minor component of the dose among all but a few 

survivors in both Hiroshima and Nagasaki (Bo82; RERF83, 84; Pr87; 

Sh87). Therefore, it is likely that most of the response attributed 

to neutrons was caused by the gamma dose, not the dose from 

neutrons. This point is discussed further in Section 6.3. 

The revised dosimetry will involve changes in individual 

absorbed doses that vary with distance from the explosion in each of 

the two cities and with shielding characteristics. As a 

consequence, though it seems clear that there will generally be a 

higher response per unit dose, there will also be an unpredictable 

change in the shape of the dose response exhibited by the data. 

Reanalysis of the Japanese experience after completion of the dose 

reassessment may then provide more definitive information on the 

dose response of the A-bomb survivors; nevertheless, it is unlikely 

to produce a consensus on the dose response at environmental levels, 

i.e., about 100 mrad/yr. This is because at low enough doses, there 

will always be sampling variations in the observed risks so that 

observations are compatible, in a statistical sense, with a variety 

of dose response functions. In the absence of empirical evidence or 

a strong theoretical basis, a choice between dose response functions 

must be based on other considerations. 

Although there is evidence for a nonlinear response to low-LET 

radiations in some, but not all, studies of animal 

radiocarcinogenesis (see below), no data on human cancers that are 

incompatible with the linear model are known. In such a case, it 

may be preferable to adopt the simplest hypothesis that adequately 

models the observed radiation effect. Moreover, EPA believes that 

risk estimates, for the purpose of assessing radiation impacts on 

public health, should be based on scientifically credible risk 

models that are unlikely to understate the risk. The linear model 

fulfills this criterion. Given the current bias in the doses 

assigned to A-bomb survivors (see Section 6.5.1), such an approach 

seems reasonable as well as prudent. Therefore, EPA has primarily 

used the BEIR I11 linear dose response model for estimating the risk 

of radiogenic cancer due to low-LET radiations. 
For low-LET radiations, the BEIR 111 Committee preferred the 

linear-quadratic dose response model. In this model, the risk from 

an acute dose, D, of low-LET radiation is assumed to be of the form 

aD + PD~. The BEIR 111 Committee assumed that the linear and 

quadratic terms were equal at 116 rads, leading to a linear 

coefficient a, which was a factor of 2.5 times lower than the 

coefficient obtained from the linear model (NAS80). At low doses, 

the quadratic term becomes negligible; at chronic low-dose rates it 

is ignored, for reasons discussed below. For environmental 

exposures, therefore, risk estimates based on the BEIR 111 linearquadratic dose response model are only 40 percent of those based on 

the BEIR 111 linear model. 

A theoretical basis for the linear-quadratic dose response 

model has been put forth by Kellerer and Rossi (Ke72). In this 

theory of "dual radiation action," events leading to "lesions" 

(i.e., permanent changes) in cellular DNA require the formation of 

interacting pairs of "sublesions." The interacting pairs can be 

produced by a single traversing particle, or track, or by two 

separate tracks, giving rise, respectively, to a linear and 

quadratic term in the dose response relationship. According to the 

theory, a sublesion may be repaired before it can interact to form a 

lesion, the probability of such repair increasing with time. 

Consequently, as dose rate is reduced, the formation of lesions from 

sublesions caused by separate tracks becomes less important, and the 

magnitude of the D' term diminishes. Hence, the theory predicts 

that at sufficiently low doses or dose rates, the response should be 

a linear function of dose. Moreover, the constant of 

proportionality is the same in both cases; i.e., a. 

Results of many animal experiments are qualitatively consistent 

with the theory: low-LET radiation often seems to have a reduced 

effectiveness per unit dose at low dose rates (NCRP80). However, it 

is usually not possible from the data to verify that the dose 

response curve has the linear-quadratic form. Another success of 

the dual action theory has been in explaining observed differences 

between the effects of low-LET and high-LET radiations. In this 

view, the densely ionizing nature of the latter results in a much 

greater production of interacting pairs of sublesions by single 

tracks, leading in turn to higher relative biological effectiveness 

at low doses and a linear dose response relationship for high-LET 

radiation (except for possible cell-killing effects). 

The dual action theory has nevertheless been challenged on 

experimental grounds, and observed variations in response with dose, 

dose rate (see below), and LET can also be explained in terms of a 

theory involving only single lesions and a "saturable" repair 

mechanism that decreases in effectiveness at high dose rates on the 

microscopic scale (Go82). One property of such a theory is that the 

effectiveness of repair, and therefore the shape of the dose 

response curve, can in principle vary substantially with cell type 
1980 NAS report. Table 6-1, taken from Table V-25 (NAS80), shows 

the range of cancer fatalities that are induced by a single lo-rad 

dose as estimated using linear, linear-quadratic, and quadratic dose 

response functions and two projection models, relative and absolute 

risk (NAS80) . 

As illustrated in Table 6-1, estimating the cancer risk for a 

given projection model on the basis of a quadratic as compared to a 

linear dose response reduces the estimated risk of fatal cancer by a 

factor of about 18. Between the more credible linear and linearquadratic response functions, the difference is less, a factor of 

about 2.2. For a given dose response model, results obtained with 

the two projection models for solid cancers differ by about a factor 

of 3. 

Differences in the estimated cancer risk introduced by the 

choice of the risk projection model are also appreciable. As 

pointed out above, the 1980 NAS analysis indicates that relative 

lifetime risk estimates exceed absolute risk estimates by about a 

factor of 3 (see Table 6-1). However, relative risk estimates are 

quite sensitive to how the risk resulting from exposure during 

childhood persists throughout life. This question is addressed in 

the next section, where risk estimates made by the 1972 and 1980 NAS 

BEIR Committees are compared with those of the ICRP and UNSCEAR. 

6.2.6 Comparison of Cancer Risk Estimates for Low-LET Radiation 

Estimates of the risk of fatal cancer following lifetime 

exposure are compared in Table 6-2. The BEIR I and BEIR I11 values 

were calculated for this table using risk model data from NAS72 and 

NAS80. The BEIR I11 values in this table differ slightly from those 

in NAS80 and Table 6-1 because of some minor calculational 

corrections including revised age-specific mortality data. 

Differences between the various estimates in Table 6-1 mainly 

reflect differences in assumptions regarding extrapolation to low 

doses and low dose rates and regarding the dependence of risk on age 

and time. 

The NAS BEIR estimates are for lifetime exposure and lifetime 

expression of induced cancers (NAS72, NAS80). Neither the age 

distribution of the population at risk nor the projection models (if 

any) have been specified by the UNSCEAR (UNSCEAR77) or the ICRP 

(ICRP77). UNSCEAR apparently presumes the same age distributions as 

in the epidemiological studies it cited, mainly the A-bomb 

survivors, and a 40-year period of cancer expression. The ICRP risk 

estimates are for adult workers, presumably exposed between ages 18 

and 65, and a similar expression period. These are essentially ageindependent absolute risk models with less than lifetime expression 
Table 6-1. Comparative estimate of cancer fatalities induced by 

a single 10-rad, low-LET radiation exposure to the 

general population (average value per rad per million 

persons exposed) . 

Dose Response Projection Modelb 

Model" 

Relative Absolute 

- 

L-L, L-L 501 

- 

LQ-L, LQ-L 226 

- 

Q-L, Q-L 2 8 

a The first model is used for leukemia and bone 

cancer, the second for other forms. 

b Leukemia and bone cancer fatalities are projected 

by means of the absolute risk model in both 

columns (NAS80). 

Source: NAS80, Table V-25. 
increase in solid cancers that has occurred among irradiated 

populations (Xa82). 

To project the number of fatalities resulting from leukemia and 

bone cancer, EPA uses an absolute risk model, a minimum induction 

period of 2 years, and a 25-year expression period. To estimate the 

number of fatalities resulting from other cancers, EPA has used a 

relative risk projection model (EPA84), a 10-year minimum induction 

period, and the remaining balance of an exposed person's lifetime as 

the expression period. 

6.2.8 Methodology for Assessing the Risk of Radiogenic Cancer 

EPA uses a life table analysis to estimate the number of fatal 

radiogenic cancers in an exposed population of 100,000 persons. 

This analysis considers not only death due to radiogenic cancer, but 

also the probabilities of other competing causes of death which are, 

of course, much larger and vary considerably with age (Bu81, Co78). 

Basically, it calculates for ages 0 to 110 the risk of death due to 

all causes by applying the 1970 mortality data from the National 

Center for Health Statistics (NCHS75) to a cohort of 100,000 

persons. Additional details of the life table analysis are provided 

in Appendix B. It should be noted that a life table analysis is 

required to use the age-dependent risk coefficients in the BEIR 111 

report. For relative risk estimates, EPA has used age-specific 

cancer mortality data also provided by NCHS (NCHS73). The EPA 

computer program used for the life table analysis was furnished to 

the NAS BEIR 111 Committee by EPA and used by the Committee to 

prepare its risk estimates. Therefore, the population base and 

calculations should be essentially the same in both the NAS and EPA 

analyses. 

Both absolute and relative risk models have been considered to 

project the observed risks of most solid radiogenic cancers beyond 

the period of current observation. As indicated in Table 6-2, the 

range of estimated fatal cancers resulting from the choice of a 

particular projection model and its internal assumptions is about a 

factor of 3. Although the relative risk model has been tested in 

some detail only for lung and breast cancer (La78), based on current 

evidence, it appears to be the better projection model for solid 

cancers. Therefore, it has been adopted for risk estimates in this 

report. Previously, EPA used an average of the risks calculated by 

the absolute and relative risk projection models (EPA84). 

To estimate the cancer risk from low-LET, whole-body, lifetime 

wosure, the analysis uses relative risk projections (the BEIR I11 

L-L model) for solid cancers and the absolute risk projection for 

leukemia and bone cancer (the BEIR TI1 L-L model). Since the 

expression period for leukemia and bone cancer is less than the 

follow-up period, the same risk values would be calculated for these 
cancers using either projection method. For a dose to the whole 

body, this procedure yields about 400 fatalities per million personrad (for the BEIR 111 linear-quadratic model, a low-LET whole-body 

dose would yield an estimated lifetime risk of about 160 fatalities 

per million person-rad). 

BEIR I11 also presented estimates of excess soft tissue cancer 

incidence risk coefficients for specific sites, as a function of age 

at exposure, in its Table V-14. By summing the site-specific risks, 

it then arrived at an estimate for the whole-body risk of cancer 

incidence (other than leukemia and bone cancer) as given in Table V30. Finally, by using the weighted incidence/mortality ratios given 

in Table V-15 of the same report (NAS80), the results in Table V-30 

can be expressed in terms of mortality to yield (for lifetime 

exposure) a risk estimate of about 242 and 776 cancer fatalities per 

10' person-rad, depending on whether an absolute or a relative risk 

projection model, respectively, is used to estimate lifetime risk. 

These values - are about 1.7 and 2.1 times their counterparts for the 

BEIR I11 L-L model and 4.2 and 5.2 times the LQ-L values. These 

models all presume a uniform dose to all tissues at risk in the 

body. In practice, such uniform whole-body exposures seldom occur, 

particularly for ingested or inhaled radioactivity. The next 

section describes how this risk estimate is apportioned for wholebody exposure when considering the risks following the exposure of 

specific organs. 

6.2.9 Organ Risks 

For most sources of environmental contamination, inhalation and 

ingestion of radioactivity are more common than external exposure. 

In many cases, depending on the chemical and physical 

characteristics of the radioactive material, inhalation and 

ingestion result in a nonuniform distribution of radioactive 

materials within the body so that some organ systems receive much 

higher doses than others. For example, since iodine isotopes 

concentrate preferentially in the thyroid gland, the dose to this 

organ can be orders of magnitude larger than the average dose to the 

body. 

To determine the probability that fatal cancer occurs at a 

particular site, EPA has performed life table analyses for each 

cancer type using the information on cancer incidence and mortality 

in NAS80. NAS80 published incidence risk coefficients (NAS80 Table 

V-14) and mortality to incidence ratios (NAS80 Table V-15). The 

data in Tables 6-3 and 6-4 are from these tables with the exception 

of the mortality to incidence ratios for thyroid and lung cancer. 

Since not all forms of thyroid cancer can be induced by radiation 

and since, for those that are, a more reasonable mortality to 

incidence ratio would be 0.1 (NRC85), EPA has used that value in its 

calculations. Lung cancer incidence and mortality have both shown 
an increasing trend between 1970 and 1980. Since incidence leads 

mortality, an uncorrected mortality to incidence ratio gives a low 

estimate of the fraction of those persons who, having been diagnosed 

with lung cancer, will die of that disease. Therefore, a mortality 

to incidence ratio of 0.94, based on long-term survival studies by 

the National Cancer Institute for lung cancer (J. Horn, private 

communication), has been used. 

Risk coefficients for a site-specific relative risk model were 

calculated as follows: 

1. Mortality risk coefficients for an absolute risk model were 

calculated using the data in Tables 6-3 and 6-4. 

2. Following the procedure used in NAS80, absolute risks at an 

absorbed dose rate of 1 mrad/y were calculated for each site for 

males and females in each age group. A 10-year minimum latency and 

a 20-year plateau - i.e., a 30-year follow up - was used for these 

calculations. 

3. The relative risk coefficients (l/rad) for each age group 

providing the same 30-year projected risk were then calculated. 

Following the NAS80 convention, the values calculated for ages 10-19 

were used for ages 0-9. For consistency, this report uses this 

convention for all cancers including lung and breast, for which the - NAS80 absolute risk coefficients are zero in the first decade. For 

calculating thyroid risks, the relevant age-specific mortality rate 

was considered to be one-tenth of the corresponding incidence rate. 

4. Male and female risks for lifetime expression of risk at 1 

mrad/y were then calculated and combined to obtain estimates for the 

general population. 

EPA used the NCHS 1970 life table and mortality data for all 

these calculations. Male and female cohort results were combined 

presuming a ma1e:female sex ratio at birth of 1.0511, consistent 

with the expected lifetimes at birth for the 1970 male, female, and 

general cohort life tables. 

The average risk for a uniform dose to all tissues was 

calculated to be 542 x 806 x and 678 x per rad for 

males, females, and the general population, respectively. 

It is generally accepted that the risk estimates for the 

individual sites are less certain than are the risk estimates for 

all sites combined. Table 6-5 summarizes the relative risk 
Table 6-3. Site-specific incidence risk coefficients 

per rad-y) . 

Age at Exposure 

Site 0-9 10-19 20-34 35-50 50+ 

Males 

Thyroid 

Breast 

Lung 

Esophagus 

Stomach 

Intestine 

Liver 

Pancreas 

Urinary 

Lymphoma 

Other 

All Sites 

Females 

Thyroid 

Breast 

Lung 

Esophagus 

Stomach 

Intestine 

Liver 

Pancreas 

Urinary 

Lymphoma 

Other 

All Sites 

Source: NAS80, Table V-14 
Table 6-4. Site-specific mortality to incidence risk ratios. 

Site Male Female 

Thyroid 

Breast 

Lung 

Esophagus 

Stomach 

Intestine 

Liver 

Pancreas 

Urinary 

Lymphoma 

Other 

Source: NASBO, Table V-15, except thyroid and lung (see text) . 

calculations for the BEIR 111 L-L model. The calculational 

procedure was the same as that outlined above. 

The risks tabulated in Table 6-5 are slightly different from 

those in NAS80. These differences reflect a correction in the 

exposure interval data for each age group and the use of final 

rather than preliminary 1970 mortality data. NAS80 also combined 

male and female risk estimates presuming a sex ratio at birth of 

1:1, which is not consistent with mortality data. 

Since the total risk for all sites is considered more certain 

than the risk for each site individually, the lifetime risks 

calculated for the L-L model have been used as a constraint for the 

sum of the individual site estimates. The relative risk coefficient 

for each site shown in Table 6-6 has been calculated by multiplying 

the coefficient for the unconstrained model for each sex* the 

quotient of the average risk for all age groups for the L-L 

unconstrained site-specific model. The constrained risk 

coefficients are about one-half of the unconstrained values. 

The L-L absolute risk model coefficients for leukemia and bone 

cancer are shown in Table 6-7. The risk coefficient for bone was 

obtained by dividing the value for alpha particles (high-LET) in 

NAS80 Table A-27 by an RBE of 8 to obtain a low-LET value of 1.25 x 

10" per rad-year. The risk coefficients for leukemia were obtained 

by subtracting the risk coefficients for bone from the risk 

coefficients for leukemia and bone from NAS80 Table V-17. EPA has 
followed the BEIR 111 Committeeb practice of using the absolute 

risk model projections for Leukemia and bone cancer with the 

relative risk projection for all other cancers. Since the 

expression period for leukemia and bone cancer is 27 years, there is 

no difference between the number of cancers projected for a 30-year 

and a lifetime follow-up period. 

Table 6-8 shows the average mortality risks per unit absorbed 

dose for the combined leukemia/bone and constrained relative risk 

models. The risk, in general, decreases with increasing age at 

exposure. For a constant, uniform absorbed dose rate to all organs 

and tissues, about 60 percent of the risk is conferred by the 

exposures in the first 20 years of life. 

The mortality to incidence ratios in Table 6-4 were used to 

convert the mortality risk estimates in Table 6-8 to incidence risk 

estimates. For leukemia and bone cancer, the incidence risks are 

considered to be equal as in NAS80. The resultant incidence risks 

are shown in Table 6-9. 

Table 6-10 summarizes the low LET cancer mortality risk per 

unit dose for a uniform, lifetime dose to each site in Table 6-8. 

Since site-specific dose factors for estimating risk of esophageal 

cancer and lymphoma are not available, these risks are added to 

those already in the category "other." The proportional risk column 

shows the fraction of the total risk associated with each site when 

all organs and tissues are uniformly irradiated. 

6.2.10 Thyroid Cancer from Iodine-131 and Iodine-129 

Iodine-131 has been reported to be only one-tenth as effective 

as x-rays or gamma rays in inducing thyroid cancer (NAS72, NCRP77, 

NCRP85). BEIR 111 reported estimates of factors of 10-80 times 

reduction for 1-131 compared to x-rays and noted the estimates were 

derived primarily from animal experiments (NAS80). However, one 

study in rats reported that 1-131 was just as effective as x-rays in 

inducing thyroid cancer, leading an NRC review group to select onethird as the minimum ratio of 1-131 to x-ray effects that is 

compatible with both old and new data (NRC85). 

It would be prudent to use this factor until further 

information from animal studies or some human data are developed. 

In this document, EPA has employed a thyroid cancer risk coefficient 

for internal exposures to 1-131 and 1-129 which is one-third that 

used for gamma rays or beta radiations from other radionuclides. 

6.2.11 Cancer Risks for a Constant Intake Rate 

The fatal cancer risks shown in the tables of this chapter 

presume a lifetime exposure at a constant dose rate. Even for a 
- Table 6-5. BEIR 111 L-L model for excess fatal cancers other 

than leukemia and bone cancer. 

Age at Exposure 

Group 0-9 10-19 20-34 35-49 50+ A1 1 

Risk Coefficients per rad-y) for Absolute Risk Model* 

Male 1.92 1.457 4.327 5.291 8.808 

Female 2.576 1.955 5.807 7.102 11.823 

Risk Coefficients per rad) for Relative Risk Model 

Male 

Female 

General 

Cohort Deaths at lo-' rad/y for Relative Risk Model 

Male 

Female 

General 

Risk per Unit Dose per rad) for Relative Risk Model 

Male 

Female 

General 

Source: NAS80, Table V-20 
Table 6-6. Mortality risk coefficients per rad) for 

the constrained relative risk model. 

Age at Exposure 

Site 0-9 10-19 20-34 35-50 501 

- Male 

Thyroid 

Breast 

Lung 

Esophagus 

Stomach 

Intestine 

Liver 

Pancreas 

Urinary 

Lymphoma 

Other 

Female 

Thyroid 

Breast 

Lung 

Esophagus 

Stomach 

Intestine 

Liver 

Pancreas 

Urinary 

Lymphoma 

Other 

General 

Thyroid 

Breast 

Lung 

Esophagus 

Stomach 

Intestine 

Liver 

Pancreas 

Urinary 

Lymphoma 

Other 
Table 6-7. BEIR 111 L-L model for excess incidence of (and mortality 

from) leukemia and bone cancer (absolute risk model). 

Age at Exposure 

Site 0-9 10-19 20-34 35-50 50+ All 

- Risk Coefficients (lo-$ per rad-y) * 

Male 

Leukemia 3.852 1.724 2.471 1.796 4.194 

Bone 0.125 0.125 0.125 0.125 0.125 

Female 

Leukemia 2.417 1.067 1.541 1.112 2.635 

Bone 0.125 0.125 0.125 0.125 0.125 

. General . 

Leukemia 3.146 1.399 2.005 1.438 3.276 

Bone 0.125 0.125 0.125 0.125 0.125 

Cohort Deaths at 10.' rad/y 

- Male 

Leukemia .0923 .0405 

Bone .0030 ,0029 

Total ,0953 .0435 

Female 

Leukemia .0588 .0257 

Bone .0030 .0030 

Total .0618 .0287 

General 

Leukemia .07 60 .0333 

Bone .0030 .0030 

Total .0790 .0363 

Risk per Unit Dose per rad) 

Male 

Leukemia 94.7 41.9 

Bone 3.1 3.0 

Total 97.8 44.9 

Female 

Leukemia 59.9 26.3 

- Bone 3.1 3.1 

Total 63.0 29.4 
Table 6-7. BEIR I11 L-L model for excess incidence of (and mortality 

from) leukemia and bone cancer (absolute risk model) 

(Continued) . 

Age at Exposure 

Site 0-9 10-19 20-34 20-34 35-49 A1 1 

Risk per Unit Dose per rad) 

General 

Leukemia 77.7 34.3 48.1 31.4 41.2 44.8 

Bone 3.1 3.1 3.0 2.7 1.6 2.5 

Total 80.8 37.3 51.1 34.1 42.8 47.3 

Source: NAS80, Table V-17 (see text) 
Table 6-8. Site-specific mortalit risk per unit dose (%.CE-6 peg rad) for combined leukemia- g one and constrained relative risk 

model. 

Age at Exposure 

Site 0-9 10-19 20-34 35-50 50+ A1 1 

Male 

Leukemia 94.68 41.86 58.46 37.52 48.64 54.19 

Bone 3.07 3.04 2.96 2.61 1.45 2.47 

Thyroid 8.76 8.76 5.08 2.69 0.80 4.32 

Breast 0.00 0.00 0.00 0.00 0.00 0.00 

Lung 145.90 146.95 107.22 61.42 22.55 84.21 

Esophagus 25.57 25.76 6.13 2.82 2.03 9.91 

Stomach 110.95 111.72 40.63 16.40 9.36 46.95 

Intestine 53.49 53.83 20.89 7.60 4.30 22.78 

Liver 168.01 168.24 35.40 9.48 2.50 58.87 

pancreas 74.36 74.90 24.21 10.34 6.55 30.78 

Urinary 40.73 40.99 13.85 5.79 2.22 16.60 

Lymphoma 33.43 33.28 9.62 2.88 0.71 12.49 

Other 37.48 37.23 33.72 13.09 6.93 22.66 

Total 796.43 746.05 358.15 172.65 108.06 366.25 

Female 

Leukemia 

Bone 

Thyroid 

Breast 

Lung 

Esophagus 

Stomach 

Intestine 

Liver 

Pancreas 

Urinary 

Lymphoma 

Other 

Total 

General 

Leukemia 

Bone 

Thyroid 

Breast 

Lung 

Esophagus 

Stomach 

Intestine 

Liver 

Pancreas 

Urinary 

Lymphoma 

Other 

Total 
Table 6-9. Site-specific incidence risk per unit dose (1.OE-6 per rad) for 

combined leukemia-bone and constrained relative risk model. 

Age at Exposure 

Site 0-9 10-19 20-34 35-50 5 0 All 

Male 

Leukemia 94.67 

Bone 3.07 

Thyroid 87.59 

Breast 0.00 

Lung 155.21 

Esophagus 25.57 

Stomach 147.94 

Intestine 102.87 

Liver 168.01 

Pancreas 81.71 

Urinary 110.08 

Lymphoma 45.80 

Other 57.66 

Total 1080.20 

Female 

Leukemia 

Bone 

Thyroid 

Luna 

Liver 

Pancreas 

Urinary 

Lymphoma 

Other 

Total 

General 

Leukemia 

Bone 

Thyroid 

Breast 

Lung 

Esophagus 

Stomach 

Intestine 

Liver 

Pancreas 

Urinary 

Lymphoma 

Other 

Total 
Table 6-10. Site-specific mortality risk per unit dose from 

low-LET radiation, averages for both sexes and all 

ages. 

Cancer EPA89"." EPA84".b UNSCEAR77 ICRP77' 

i. 

Leukemia .I14 .I50 .13 .16 

Bone .006 .009 .02 .04 

Thyroid .016 .099 .07 .04 

Breast .I42 .I30 .16 .20 

Lung .I79 .207 .25 .16 

Stomach .I17 .084 .08 - 

Intestine .059 .039 .12 - 

Liver .I26 .085 .08 - 

Pancreas .088 .059 .02 - 

Urinary .045 .025 .02 - 

Other .I08 .I13 .05 (.40)d 

" Lifetime exposure and cancer expression. UNSCEAR and ICRP 

estimates use different age distributions and periods of expression. 

EPA Radionuclides Background Information Document; EPA 

520/1-84-022-1 (EPA84). 

Normalized for risk of fatal cancer (excludes genetic 

effects) . 

Five additional target organs that have the highest doses 

are assigned 0.08 each for a total of 0.4. 

Estimated proportions for this document. 

- 
dosimetric mode: with age invariant parameters, dose rates vary with 

time for a constant intake rate. This variation reflects the tirnedependent activity levels associated with the retention of the 

radionuclide in the organs and tissues. The ingrowth of radioactive 

decay products can also contribute further to the time-dependence of 

dose rates. 

Traditionally, risk estimates for chronic intake of a 

radionuclide have been determined using a dose commitment modelto 

calculate dose rates following a fixed period (e.g., a 70-year 

average lifespan). For the purpose of estimating risk, these dose 

rates are considered to be invariant over the individual's lifetime. 

This approach is overly conservative for estimating risk for many 

long-lived radionuclides. Therefore, EPA estimates risks for 

constant radionuclide intakes by first determining dose rates to 

each radiosensitive organ or tissue as a function of time. Then 

these dose rates and the xisk models of this chapter are used to 

calculate lifetime risk based on 1970 life table data. The 

resulting risks are consistent with both the dosimetric and risk 

models, and the arbitrary choice of a dose commitment period is 

avoided. 

6.3 FATAL CANCER RISK RESULTING FROM HIGH-LET RADIATION 

This section explains how EPA estimates the risk of fatal 

cancer resulting from exposure to high-LET radiations. Unlike 

exposures to x-rays and gamma rays where the resultant charged 

particle flux results in linear energy transfers (LET) of the order 

of 0.2 to 2 keV per pm in tissue, 5-MeV alpha particles result in 

energy deposition of more than 100 keV per pm. High-LET radiations 

have a larger biological effect per unit dose (rad) than low-LET 

radiations. How much greater depends on the particular biological 

endpoint being considered. For cell killing and other readily 

observed endpoints, the relative biological effectiveness (RBE) of 

high-LET alpha radiations is often 10 or more times greater than 

low-LET radiations. The RBE may also depend on the dose level; for 

example, if linear and linear-quadratic dose response functions are 

appropriate for high- and low-LET irradiations, respectively, then 

the RBE will decrease with increasing dose. 

6.3.1 Quality Factors and RBE for Alpha Particles 

For purposes of calculating dose equivalent, each type of 

biologically important ionizing radiation has been assigned a 

quality factor, Q, to account for its relative efficiency in 

producing biological damage. Unlike an RBE value, which is for a 

specific tissue and well-defined endpoint, a quality factor is based 

on an overall assessment by radiation protection experts of 

potential harm of a given radiation relative to x or gamma 

radiation. In 1977, the ICRP assigned a quality factor of 20 to 

alpha particle irradiation from radionuclides (ICRP77). However, 

the appropriateness of this numerical factor for estimating fatal 
radiogenic cancers is still unclear, particularly for individual 

sites. 

The dose equivalent (in rem) is the dose (in rad) tlmes the 

appropriate quality factor for a specified kind of radiation. For 

the case of internally deposited alpha-particle emitters, the dose 

equivalent from a one-rad dose is 20 rem. It should be noted that 

prior to ICRP Report 26 (ICRP79), the quality factor assigned to 

alpha particle irradiation was 10. That is, the biological effect 

from a given dose of alpha particles was estimated to be 10 times 

that from an acute dose of low-LET x-rays or gamma rays of the same 

magnitude in rad. The ICRP decision to increase this quality factor 

to 20 followed from its decision to estimate the risk of low-LET 

radiations, in occupational situations, on the assumption that 

biological effects were reduced at low dose rates. There is general 

agreement that dose rate effects do not occur for high-LET (alpha) 

radiations. Implicit in ICRP's risk estimates for low dose/dose - rate gamma radiation is a dose rate reduction factor of about 2.5. 

The EPA (linear) risk model for low-LET radiation does not involve - such a DREF; therefore, in order to avoid an artifactual inflation - in high-LET risk estimates, EPA has assumed an RBE of 8 (20/2.5) for 

calculating the risks from alpha particles (see Section 6.3.3). 

. In 1980, the ICRP published the task group report "Biological 

Effects of Inhaled Radionuclides," which compared the results of 

animal experiments on radiocarcinogenesis following the inhalation 

of alpha-particle and beta-particle emitters (ICRP80). The task 

group concluded that: "...the experimental animal data tend to 

support the decision by the ICRP to change the recommended quality 

factor from 10 to 20 for alpha radiation." 

6.3.2 Dose Response Function 

In the case of high-LET radiation, a linear dose response is 

commonly observed in both human and animal studies. This response 

is not reduced at low dose rates (NCRPBO). Some data on human lung 

cancer indicate that the carcinogenic response per unit dose of 

alpha radiation is maximal at low doses (Ar81, Ho81, Wh83); in 

addition, some studies with animals show the same response (Ch81, 

Ul82). EPA agrees with the NAS BEIR 111 Committee that: "For highLET radiation, such as from internally deposited alpha-emitting 

radionuclides, the linear hypothesis is less likely to lead to 

overestimates of the risk and may, in fact, lead to underestimates" 

(NAS80). However, at low doses, departures from linearity are small 

compared to the uncertainty in the human epidemiological data, and 

EPA believes a linear response provides an adequate model for 

evaluating risks in the general environment. 

A possible exception to a linear response is provided by the 

- data for bone sarcoma (but not sinus carcinoma) among U.S. dial 

painters who ingested alpha-emitting Ra-226 (NAS80). These data are 

consistent with a dose-squared response (Ro78). Consequently, the 

NAS BEIR I11 Committee estimated bone cancer risk on the basis of 
both linear and quadratic dose response functions. However, as 

pointed out in NASRO, the number of U.S. dial painters at risk who 

received less than 1,000 rads was so small that the absence of 

excess bone cancer at low doses is not inconsistent with the linear 

response model. Therefore, the consistency of these data with a 

quadratic (or threshold) response is not remarkable and, perhaps, 

not relevant to evaluating risks at low doses. In contrast to the 

dial painter data, the incidence of bone cancer following shortlived radium-224 irradiation, observed in spondylitics by Mays and 

Spiess (Ma83, NAS80) in a larger sample at much lower doses, is 

consistent with a linear response. Therefore, for high-LET 

radiations, EPA has used a linear response function to evaluate the 

risk of bone cancer. 

Closely related to the choice of a dose response function is 

what effect the rate at which a dose of high-LET radiation is 

delivered has on its carcinogenic potential. This is an area of 

active current research. There is good empirical evidence, from both 

human and animal studies, that repeated exposures to radium-224 

alpha particles are 5 times more effective in inducing bone sarcomas 

than a single exposure that delivers the same dose (Ma83, NAS80). 

The 1980 NAS BEIR Committee took this into account in its estimates 

of bone cancer fatalities, which EPA is using. 

6.3.3 Assump~ions Made by EPA for Evaluating ~he Dose from 

Alpha-Parcicle Emit~ers 

EPA has evaluated the risk to specific body organs by applying 

an RBE of 8 for alpha radiations to the risk estimates for low dose 

rate, low-LET radiations as described above. For some organs, this 

factor may be too large. Several authors have noted that estimates 

of the risk of leukemia based on an RBE of 20 for bone marrow alpha 

irradiation (relative to a low dose rate, low-LET risk model which 

includes a DREF of 2.5) overpredicts the observed incidence of 

leukemia in persons receiving thorotrast (thorium oxides) (Mo79) and 

in the U.S. radium dial painters (Sp83). Nevertheless, in view of 

the paucity of applicable human data and the uncertainties discussed 

above, the ICRP quality factor provides a reasonable and prudent way 

of evaluating the risk due to alpha emitters deposited within body 

organs. 

All EPA risk estimates for high-LET radiations are based on a 

linear dose response function. For bone cancer and leukemia, EPA 

uses the absolute risk projection model described in the previous 

section. For other cancers, the Agency uses relative risk 

projections. 

The Agency's estimates of the risk of fatal cancer due to a 

uniform organ dose in various organs from internally deposited 

alpha-particle emitters are given in Table 6-11. The risk values in 

Table 6-11 were obtained by multiplying the average risk based on 

the linear model for a uniformly distributed whole-body dose of low- 
LET radiation by an RBE of 8 and then apportioning this risk by 

organ, as indicated. 

These estimates are for lifetime doses at a constant dose rate. 

This procedure was not followed for bone cancer. As outlined above, 

the risk estimate for this cancer in the BEIR I11 report is based 

directly on data for high-LET (alpha) radiation. Some readers may 

note that the risk estimate in Table 6-11, 19 bone cancer fatalities 

per lo6 person-rad, is less than the 27 fatalities listed in Table 

A-27 of NAS80 for alpha particles. This is because the analysis in 

Appendix A of NAS80 (but not Chapter V of that report) assumes that 

in addition to a 2-year minimum induction period, 25 years are 

available for cancer expression. This is usually not the case for 

doses received beyond about age 50. Hence, the estimated lifetime 

risk is smaller when it is based on a life table analysis that 

considers lifetime exposure in conjunction with competing causes of 

death. 

The next section describes how EPA estimates the risk due to 

inhalation of alpha-emitting radon progeny, a situation where the - organ dose is highly nonuniform. 

6.4 ESTIMATING THE RISK FROM LIFETIME POPULATION EXPOSURES FROM 

RADON-222 PROGENY 

- The Agency's estimates of the risk of lung cancer due to 

inhaled radon progeny do not use a dosimetric approach, but rather 

are based on what is sometimes called an epidemiological approach: 

that is, on the excess human lung cancer in groups known to have 

been exposed to radon progeny. 

When Rn-222, a radioactive noble gas, decays, a number of 

short half-life radionuclides (principally polonium-218, lead-214, 

bismuth-214, and polonium-214) are formed. These decay products, 

commonly referred to as "progeny" or "daughters," readily attach to 

inhalable aerosol particles in air. When inhaled, the radon progeny 

are deposited on the surfaces of the larger bronchi of the lung. 

Since two of these radionuclides decay by alpha-particle emission, 

the bronchial epithelium is irradiated by high-LET radiation. A 

wealth of data indicate that a range of exposures to the bronchial 

epithelium of underground miners causes an increase in bronchial 

lung cancer, both in smoking and in nonsmoking miners, and in some 

members of the general public. Recently the National Academy of 

Sciences, BEIR IV Committee, and the International Commission on 

Radiological Protection reviewed the question of radon risks and 

reported their conclusions (NAS88, ICRP87). 

Although considerable progress has been made in modeling the 

deposition of radon daughters in the lung, it is not yet possible to - characterize adequately the bronchial dose delivered by alpha 

radiation from inhaled Rn-222 progeny (NAS88). This is in part due 

to the uncertainty concerning the kinds of cells in which bronchial 
Table 6-1.1. Site-specific mortality risk per unit dose from 

internally deposited alpha emitters averaged for both 

sexes and ali ages. 

Site 

Mortality risk 

per lo6 personProportional risk" rad 

Leukemia 

Bone 

Thyroid 

Breast 

Lung 

Stomach 

Intestine 

Liver 

Pancreas 

Urinary 

Other 

Total 

" Proportion of whole-body risk from Table 6-10. 
cancer i.s i.niti.ated and the depth of these cells in the bronchial 

epithelium. 

Aside from the uncertainties in the dose calculations, a 

purely dosimetric approach to radon risk estimation appears 

untenable. Such an approach relates the risk from a given absorbed 

dose to the lung resulting from radon progeny exposure to that from 

gamma or x-ray exposure. This approach ignores the extensive 

epidemiological data on radon exposed miners and bases risk 

estimates indirectly on epidemiological studies of populations 

exposed to low-LET radiation. It must also, therefore, make use of 

an RBE for alpha particles estimated from animal studies. Given the 

uncertainties in the latter epidemiological studies and in the RBE, 

there would seem to be no advantage to this approach. Consequently, 

EPA agrees with the BEIR IV Committee conclusion that radon decay 

product dosimetry in the lung is only useful for extrapolating radon 

risk estimates from one exposure situation to another (NAS88). 

6.4.1 Characterizinq Exposures to the General Population 

vis-a-vis Underground Miners 

Exposures to radon progeny under working conditions are 

commonly reported in a special unit called the working level (WL). 

One working level is any combination of short half-life Rn-222 

progeny having 1.3 x 105 MeV per liter of potential alpha energy 

(FRC67). This value was chosen because it is the alpha energy 

released from the total decay of the short-lived radon progeny at 

radioactive equilibrium with 100 pCi/L of Rn-222. The WL unit was 

developed because the concentration of specific radon progeny 

depends on ventilation rates and other factors. A working level 

month (WLM) is the unit used to characterize a miner's exposure to 

one working level of radon progeny for a working month of about 170 

hours. Because the results of epidemiological studies are expressed 

in units of WL and WLM, the following outlines how they can be 

interpreted for members of the general population exposed to radon 

progeny. 

There are age- and sex-specific respiratory rate and volume 

differences, as well as differences in duration of exposure, in a 

general population as compared to a mining population. In earlier 

reports, EPA used an "exposure equivalent," a modified WLM in which 

adjustments were made for age-specific differences in airway 

dimensions and surface area, respiratory frequency, and tidal 

volume. These factors were expected to influence aerosol deposition 

and, therefore, radiation dose from radon daughters. This approach 

to quantifying exposure, correcting for differences in these 

factors, was recommended by Evans (Ev69) and is consistent with the 

original derivation of the working level (Ho57). 

- The BEIR IV Cornittee, however, concluded that the tracheobronchial "dose per WLM in homes, as compared to that in mines, 

differs by less than a factor of 2," and advised that the dose and 

risk per WLM exposure in residences and in mines should be 
considered to be identlical until better dosimetric estimates are 

developed (NAS881. EPA will follow the lead of the BEIR IV 

Committee in this regard and will not use the "exposure equivalentr' 

correction employed to compensate for age- and sex-specific tracheobronchial deposition in earlier EPA reports. In this report, 

exposure of any individual to 1 WL for 170 hours is 1 WLM and for l 

year is 51.56 WLM. This change puts EPA risk estimates in standard 

units generally used for this purpose, still without requiring dose 

calculations. 

For indoor exposure, an occupancy factor of 0.75 is still 

employed. Discussion of the support for this estimate can be found 

in EPA8 6. 

6.4.2 The EPA Model 

The initial EPA method for calculating radon risks has been 

described in detail (EPA79, E179). As new data were reported, the 

EPA revised its model to reflect changes, as contained in 

consecutive reports (EPA79, EPA82, EPA83a, EPA83b, EPA84, EPA85,and 

EPA86). The Agency initially projected radon lung cancer deaths for 

both absolute and relative risk models, but, since 1978, EPA has 

based risk estimates due to inhaled Rn-222 progeny on a linear dose 

response function, a relative risk projection model, and a minimum 

induction period of 10 years. A life table analysis has been used 

to project this risk over a full life span. Lifetime risks were 

initially projected on the assumption that an effective exposure of 

1 WLM increased the age-specific risk of lung cancer by 3 percent 

over the age-specific rate in the U.S. population as a whole 

(EPA79). In the most recent documents, lifetime risks were 

calculated for a range of risk coefficients from 1 percent to 4 

percent per WLM (EPA86). 

Although occupational exposures to pollutants other than Rn-222 

progeny are probably not important factors in the observed lung 

cancer risk for underground miners (E179, Th82, Mu83, Ra84, Se88) , the use of occupational risk data to estimate the risk of a general 

population is far from optimal, as it provides no information on the 

effect of radon progeny exposures for children and women. While for 

most estimates, it is assumed that the risk per unit dose received 

by children is no higher than that received by adults, this 

assumption may not be correct. 

The A-bomb survivor data indicate that, in general, the risk 

from childhood exposure to low-LET radiation is greater than from 

adult exposure and continues for at least 33 years, the time over 

which A-bomb survivors have been observed (Ka82). There are not, as 

yet, adequate age-specific data on occurrence of lung cancer in 

those under 10 years of age at the time of exposure (Ka82). Another 

limitation of the underground miner data is the absence of women in 

the studied populations. The A-bomb survivor data indicate women 

are as sensitive as men to radiogenic lung cancer from low-LET 
radiation even though, on the whole, they smoke less (Pr83). 

These data are not conclusive, however. 

6.4.3 Comparison of Earlier Risk Estimates 

Several estimates of the risk due to radon progeny have been 

published since the original EPA model was developed. These risk 

estimates were reviewed recently in a number of EPA reports 

(EPA894, EPA85, and EPA86) 

The recent EPA risk estimates for lifetime exposure to a 

general population, along with AECB, NAS, UNSCEAR, ICRP, and NCRP 

estimates of the risk of lung cancer resulting from inhaled radon 

progeny, are listed in Table 6-12. The AECB estimate for 

lifetime exposure to Canadian males is 830 fatalities per million - person-WLM (Th82). In Table 6-12, this estimate has been 

adjusted for the U.S. 1970 male and female population. . 

- The National Institute for Occupational Safety and Health 

reviewed published data on miner studies used as a basis for 

estimated risk coefficients and pointed out some of the strengths - and limitations of selected studies (NIOSH87). 

. The occupational exposure groups that constitute the 

epidemiological database for the risk estimates are as follows: 

I. U.S. Uranium Miners (NIOSH87) 

(a) Strengths: A large, clearly defined, well-traced cohort 

with some smoking histories and exposure records on the 

same persons. Standard sampling techniques were used to 

make measurements. 

(b) Limitations: There were few measurements in small mines, 

work histories were self-reported, exposures were high, 

and potential error due to excursions in exposure levels 

is high. 

(c) Follow-up: 19 years in 1977. 

2. Czechoslovakian Uranium Miners (NIOSH8 

(a) Strengths: Extensive exposure data with a large number 

of low level exposures and limited exposure to other 

underground mining. Many possible confounding 

factors have been investigated and eliminated. 
(b) Limitations: Exposure estimates prior to 1960 based on 

radon gas measurements. Person years at risk not 

determined in standard manner. Smoking effect 

neglected. Elevated levels of arsenic in ore. 

(c) Follow-up: 26 years in 1975. 

3. Ontario Uranium Miners (NIOSH87) 

(a) Strengths: Miners received low mean cumulative exposures. 

Prior mining experience was carefully traced. Exposures 

prior to 1967 may be disputed. 

(b) Limitations: Median age of the cohort was 39 years in 

1977. Thoron and gamma exposures may have been high but 

not accounted for. Smoking history is limited. 

(c) Follow-up: 18 years in 1977 

4. Malmberget Iron Miners (NIOSH87) 

(a) Strengths: Low exposure levels, long follow-up and 

stability of work force. Complete ascertainment of vital 

status and confirmation of diagnosis. Risk from 

confounders was examined and ruled out. 

(b) Limitations: Relatively small cohort with limited 

exposure data and an unclear cohort definition. 

(c) Follow-up: 44 years in 1976 

5. Eldorado - Uranium Miners (NAS88) 

(a) Strengths: Very low exposure rates, miners screened 

for prior mining experience, roughly equal groups 

of surface only and underground only miners, Silica 

and diesel exhaust exposures low. Potential 

confounders investigated. 

(b) Limitations: Exposure estimates are disputed. Sixteen 

percent of the miners excluded for incorrect or missing 

data. Average age in 1980 was 43 years. 

(c) Follow-up: 14 years in 1980. 
Table 6-12. Risk estimate for exposures to radon progeny 

Organization Model Fatalities per Exposure period Expression 

lo6 person-WLM period 

EPA Rel. 760 (460)" Lifetime Lifetime 

NAS* A-S Abs. 730 (440)" Lifetime Lifetime 

AECB~ Re1 . 600 (300)" Lifetime Lifetime 

ICRP - 150-450 Working Lifetime 30 years 

UNSCEAR - 200-450 Lifetime 40 years 

NCRP' Dec. Abs. 130 Lifetime 

Lifetime 

" EPA and AECB based their estimates of risk for the general 

population on an exposure equivalent, corrected for breathing 

rate (and other factors). For comparison purposes, the values 

in parentheses express the risk in more customary units, in 

which a continuous annual exposure to 1 WL corresponds to 51.6 

WLM. 

Adjusted for U.S. General Population: see text. 

NCRP84: Table 10.2; assumes risk diminishes exponentially with 

a 20-year halftime, and no lung cancer risk is expressed before 

age 40. 

Sources: EPA83b; NAS80; Th82; 1CRP81; EPA86; UNSCEAR77; NCRP84; 

USRPC80. 

Models: Rel. - Relative Risk Projection 

A-S Abs. - Age-Specific Absolute Risk Projection 

Dec. Abs. - Decaying Absolute Risk Projection 
6.4.4 Recent Radon Risk Estimates 

6.4.4.1 BEIR IV 

At the beginning of 1988, the National Academy of Sciences 

released the BEIR IV Committee report, reviewing information on the 

risks from radon and other alpha-emitting radionuclides (NAS88). 

With the cooperation of the principal investigators, BEIR IV 

examined in detail the mortality experience of four cohorts of 

underground miners (the U.S., Ontario, and Eldorado uranium miners 

and the Malmberget iron miners) and how the mortality related to 

radon daughter exposure. The Committee calculated the relationship 

of age-specific relative risk to exposure level and time-sinceexposure (TSE) in two analyses. The first used internal cohort 

comparisons and was a grouped-data analog of a Cox relative-risk 

regression (NAS88). The second analysis compared the cohorts with 

external rates and was a generalization of standard SMR methods. 

Separate parallel analyses were carried out to establish a single 

combined value for each parameter. 

The mathematical form of the Committee's preferred TSE model for 

the radon related age-specific mortality rate at age a is 

where r,(a) = age-specific lung cancer mortality rate 

y(a) = 1.2, if a is less than 55 years 

1.0, if a is between 55 and 64 years 

0.4, if a is greater than 64 years 

W, = WLM incurred between 5 and 15 years prior to age a 

W, = WLM incurred more than 15 years prior to age a 

The Committee model is, therefore, an age-specific, relative-risk 

projection model with a 5-year latent period prior to expression of 

risk. 

The BEIR IV Committee also estimated what the lung cancer risk 

coefficient would be for an age-constant, relative-risk model. The 

results of this analysis are summarized below: 
Cohort Excess Risk 

per WLM 

95% Confidence 

Limits 

u. S. 

Ontario 

Eldorado 

Malmberget 

Combined 

In its analysis, the BEIR IV Committee identified two major 

areas of uncertainty affecting its conclusions: (1) uncertainty 

related to the Committee's analysis of cohort data and (2) 

uncertainty related to projection of the risk to other groups. The 

Committee's TSE model uses risk coefficients derived from analysis 

of data from four miner cohorts. Random or systematic errors, 

particularly systematic errors, could bias the conclusions. Sources 

of error in addition to basic sampling variation include: (1) errors 

in exposure estimates, particularly since the magnitude of error may 

differ among the studies; (2) errors of misclassification of cause 

of death; (3) errors in smoking status of individual miners, and (4) 

modeling uncertainty--i.e., does the model properly address all 

parameters that are determinants of risk? 

Having developed the TSE model for miners, the Committee 

anticipated the folLowing sources of uncertainty in projecting the 

model across other groups: (1) effect of gender (miner data all for 

males); (2) effect of age (miner data contain no information on 

exposures before about age 20); (3) effect of smoking (miner data 

contain poor information on smoking status); (4) temporal expression 

of risk (not enough miners have died to establish accurately the 

pattern of lifetime risk from radon exposure), and (5) extrapolation 

from mining to indoor envirenments (what are significant differences 

in the air in mines compared to air indoors?). After reviewing the 

various sources of uncertainty, the BEIR IV Committee concluded 

[p42]," ... The imprecision that results from sampling variation can 

be readily quantified, but other sources of variation cannot be 

estimated in a quantitative fashion. Therefore, the Committee chose 

not to combine the various uncertainties into a single numerical 

value" (NAS88) . 

The question of errors in exposure estimates is particularly 

interesting since the modeling is strongly influenced by the U.S. 

uranium miner data. In fact, the model risk estimates would be 33 

percent higher if the U.S. cohort was removed. Exposure in the U.S. 

- cohort is poorly known: cumulative WLM (CWLM) are calculated from 

measured radon levels for only 10.3 percent of the miners, varying 

amounts of estimation are required for about 36.1 percent of the 
miners, and guesswork is used for about 53.6 percent of the miners 

(NAS88, Lu71). Only 26.1 percent of the U.S. uranium miner exposure 

data are based on measured values (Lu71). 

The Ontario cohort exposure estimates also are not well 

founded. Upper and lower estimates were developed: the Lower from 

measured values, the upper based on engineering judgment (NAS88). 

Eldorado cohort estimates of CWLM were based almost entirely on 

measured values, while Malmberget cohort estimates were based on a 

reconstruction of past ventilation conditions (NAS88). Of the four 

cohorts, the United States has one of the poorest bases for CWLM 

estimates. One serious problem is the potential error due to large 

excursions in radon daughter concentrations (NIOSH87). The 

uncertainties in exposure estimates are particularly significant in 

view of the rather large impact the U.S. cohort has on the form of 

the model. 

When the BEIR IV model is run with the 1980 lifetable and 

vital statistics at an exposure level of 0.001 WLM per year, the 

reference risk can be calculated. 

BEIR IV Relative Risk Model - Lifetime Exposure and Lifetime Risk 

Group Risk (10'6/WLM) 

Male 

Female 

Combined 

6.4.4.2 ICRP 50 

The International Commission on Radiological Protection, in 

its Publication 50, addressed the question of lung cancer risk from 

indoor radon daughter exposures. The ICRP Task Group took a 

direction quite different from the BEIR Committee. The Task Group 

reviewed published data on three miner cohorts: U.S., Ontario, and 

Czech uranium miners. The estimated risk coefficients by cohort 

are : 

Cohort Follow-up Relative model Absolute model 

u.S. 1950-1977 0.3%-1.0% 2-8 cases/106Pw~~~ 

Czech 1948-1975 1.0%-2.0% 10-25 cases/106 PWLMY 

Ontario 1958-1981 0.5%-1.3% 3-7 cases/106 PWLMY 

Average 1 % 10 cases/106 PWLMY 
The relative risk model then developed for a constant exposure rate 

is : 

= the mortality rate at age t 

where : 

h,(t) = the age-specific lung cancer rate at age t 

r(t,) = risk coefficient at age of exposure t, 

E(t.) = age-dependent exposure rate 

7 = time lag (minimal latency) = 10 year 

In the case of a constant exposure rate or constant annual 

exposure, the equation collapses to: 

h(t) = h,(t) [l + r E(t - 7)l 

where: 

- 

r = age averaged relative risk coefficient 

= cumulative exposure to radon daughters to age 

t -7 

Since ICRP recommends the use of the relative risk model, the 

ICRP 50 absolute risk model will not be addressed further in this 

document. 

To adapt the relative risk model derived from studies of 

underground miners for the general population, the ICRP Task Group 

introduced several adjustments. The first was to correct for cocarcinogenic influences in mines. To account for unidentified, 

unproven carcinogens that might be present in mine environments but 

not elsewhere, only 80 percent of the risk was attributed to radon. 

The second adjustment was for dosimetric corrections. The dose to 

bronchial epithelium used by the Task Group for persons indoors was 

estimated to be only 80 percent as large as that for persons in 

- mines; therefore, the risk to the public from radon was considered 

to be 80 percent of the risk of miners. 
Adjusting the average relative risk coefficient of 

1 percent per WLM by these two factors gives a risk coefficient of 

0.64 percent per WLM: 

The third adjustment made by the Task Group is related to age. 

Since reports of Japanese A-bomb survivors and some other radiationexposed groups support an elevated estimate of risk in children 

compared to adults, the Task Group increased the risk coefficient of 

persons between birth and age 20 by a factor of 3. 

The final relative risk coefficients in the ICRP 50 model are: 

1.9 percent per WLM if the age at time of exposure is between birth 

and 20 years, and 0.64 percent per WLM if age at time of exposure 

exceeds 20 years. 

When the ICRP 50 relative risk model is run with 1980 U.S. 

lifetable and vital statistics at an exposure level of 0.001 WLM per 

year, the reference risk calculated is: 

Group 

Male 

Female 

Combined 

Risk (10-6/~LM) 

6.4.5 Selection of Risk Coefficients 

To estimate the range of reasonable risks from exposure to Rn222 progeny for use in the Background Information Document for 

Underground Uranium Mines (EPA85), EPA averaged the estimates of 

BEIR 111, the EPA model, and the AECB to establish an upper bound of 

the range. The lower bound of the range was established by 

averaging the UNSCEAR and ICRP estimates. The Agency chose not to 

include the NCRP estimate in its determination of the lower bound 

because this estimate was believed to be outside the lower bound. 

With this procedure, the EPA arrived at relative risk coefficients 

of 1.2 percent to 2.8 percent per WLM exposure equivalent (300 to 

700 fatalities per million person-WLM exposure equivalent) as 

estimates of the possible range of effects from inhaling Rn-222 

progeny for a full lifetime. Although these risk estimates did not 

encompass the full range of uncertainty, they seemed to illustrate 

the breadth of much of current scientific opinion. 

The lower limit of the range of 1985 EPA relative risk 

coefficients, 1.2 percent per effective WLM, was similar to that 

derived by the Ad Hoc Working Group to Develop Radioepidemiological 

Tables, which also used 1.2 percent per WLM (NIH85). However, some 

other estimates based only on U.S. and Czech miner data averaged 

1 percent per WLM (Ja85) or 1.1 percent per WLM (St85). On the 
other hand, three studies - two on miners (Ra84, Ho86) and one on 

residential exposure (Ed83, 84) - indicated a relative risk 

coefficient greater than 3 percent per WLM, perhaps as large as 3.6 

percent. 

The EPA therefore increased the upper limit of its estimated 

range of relative risk coefficients. To estimate the risk due to 

Rn-222 progeny, the EPA used the range of relative risk coefficients 

of 1 to 4 percent per WLM. (See EPA86 for a more detailed 

discussion.) Based on 1980 vital statistics, this yielded, for 

members of the general public, a range of lifetime risks from 380 to 

1,520 fatal cases per lo6 WLM (expressed in exposure equivalents). 

In standard exposure units, uncorrected for breathing rate and age, 

this corresponds to 230 to 920 cases per lo6 WLM. Coincidentally, 

the geometric mean estimate obtained in this way with 1980 vital 

statistics, 4.6xlO-'/~LM in standard units of exposure, is 

numerically the same as that obtained using a 3 percent relative 

risk coefficient and 1970 vital statistics (see Table 6-9). . 

However, in light of the two recently published consensusbased reports, BEIR IV and ICRP 50, and a recent report on the Czech 

miner groups (Se88), the Agency has reviewed its basis for radon . 

risk estimation. Comparable relative risk coefficients for miners 

(age-constant relative risk) yield a coefficient of around 1 percent . in ICRP 50, 1.34 percent in BEIR IV, and 1.5 percent in the Czechs. 

This suggests that the range, 1 percent to 4 percent, used by EPA 

may be too wide. Nevertheless, note that only 5 of the 20 or so 

studies for which there are some data are included in these 

estimates. 

The BEIR IV Committee noted and modeled a drop in relative risk 

with increasing time of exposure and a decreasing relative risk with 

increasing age after exposure (NAS88). The Czech miners show a 

similar response pattern (Se88). Though the Committee did note a 

dose rate effect in the U.S. uranium miner cohort, i.e., a decrease 

in risk per unit exposure at high dose rates, it was not included in 

the model (NAS88). The possibility of a similar dose-rate effect 

was found recently in a study on Port Radium uranium miners (Ho87). 

The ICRP 50 Task Group worked from a different database and 

developed a simpler model with fewer age- and time-dependent 

parameters. The Task Group provided a 3 times higher risk for 

exposure between birth and 20 years of age than after 20 years of 

age (ICRP87). The finding in the recent Czech report that risk 

prior to age 30 is 2 to 2.5 times greater than after age 30 lends 

some support to the ICRP conclusions (Se88). 

Both BEIR IV and ICRP 50 models treat radon and smoking risks 

- as multiplicative. This conclusion is based primarily on data from 

the U.S. uranium miner cohort. Although apparently based on weaker 

evidence, the report on Malmberget miners and the recent report on 
Czech miners both concluded that the interaction of smoking and 

radon exposure is small (Ra84, Se88). The attributable risk per 

unit exposure in smokers and non-smokers was essentially the same 

(Se88). The true interaction of radon and cigarette smoking is 

controversial. Both antagonistic (Ax78, Lu79, Ax80) and 

multiplicative (Lu69, Wh83) interactions have been reported in man, 

and animal studies can be found to justify any position (Ch81, Ch85, 

Cr78). In prior calculations, EPA has always treated the 

interaction between radon daughters and cigarette smoke as 

multiplicative. EPA will continue to treat the radon daughter-smoke 

interaction as multiplicative at this time. 

Important unresolved issues pertaining to the risks from 

inhaled radon progeny remain. At the advice of the Radiation 

Advisory Committee of EPA's Science Advisory Board, EPA will 

continue to use relative risk models but shall include both BEIR IV 

and ICRP 50 model calculations to illustrate the difference in 

results from the two models. The ICRP 50 model will be slightly 

modified. The risk reduction factor of 0.8 to compensate for 

differences in dosimetry will be removed to place the ICRP 50 model 

and BEIR IV model on a comparative basis. Calculations in the ICRP 

50 model will be made using risk coefficients of 2.4 percent per WLM 

from birth to age 20 and 0.8 percent per WLM for ages greater than 

20 years, yielding estimates listed in Table 6-13. 

Table 6-13 summarizes risk estimates based on the BEIR IV and 

the ICFiP 50 model, modified as described above. For the 

calculations in this document, both models were adjuste for the 

effect of background radon exposure (see section below). 

Table 6-13. Lifetime risk from radon daughter exposure of lung 

cancer death (per 106 WLM) . 

Model 

Group BEIR IV ICRP 50 

Men 

Women 

Combined Population 350 

(Range) -- 

The ICRP Task Group concluded that, all things considered, the 

range of variation of the mean relative risk coefficient is from 

about 0.3 up to 2 times the value stated (ICRP87). The range of 

risk cited in Table 6-13 for the ICRP model reflects this 
The Committee analyzed the A-bomb survivor data in two separate 

sets: (1) leukemia and (2) all cancers excluding leukemia (solid 

cancers). Its treatment of these two cases was not equivalent. The 

analysis of leukemia considered the Nagasaki and Hiroshima data 

separately. The Committee's regression analysis of the leukemia 

mortality data provided stable values for all of the coefficients in 

Equation 6-3, and hence for the neutron RBE and the ratio of linear to 

dose-squared terms for leukemia induction by gamma rays, as a function of 

dose. Estimating the linear-quadratic response coefficients for solid 

cancers proved to be less straightforward, however, and it was decided 

that the observations on solid cancers were "not strong enough to provide 

stable estimates of low dose, low-LET cancer risk when analyzed in this 

fashion" (NAS80,p.186). The Committee decided to use a constrained 

regression analysis, carrying over some of the parameters for Equation 6- 

3 found in its analysis of leukemia deaths to the regression analysis of 

the dose response for solid cancers. Specifically, both the neutron RBE 

at low dose (the ratio of the coefficient K, to C,) and the ratio of C, to 

C,, as estimated from the leukemia data, were assumed to apply to the 

induction of fatal solid cancers. These estimates became the basis for 

the "preferred" linear-quadratic (LQ-L) risk estimates for solid cancers 

presented in BEIR I11 

(NAS80, p. 187). 

6.5.2 Uncertainty in the Dose Response Models 

Due to Bias in the A-bomb Dosimetry 

Investigators from Oak Ridge National Laboratory carried out careful 

state-of-the-art evaluation of the dose to A-bomb survivors in the early 

1960s (Au67, Au77). The results of these studies resulted in a "T65" 

dose being assigned to the dose (kerma) in free air at the location of 

each survivor for both gamma rays and neutrons. A major conclusion of 

the ORNL study was that the mix of gamma ray and neutron radiations was 

quite different in the two cities where A-bombing occurred. These 

results indicated that at Hiroshima the neutron dose was more important 

than the gamma dose when the greater biological efficiency of the highLET radiations produced by neutrons was taken into account. Conversely, 

the neutron dose at Nagasaki was shown to be negligible compared to the 

gamma dose for that range of doses where there were significant numbers 

of survivors. Therefore, the 1980 BEIR Committee evaluated the cancer 

risks to the survivors at Hiroshima on the assumption that the combined 

effects of gamma rays and particularly neutrons caused the observed 

cancer response. 

Serious inadequacies in the T65 dosimetry system were discovered in 

the late 1970s. A comprehensive reevaluation of the doses to survivors 

was carried out under the auspices of the U.S.-Japan Joint Committee for 

Reassessment of Atomic Bomb Dosimetry in Hiroshima and Nagasaki. In 

1986, this committee provided results to the Radiation Effects Research 

Foundation (RERF) from which a revised dosimetry system, termed "DS86," 

was developed. Although work on the DS86 is largely complete, small 

adjustments in dose estimates are anticipated over the next few years 

(Pr87). In addition, about 1,000 survivors from Nagasaki, who were 

shielded by terrain or were in factories, have so far been excluded from 
* * * DRAFT * * * 

the analysis because of difficulties in estimating their doses. It is 

anticipated that dose estimates for some of these survivors will be 

forthcoming in the near future (Pr87). 

The major differences between T65 and DS86 are: (1) the neutron dose 

in DS86 is decreased to 10 percent of its former value in Hiroshima and 

30 percent in Nagasaki (as a result, neutrons now contribute relatively 

little to the estimated excess of cancers in the two cities); (2) the 

DS86 free-in-air gamma dose increases somewhat in Hiroshima but decreases 

in Nagasaki relative to T65; (3) transmission of gamma rays through 

wooden structures is decreased by about a factor of 2 in DS86; and (4) 

Transmission of gamma rays through the body to internal organs is 

generally increased, partially nullifying the change associated with the 

decreased transmission through structures (Pr87, Sh87). 

Analysis of the A-bomb survivor data using the DS86 dosimetry is 

continuing. Preliminary indications are that risk estimates 

corresponding to a given dose-response model (linear or linear quadratic) 

will be increased by more than a factor of 2 as compared to BEIR I11 

estimates. This increase arises not only from changes in dosimetry, but 

also from further epidemiological follow-up and new statistical 

procedures employed (Pr87, 88). A preliminary estimate of low-LET 

radiation risk to the general population based on DS86 dosimetry and the 

linear, relative risk model is 1.2 x 10-"atal cancers per rad (Pr88) - approximately 3 times the corresponding BEIR 111 estimate. 

It appears that either a linear or linear-quadratic dose response is 

consistent with the survivor data, analyzed according to DS86 (Pr87). It 

would also appear that the residual difference in risk per unit dose 

between Hiroshima and Nagasaki is no longer statistically significant 

under DS86 dosimetry (Sh87) . 

From the standpoint of estimating risks from low-level, low-LET 

radiation, however, the most important result of the new dosimetric 

calculations may be in helping to determine which models best describe 

the data on human radiation carcinogenesis. After all, the greatest 

uncertainties in radiation risk estimation generally reflect model 

uncertainties, not uncertainties in the magnitude of risk coefficients. 

6.5.3 Sampling Variation 

Besides the systematic bias in the BEIR I11 risk estimates for lowLET radiation outlined above, the precision of the estimated linear and 

linear-quadratic risk coefficients in the BEIR I11 report is limited by 

statistical fluctuations due to sample size. The uncertainty bounds (2 1 

SD) attached to the gamma-ray risk coefficient in the BEIR 111 linear 

model are about +25 percent, for either leukemia (Table V-8) or for all 

other cancers able V-11). For the latter groups of cancers, however, 

the neutron RBE was constrained to the value obtained from analysis of 

the leukemia data. If this constraint is removed, the uncertainty in the 

estimate increases to +I50 percent (Table V-9). This increase reflects 

the large uncertainty associated with the neutron contribution in the 

analysis and the strong correlation between neutron and gamma-ray doses. 
Following the dosimetry reassessment, neutron doses will decrease 

markedly but will remain correlated with gamma-ray doses. 

Finally, it should be noted that random errors in dosimetry tend to 

bias the slope of the dose response curve downward, reducing the estimate 

of risk (Da75, Gi84, Ma59). The amount of bias introduced depends on the 

size of the random errors in the dose estimates and their distribution, 

which are unknown at present. 

6.5.4 Low Dose Extrapolation 

Results from animal and cellular studies often show decreasing 

effects (e.g., cancers, mutations, or transformations) per rad of low-LET 

radiation at low doses and dose rates. Based on a review of this 

literature, the National Council on Radiation Protection (NCRP80) has 

concluded that "linear interpolation from high doses (150 to 350 rads) 

and dose rates (>5 rads min-') may overestimate the effects of either low 

doses (0-20 rads or less) or of any dose delivered at dose rates of 5 rad 

y-l or less by a factor of 2 to 10." Judged solely from laboratory 

experiments, therefore, about a factor of 10 reduction from the linear 

prediction would seem to constitute a plausible lower limit on the 

effectiveness of low-LET radiation under chronic low dose conditions. 

Epidemiological evidence, however, would seem to argue against such a 

large DREF for human cancer induction. 

Data on the A-bomb survivors and patients irradiated for medical 

reasons indicate that excess breast cancer incidence is proportional to 

dose and independent of dose fractionation (NAS80, NIH85). The evidence 

regarding thyroid cancer induction is less firm, but the data would again 

suggest a linear dependence on dose (NASBO, NIH85). Another cancer for 

which human data provide a test of dose response models is leukemia. An 

analysis of the A-bomb survivor data based on T65 dosimetry suggests a 

quadratic component. The simple linear model, however, fits much better 

under DS86 than T65 (Sh87). The best estimate of the linear coefficient 

obtained from the linear quadratic fit to the data, under either T65 or 

DS86, is only about a factor of 2.5 less than the coefficient derived 

from the linear model. Thus, while the animal data indicate a reduced 

effectiveness of radiation at low doses and. low dose rates, the available 

human data would suggest that the linear dose response model does not 

overestimate the risk by more than about a factor of 3 under these 

conditions. 

6.5.5 Other Uncertainties Arising from Model Selection 

In addition to a dose response model, a "transportation model" is 

needed to apply the risks from an observed irradiated group to another 

population having different demographic characteristics. A typical 

example is the application of the Japanese data for A-bomb survivors to 

Western people. Seymour Jablon (Director of the Medical Follow-up Agency 

of the National Research Council, NAS) has called this the 

"transportation problem," a helpful designation because it is often 

confused with the risk projection problem described below. However, 

there is more than a geographic aspect to the "transportation problem," 
Risk estimates for one sex must sometimes be based on data for the other. 

In transporting risk estimates from one group to another, one may have to 

consider habits influencing health statcs, such as differences between 

smokers and nonsmokers, as described in Section 6.4 for the case of risk 

estimates for radon progeny. 

The BEIR 111 Committee addressed this problem in its 1980 report and 

concluded, based largely on the breast cancer evidence, that the 

appropriate way to transport the Japanese risk to the U.S. population was 

to assume that the absolute risk over a given observation period was 

transferrable but that relative risk was not. Therefore, the Committee 

calculated what the relative risk would be if the same number of excess 

cancer deaths was observed in a U.S. population having the same age 

characteristics as the A-bomb survivors. A constant absolute risk model, 

as postulated by the Committee, would imply that, whatever the factors 

are that cause Japanese and U.S. baseline cancer rates to differ, they 

have no effect on the incidence of radiation-induced cancers; i.e., the 

effects of radiation and these factors are purely additive. 

An alternative approach to the "transportation problem" is that of 

the 1972 NAS BEIR-I Committee. This Committee assumed relative risks 

would be the same in the United States and Japan and transferred the 

observed percentage increase directly to the U.S. population. Since the 

U.S. and Japanese baseline rates differ drastically with respect to 

mortality from specific cancers, this approach implies some large 

differences in the predicted number of specific cancers resulting from a 

given dose of radiation in the two countries. The most important 

differences relate to cancers of the breast, lung, and stomach. Baseline 

rates of breast and lung cancers are higher in the United States by 

factors of about 4 and 2, respectively, while the risk of stomach cancer 

is about 8 times higher in Japan (Gi85). As noted above, it now appears 

that the absolute risk should be transported for breast cancer. Evidence 

is lacking regarding the other diseases, however. If lung cancer risk 

were to be transported with a relative risk model, retaining the absolute 

model for other cancers, the estimated risk from a whole-body exposure 

would increase by about 20 percent; on the other hand, applying the 

relative risk model to stomach cancer alone would lower the whole-body 

risk by about 8 percent. Based on these considerations, including the 

tendency for changes in specific cancers to cancel one another, EPA 

believes that using the absolute risk "transportation model" is unlikely 

to cause errors of more than +20 percent in the total risk estimate. 

Thus, in the case of uniform whole-body doses, the amount of uncertainty 

introduced by transporting cancer risks observed in Japan to the U.S. 

population appears to be small compared to other sources of uncertainty 

in this risk assessment. 

Also needed to estimate risk is a risk projection model. As outlined 

in Section 6.2, such models are used to project future risks as an 

exposed population ages. For leukemia and bone cancer, where the 

expression time is not for a full lifetime but rather 25 years, absolute 

and relative risk projection models yield the same number of radiogenic 

cancers, but would distribute them somewhat differently by time after 

exposure, and hence by age. For solid cancers or other than bone, the 
BEIR ZII Committee assumed that radiogenic cancers would occur throughout 

the estimated lifetime. This makes the choice of projection model more 

critical because the relative risk projection yields estimated risks 

about three times larger than those obtained with an absolute risk 

projection, as shown in Table 6-2. Recent follow-up of the A-bomb 

survivor population strongly suggests that the relative risk projection 

model better describes the variation in risk of solid tumors over time 

(NIH85). However, there may be some cancers, apart from leukemia and 

bone cancers, for which the absolute risk projection model is a better 

approximation to reality. For other cancers, the relative risk may have 

been roughly constant for the current period of follow-up but may 

eventually decrease over time. Thus, while the relative risk model was 

used in this report for calculating a "best estimate" of the lifetime 

risk of solid tumors, it may overestimate the risk by as much as a factor 

of 2. 

Similarly, there is as yet insufficient information on - 

radiosensitivity as a function of the age at exposure. The age-dependent 

risk coefficients used here are those presented in the BEIR 111 report. - As yet, there is little information on the ultimate effects of exposure 

during childhood. As the A-bomb survivor population ages, more 

information will become available on the cancer mortality of persons 

irradiated when they were young. Table 6-2 indicates that the more 

conservative BEIR-I assumption for the effect of childhood exposures 

would increase BEIR I11 risk estimates by about 40 percent. Recent - follow-up supports the view that relative risks are highest in those aged 

0-9 yr at exposure. Full inclusion of the projected effects on this 

group was a major contributor to the increase in risk found with the 

recent analysis based on DS86 dosimetry (Pr87, 88). The BEIR I11 

Committee did not include -- in utero exposures when calculating population 

risks for radiogenic cancer because it felt the estimate of the effect of 

-- in utero radiation is highly uncertain. EPA has deferred to the 

committee's judgment in this regard. The BEIR-I report did include in 

utero cancer risk, but this had little effect (1 to 10 percent) on ts 

lifetime risk of cancer from lifetime exposure. An effect thls small is 

not significant relative to other sources of uncertainty in the risk 

assessment. 

6.5.6 Summary and Conclusions Regarding Uncertainties 

in Cancer Risk Estimates 

The most important uncertainties in estimating risk from whole body, 

low-LET radiation appear to relate to: (1) the extrapolation of risks 

observed in populations exposed to relatively high doses, delivered 

acutely, to populations receiving relatively low dose chronic exposures 

and (2) the projection of risk over a full lifespan - most critically, 

the extent to which high relative risks seen over a limited follow-up 

period among individuals exposed as children carry over into later years 

of life when baseline cancer incidence rates are high. 

The EPA central estimate of average lifetime risk, approximately 400 

fatal cancers per lo6 person-rad, is taken from the NAS BEIR 111 Committee 

report (NAS1980), incorporating the most conservative model assumptions 
utilized by the Committee - i.e., a linear dose response and age-specific 

relative risks projected over a lifetime for solid tumors (L-RR model). 

For reasons discussed above, it would now appear that estimates of 

average lifetime risk based on the L-RR model assumptions must be revised 

upwards - to roughly 1,200 fatal cancers/106 person-rad. Although further 

analysis of the A-bomb survivor data may increase this estimate, the 

conservatism inherent in the model's assumption supports the view that 

the 1,200/1O6 value is an upper bound, pending release of NAS BEIR V 

report now in preparation. 

Animal data would suggest that the linear dose response may 

overestimate risk by roughly a factor of 3. Likewise, while the 

epidemiological data clearly indicate an increase in risk with age at 

expression, the (age-specific) constant relative risk projection may 

overstate lifetime risk by about a factor of 2 or 3. Allowing even for 

additional sources of uncertainty discussed above, it would appear that 

the upper bound (L-RR) model estimate may be high by a factor of 5 to 10. 

Therefore, as a lower bound estimate of the average lifetime risk, a 

value which is one-tenth the upper bound, or 120 fatal cancers/106 personrad, has been adopted. 

The L-RR model estimate from BEIR 111, about 400 fatal cancers/106 

person-rad, falls near the geometric mean of what tentatively appears to 

be a reasonable range for the estimate of risk, based on current 

information. EPA has chosen the BEIR 111, L-RR model value as its 

"central estimate." It should be emphasized that this estimate cannot be 

regarded as conservative in the sense of providing any significant margin 

of safety with respect to public health protection. The decision by EPA 

to employ the central estimate of 400 fatalities/106 person-rad and a 

range of 120-1,200 fatalities/106 person-rad was reviewed and approved by 

a special panel set up by the Agency's outside Radiation Advisory 

Committee and by the Committee itself, as an interim measure for this 

proposed rulemaking. 

The uncertainty in risks for specific cancer sites may be 

substantially larger than the uncertainty in the whole-body risk. One 

reason is that the epidemiological data pertaining to some sites may be 

very sparse. In addition, the uncertainty in projecting risk from one 

population to another (e.g., Japanese to U.S.) is important at sites for 

which incidence rates differ markedly between populations. 

The uncertainties in risk associated with internally deposited alpha 

emitters are often greater than for low-LET radiation. Human 

epidemiological data on the risks from alpha emitters are largely 

confined to: (1) lung cancer induced by radon decay products (see 

below); (2) bone cancer induced by radium; and (3) liver cancer induced 

by injected thorotrast (thorium). Many of the risk estimates presented 

here for alpha irradiation are extrapolated from those for low-LET 

irradiation, assuming an RBE of 8, as determined from high dose 

experiments on animals. The available evidence on cells, animals, and 

humans points to a linear dose response relationship for the risk from 

alpha emitters (NAS88). The extrapolation to low doses is therefore 

considered to be less important as a source of uncertainty for alpha 
irradiation than for low-LET irradiation. There is, however, 

considerable variability in the RBE determined from animal studies; the 

extrapolation of these results to humans is also problematic. 

For many alpha-emitting radionuclides, the most important source of 

uncertainty in the risk estimate is the uncertainty in dose to target 

cells. Contributing to this uncertainty is uncertainty in the location 

of these cells, ignorance regarding the metabolism of the radionuclide, 

uniformity of radionuclide deposition in an organ, and the short range of 

alpha particles in tissue (see Chapter 5). 

In the case of alpha irradiation of the lung by radon decay products, 

there are human epidemiological data that allow direct estimation of the 

risk per unit exposure. Knowledge of RBE and the actual dose to target 

cells is therefore not important except as the dose per unit exposure 

might differ between mine and indoor environments. As a consequence, the 

estimated uncertainty in average radon risk estimates is similar to that 

for low-LET radiation. As discussed in Section 6.4.5, the EPA is 

employing a central risk estimate for excess radon exposure of 360 fatal 

lung cancers/106 WLM and an uncertainty range of 160-720 fatal lung 

cancers/106 WLM. 

6.6 OTHER RADIATION-INDUCED HEALTH EFFECTS 

The earliest report of radiation-induced health effects was in 1896 

(Mo67), and it dealt with - acute effects in skin generally caused by very 

large x-ray exposures. Within the six-year period following, 170 

radiation-related skin damage cases had been reported. Such injury, like 

many other acute effects, is the result of exposure to hundreds or 

thousands of rads. Under normal situations, environmental exposure does 

not cause such large doses, so possible acute effects will not need to be 

considered in assessing the risk to the general population from routine 

radionuclide emissions. 

Radiation-induced carcinogenesis was the first delayed health effect 

described: the first case was reported in 1902 (Vo02), and 94 cases of 

skin cancer and 5 of leukemia were reported by 1911 (Up75). Radiationinduced genetic changes were noted soon afterward. In 1927, H.J. Muller 

described x-ray-induced mutations in animals (in the insect, Drosophila), 

and in 1928, L.J. Stadler reported a similar finding in plants (Ki62). 

At about the same time, radiation effects on the developing human embryo 

were observed. Case reports in 1929 showed a high rate of microcephaly 

(small head size) and central nervous system disturbance and one case of 

skeletal defects in children irradiated -- in utero (UNSCEAR69). These 

effects, at unrecorded but high exposures and at generally unrecorded 

gestational ages, appeared to produce central nervous system and eye 

defects similar to those reported in rats as early as 1922 (Ru50). 

For purposes of assessing the risks of environmental exposure to 

radionuclide emissions, the genetic effects and -- in utero developmental 

effects are the only health hazards other than cancer that are addressed 

in this Background Information Document (BID). 
6.6.1 Types of Genetic iiarm and Duration of Expression 

Genetic harm (or the genetic effects) of radiation exposure is 

defined as stable, heritable changes induced in the germ cells (eggs or 

sperm) of exposed individuals, which are transmitted to and expressed 

only in their progeny and in future generations. 

Of the possible consequences of radiation exposure, the genetic risk 

is more subtle than the somatic risk, since it affects not the persons 

exposed, but relates only to subsequent progeny. Hence, the time scales 

for expression of the risk are very different. Somatic effects are 

expressed over a period on the order of a lifetime, while about 30 

subsequent generations (nearly 1,000 years) are needed for near comp1.ete 

expression of genetic effects. Genetic risk is incurred by fertile 

people when radiation damages the nucleus of the cells which become their 

eggs or sperm. The damage, in the form of a mutation or a chromosomal 

aberration, is transmitted to, and may be expressed in, a child conceived 

after the radiation exposure. However, the damage may also be expressed 

in subsequent generations or only after many generations. Alternatively, 

it may never be expressed because of failure to reproduce or failure of 

the chance to reproduce. 

EPA treats genetic risk as independent of somatic risk even though 

somatic risk may be caused by mutations in somatic cells because, whereas 

somatic risk is expressed in the person exposed, genetic risk is 

expressed only in progeny and, in general, over many subsequent 

generations. Moreover, the types of damage incurred often differ in kind 

from cancer and cancer death. Historically, research on genetic effects 

and development of risk estimates have proceeded independently of the 

research on carcinogenesis. Neither the dose response models nor the 

risk estimates of genetic harm are derived from data on studies of 

carcinogenesis. 

Although genetic effects may vary greatly in severity, the genetic 

risks considered by the Agency in evaluating the hazard of radiation 

exposure include only those "disorders and traits that cause a serious 

handicap at some time during lifetime" (NAS80). Genetic risk may result 

from one of several types of damage that ionizing radiation can cause in 

the DNA within eggs and sperm. The types of damage usually considered 

are: dominant and recessive mutations in autosomal chromosomes, 

mutations in sex-linked (x-linked) chromosomes, chromosome aberrations 

(physical rearrangement or removal of part of the genetic message on the 

chromosome or abnormal numbers of chromosomes), and irregularly inherited 

disorders (genetic conditions with complex causes, constitutional and 

degenerative diseases, etc.). 

Estimates of the genetic risk per generation are conventionally 

based on a 30-yr reproductive generation. That is, the median parental 

age for production of children is defined as age 30 (one-half the 

children are produced by persons less than age 30, the other half by 

persons over age 30). Thus, the radiation dose accumulated up to age 30 

is used to estimate the genetic risks. EPA assessment of risks of 
genetic effects includes both first generation estimates and total 

genetic burden estimates. 

In the EPA Background Information Document for Radionuclides 

(EPA84), direct and indirect methods for obtaining genetic risk 

coefficients are described, and some recent estimates based on these 

methods are tabulated. Briefly, the direct method takes the frequency of 

mutation or occurrence of a heritable defect per unit exposure observed 

in animal studies and extrapolates to what is expected for humans. 

Direct estimates are usually used for first generation effects estimates. 

The indirect method, on the other hand, uses animal data in a different 

way. The estimated human spontaneous mutation rate per gene site is 

divided by the average radiation-induced mutation rate per gene observed 

in mouse studies, to obtain the relative radiation mutation risk in 

humans. The inverse of this relative radiation mutation risk is the 

expected "doubling dose" for radiation-induced mutations in man. The 

doubling dose is the exposure in rads which will double the current 

genetic malformation level in man and usually is used to estimate 

equilibrium effects or all future generation effects. 

A doubling dose estimate assumes that the total population of both 

sexes is equally irradiated, as occurs from background radiation, and 

that the population exposed is large enough so that all genetic damage 

can be expressed in future offspring. Although it is basically an 

estimate of the total genetic burden across all future generations, it 

can also provide an estimate of effects that occur in the first 

generation. Usually a fraction of the total genetic burden for each type 

of damage is assigned to the first generation using population genetics 

data as a basis to determine the fraction. For example, the BEIR 111 

Committee geneticists estimated that one-sixth of the total genetic 

burden of x-linked mutations would be expressed in the first generation 

and five-sixths across all subsequent generations. EPA assessment of 

risks of genetic effects includes both first generation estimates and 

total genetic burden estimates. 

6.6.2 Estimates of Genetic Harm Resulting from Low-LET Radiations 

A number of committees have addressed the question of genetic risk 

coefficient (NAS72,80; UNSCEAR 58,62,66,72,77,82; Of80) . The detailed 

estimates of the BEIR I11 Committee (NAS80) are listed in Table 6-17, and 

a summary of estimates of the various committees is listed in Table 6-18. 

Although all of the reports cited above used somewhat different 

sources of information, there is reasonable agreement in the estimates. 

However, all these estimates have a a considerable margin of error, both 

inherent in the original observations and in the extrapolations from 

experimental species to man. Some of the committee reports assessing the 

situation have attempted to indicate the range of uncertainty; others 

have simply used a central estimate (see Table 6-18). The same 

uncertainties exist for the latter (central estimates) as for the former. 

Most of the difference is caused by the newer information used in 

each report. Note that all of these estimates are based on the 
extrapolation of animal data to humans. Groups differ in their 

interpretation of how genetic experiments in animals might be expressed 

in humans. While there are no comparable human data at present, 

information on hereditary defects among the children of A-bomb survivors 

provides a degree of confidence that the animal data do not lead to 

underestimates of the genetic risk following exposure to humans. (See 

"Observations on Human Populations," which follows.) 

It should be noted that the genetic risk estimates summarized in 

Table 6-18 are for low-LET, low-dose, and low-dose-rate irradiation. 

Much of the data was obtained from high dose rate studies, and most 

authors have used a sex-averaged factor of 0.3 to correct for the change 

from high-dose rate, low-LET to low dose rate, low-LET exposure (NAS72, 

80, UNSCEAR72,77). However, factors of 0.5 to 0.1 have also been used in 

estimates of specific types of genetic damage (UNSCEAR72,77,82). 

Studies with the beta-particle-emitting isotopes carbon-14 and 

tritium yielded RBEs of 1.0 and 0.7 to about 2.0, respectively, in 

comparison to high-dose rate, high-dose exposure to x-rays 

(UNSCEAR82). At present, the RBE for genetic endpoints due to beta 

particles is taken as 1 (UNSCEAR77,82). 

6.6.3 Estimates of Genetic Harm from High-LET Radiations 

Although genetic risk estimates are made for low-LET radiation, 

some radioactive elements, deposited in the ovary or testis, can 

irradiate the germ cells with alpha particles. The relative 

biological effectiveness (RBE) of high-LET radiation, such as alpha 

particles, is defined as the ratio of the dose (rad) of low-LET 

radiation to the dose of high-LET radiation producing the same 

specific patho-physiological endpoint. 

In the Background Information Document for Radionuclides 

(EPA84), an RBE of 20 was assigned to high-LET radiation when 

estimating genetic effects. It was noted that studies comparing 

cytogenetic endpoints after chronic low-dose-rate gamma radiation 

exposure, or incorporation of plutonium-239 in the mouse testis, 

have yielded RBEs of 23 to 50 for the type of genetic injury 

(reciprocal translocations) that might be transmitted to liveborn 

offspring (NAS80, UNSCEAR77,82). Neutron RBE, determined from 

cytogenetic studies in mice, also ranged from about 4 to 50 

(UNSCEAR82, Gr83a, Ga82). However, an RBE of 4 for plutonium-239 

compared to chronic gamma radiation was reported for specific locus 

mutations observed in neonate mice (NASBO). 

Most recently, the NAS BEIR IV Committee reviewed the effects of 

alpha-emitting radionuclides and estimated the genetic effects (See 

Table 6-19). The BEIR IV genetic risk estimates for alpha-emitters 

were based on the low-LET estimates given in Table IV-2 in the 1980 

BEIR I11 report, applying an RBE of 15 for chromosome aberrations 

and 2.5 for all other effects. 
Table 6-17. BEIR 111 estimates of genetic effects of an average 

population exposure of 1 rem per 30-yr generation 

(chronic x-ray or gamma radiation exposure). 

Type of genetic Current incidence Effects per lo6 liveborn 

disorder per 10' liveborn per rem per generation 

First generation* Equilibrium** 

Autosomal dominant 

and x-linked 10,000 5-65 40-200 

Irregularly inherited 90,000 (not estimated) 20-900 

Recessive Very few Very slow 

increase 

Chromosomal aberrations 6,000 Fewer than 10 Increases 

only 

- slightly 

Total 107,100 5-75 60-1100 

* First-generation effects estimates are reduced from acute fractionated 

exposure estimates by a factor of 3 for dose rate effects and 1.9 for - fractionation effects (NAS80, p. 117). 

** Equilibrium effects estimates are based on low dose rate studies in mice 

(NAS80, pp. 109-110) . 

Source: NAS80. 
Table 6-18. Summary of genetic zisk estimates per lo6 liveborn of 

low-dose rate, low-LET radiation in a 30-yr generation. 

Source 

Serious hereditary effects 

First generation Equilibrium 

(all generations) 

BEAR, 1956 (NAS72) --- 

BEIR I, 1972 (NAS72) 49" (12-200) 

UNSCEAR, 1972 (UNSCEAR72) 9" (6-15) 

UNSCEAR, 1977 (UNSCEAR77) 6 3 

ICRP, 1980 (Of80) 8 9 

BEIR 111, 1980 (NAS80) 19" (5-75) 

UNSCEAR, 1982 (UNSCEAR82) 22 

Numbers in parentheses are the range of estimates. 

" Geometric mean of the lower and upper bounds of the estimates 

The geometric mean of two numbers is the square root of their 

product. 
Table 6-19. Genetic risk estimates per 106 live-born for an average 

population exposure of 1 rad of high-LET radiation in a 

30-year generation. 

Serious Hereditary Effects 

First Generation Equilibrium 

(all generations) 

Range 28 - 298 

Geometric Mean 9 1 

- - 

Source: NAS88 

These risk estimates, to a first approximation, give an 

average RBE of about 2.7 relative to the BEIR I11 low-LET estimates. 

This is numerically similar to the dose rate effectiveness factor 

for high dose rate. Therefore, for simplicity, it would be possible 

to use the same genetic risk coefficients per rad of high dose-rate, 

low-LET and per rad of high-LET radiation. 

6.6.4 Uncertainty in Estimates of Radiogenic Harm 

Chromosomal damage and mutations have been demonstrated in 

cells in culture, in plants, in insects, and in mammals 

(UNSCEAR'72,77,82), and in peripheral blood lymphocytes of persons 

exposed to radiation (UNSCEAR82, Ev79, Po78). However, they cannot 

be used for predicting genetic risk in progeny of exposed persons. 

Some believe such changes to be a direct expression of damage 

analogous to that induced by radiation in germ cells. At least, 

aberrations in peripheral lymphocytes show that radiation-induced 

chromosome damage can occur in vivo in humans. 

Since human data are so sparse, they can be used only to 

develop upper bounds of some classes of genetic risks following 

radiation exposure. Most numerical genetic risk estimates are based 

on extrapolations from animal data. 

Data below (Table 6-20), collected by Van Buul (Va80), on 

induction of reciprocal translocations in spermatogonia in various 

species, indicate that animal-based estimates for this type of 

genetic effect may be within a factor of 4 of the human value. 

However, if there were no human data on this genetic injury in the 

majority of cases, assuming that animal results and human results 

would be similar would underestimate the risk in humans. 
Table 6-20. Radiation-induced reciprocal translocations in 

several species 

Species Translocations 

(lo-' per rad) 

Rhesus monkey 0.86 + 0.04 

Mouse 1.29 7 0.02 to 2.90 + 0.34 - Rabbit 1.48 T 0.13 

Guinea pig 0.91 7 0.10 

Marmoset 7.44 T 0.95 

Human 3.40 - 0.72 

A basic assumption in the doubling-dose method of estimation is 

that there is a proportionality between radiation-induced and 

spontaneous mutation rates. Some of the uncertainty was removed in 

the 1982 UNSCEAR report with the observation that in two-test 

systems (fruit flies and bacteria), there is a proportionality 

between spontaneous and induced mutation rates at a number of 

individual gene sites. There is still some question as to whether 

or not the sites that have been examined are representative of ail sites and a11 gene loci, with developing evidence 

that the mouse 7-locus system is more sensitive to radiation than 

other members of the mouse genome (Ne88). The doubling-dose 

estimate, however, seems better supported than the direct estimate. 

There is some uncertainty as to which hereditary conditions 

would be doubled by a doubling dose; future studies on genetic 

conditions and. diseases can apparently, only increase the total 

number of such conditions. Every report, from the 1972 BEIR and 

UNSCEAR reports to the most recent, has listed an increased number 

of conditions and diseases that have a genetic component and hence 

may be increased by exposure to ionizing radiations. 

6.6.4.1 Observations on Human Populations 

A study of the birth cohort consisting of children of the 

Japanese A-bomb survivors was initiated in mid-1946. In a detailed 

monograph, Nee1 and Schull (Ne56) outlined the background of this 

first study and made a detailed analysis of the findings to January 

1954 when the study terminated. The study was designed to 

determine: (1) if during the first year of life, any differences 

could be observed in children born to exposed parents when compared 

to children born to suitable control parents, and (2) if differences 

existed, how they should be interpreted (Ne56). 
'This study addressed a number of endpoints, inclridirlg sex rati.0, 

malformations, perinatal data, and anthropometric data; subsequent 

studies have addressed other endpoints. Recent reports on this 

birth cohort of 70,082 persons have reported data on si.x endpoints. 

Frequency of stillbirths, major congenital defects, prenatal death, 

and frequency of death prior to age 17 have been examined in the 

entire cohort. Frequency of cytogenetic aberrations (sex chromosome 

aneuploidy) and frequency of biochemical variants (a variant enzyme 

or protein electrophoresis pattern) have been measured on large 

subsets of this cohort. 

There were small but statistically insignificant differences 

between the number of effects in the children of the proximally and 

distally exposed with respect to these various indicators. These 

differences are in the direction of the hypothesis that mutations 

were produced by the parental exposure. Taking these differences 

then as the point of departure for an estimate of the human doubling 

dose, an estimated doubling dose for low-LET radiation at high doses 

and dose rates for human genetic effects of about 156 rem (Sc81) or 

250 rem (Sa82) was obtained as an unweighted average. When each 

individual estlmate was weighted by the inverse of its variance, an 

average of 139 rem was found (Sc84). Because of the assumptions 

necessary for these calculations, as well as the Inherent 

statistical errors, the errors associated with these estimates are 

rather large. As a result, a reasonable lower bound to the human 

estimate overlaps much of the range based on extrapolation from 

mouse data. 

The most recent report evaluated the following possible genetic 

effects: (1) .untoward pregnancy outcomes, (2) all causes of early 

mortality, (3) balanced chromosomal exchanges, (4) sex-chromosome 

aneuploids, (5) early onset cancer, and (6) protein mutations. On 

the basis of the findings of the study, the authors concluded that 

the gametic doubling dose measured in humans for acute penetrating 

radiation exposure from atomic bombs is 150 rem to 190 rem (Ne88). 

The EPA is using the geometric mean of the BEIR 111 range of 

doubling doses: about 110 rads. EPA believes thLs estimate of 

doubling dose probably overstates the risk; however, it is 

compatible with both human and mouse data and should not be changed 

ac this time. EPA estimates of genetic risks will be reviewed and 

revised, if necessary, when more complete reports on the Japanese Abomb survivors are published. 

6.6.4.2 Ranges of Estimates Provided by Various Models 

Following recommendations of the 1980 BEIR I11 and earlier 

committees, EPA has continued to use a linear nonthreshold model for 

estimating genetic effects, although some data on specific genetic 

- endpoints obtained with acute low-LET exposures are equally well 

described by a linear-quadratic function. Moreover, in some of 

these cases, it has been found that a reduction in dose rate (or 

fractionation of dose) produced a reductjon in the quadratic term 
seen at high doses with little or no effect on the linear component. 

Such observations can be quali-tatively explained, as previously 

discussed in reference to somatic effects (Section 6.2.21, in terms 

of the dual radiation action theory of Kellerer and Rossi (Ke72), as 

well as alternative theories, e.g., one involving enzyme saturation 

(Go80, Ru58). 

Even though genetic risk estimates made by different committees 

based on the linear non-threshold model vary, the agreement is 

reasonably good. Some of the committees made estimates in terms of 

a range. These ranges are expressed as a single value by taking the 

geometric mean of the range. This method was recommended and first 

used by UNSCEAR (UNSCEAR58) for purposes of expressing genetic risk 

estimates. While the authors of the reports used different animal 

models, interpreted them in different ways, and had different 

estimates of the level of human genetic conditions in the 

population, the range of risk coefficients is about an order of 

magnitude (see Table 6-18). For the most recent, more comparable 

estimates, the range is a factor of 2 to 4 (see ICRP, BEIR 111, and 

UNSCEAR 1982 in Table 6-18) . 

6.6.5 The EPA Genetic Risk Estimates 

EPA has used the estimates from BEIR 111 (NAS80) based on a 

"doubling dose" range with a lower bound of 50 rem and an upper 

bound of 250 rem. The reasons are as follows: mutation rates for 

all gene loci affected by ionizing radiation are not known nor have 

all loci associated with "serious" genetic conditions been 

identified. Because the risk estimated by the direct method is 

incomplete, even for the subject animal species, and does not 

include the same types of damage estimated by doubling doses, EPA 

does not consider it further. Moreover, the BEIR 111 genetic risk 

estimates provide a better estimate of uncertainty than the UNSCEAR 

1982 and ICRP estimates because the BEIR I11 Committee assigned a 

range of uncertainty for multifactorial diseases (1 5 percent to < 

50 percent) that reflects the uncertainty in the numbers better than 

the other estimates (5 percent and 10 percent, respectively). 

The BEIR 111 estimates for low-LET radiations give a 

considerable range. To express the range as a single estimate, the 

geometric mean of the range is used, a method first recommended by 

UNSCEAR (UNSCEAR58) for purposes of calculating genetic risk. The 

factor of 3 increase in risk for high-dose rate, low-LET radiation, 

noted earlier, is also used. The weighted RBE for high-LET 

radiation as estimated in BEIR IV is about 3, which is numerically 

the same as the dose rate factor noted above. 

Genetic risk estimates used by EPA for high- and low-LET 

radiations are listed in Table 6-23.. As noted above (Section 

6.6.1), EPA uses the dose received before age 30 in assessing 

genetic risks. 
The EPA estimates in Table 6-21 are limited, like all other 

human genetic risk estimates, by the lack of confirming evidence of 

genetic effects in humans. These estimates depend on a presumed 

resemblance of radiation effects in animals to those in humans. The 

largest human source of data, the Japanese A-bomb survivors, appears 

at best to provide an estimate of the doubling dose for calculating 

the genetic risk in man which is not statistical.1y significant 

(Ne88) . 

In developing the average mutation rate for the two sexes 

used in the calculation of the relative mutation risk, the BEIR 

I11 Committee postulated that the induced mutation rate in 

females was about 40 percent of that in males (NAS80). Studies 

by Dobson, et al., show that the basis for the assumption was 

invalid and that human oocytes should have a risk equivalent to 

that of human spermatogonia. This would increase the risk 

estimate obtained from doubling-dose methods by a factor of 1.43 - 

Table 6-21. Estimated frequency of genetic disorders in a birth 

- cohort due to exposure of the parents to 1 rad per . generation. 

Serious heritablg disorders - 

(Cases per 10 liveborn) 

. 

Radiation First generation All generations 

Low Dose Rate, 

Low-LET 20 

High Dose Rate, 

Low-LET 60 

(Do83, Do84, D088) . Recently Dobson et al. (Do88) have shown that 

mouse oocytes are very sensitive to radiation, doses of 4 to 12 rads 

killing 50 percent of the immature mouse oocytes. Immature oocytes 

in women are not so easily killed. Dobson et al. (Do88) have also 

shown the existence of a special, hypersensitive, non-DNA lethality 

target (apparently the plasma membrane) in immature mouse oocytes. 

Irradiation with low energy neutrons, whose recoil protons have 

track lengths less than a cell diameter, induces genetic effects in 

immature mouse oocytes and yields effec~s similar to those observed 

in other cells (Do88). Immature human oocytes do not have the same 

hypersensitive target as mouse oocytes and so should be as 

- susceptible as spermatogonia to genetic effects of radiation. 

Unfortunately, BEIR 111 and, since it is based on BGIR 111, BEIR 

IV have embedded sex-sensit.iv3.t~ differences in their risk 
estimates. In BEIR 111: (1) autosomal dominants and X-iinked 

effects are based on a lower estimate where the oocyte has zero 

sensitivity and an upper estimate where the oocyte is 44 percent as 

sensitive as spermatogonia (p. 118); (2) irregularly inherited 

effects are based on an estimate where the oocyte is 44 percent as 

sensitive as spermatogonia (pp. 114 and 110); and (3) chromosomal 

aberrations estimates are based on oocytes and spermatogonia of 

equal sensitivity (p. 123, NASBO) . 

Since the sex-specific differences are in both BEIR I11 and BEIR 

IV, no attempt is made at this time to correct them. After BEIR V 

is published, EPA's genetic risk estimates will be reviewed and may 

then be revised. 

The combined uncertainties in doubling-dose estimates and the 

magnitude of genetic contributions to various disorders probably 

introduce an overall uncertainty of about an order of magnitude in 

the risk estimates. Moreover, the BEIR Committee, in deriving its 

estimate, has assumed that almost all of the risk was due to 

irregularly inherited mutations which would be eliminated slowly. 

They may include mild mutations which are but slightly detrimental 

in their heterozygous state. However, they may be sustained by 

advances in medical science, thus persisting and accumulating for 

generations. To what extent this occurs will depend on medical 

practices in the future. 

6.6.6 Effects of Multigeneration Exposures 

As noted earlier, while the somatic effects (cancer) occur in 

persons exposed to ionizing radiation, the genetic effects occur in 

progeny, perhaps generations later. The number of effects appearing 

in the first generation is based on direct estimates of the 

mutations induced by irradiation and should not change appreciably 

regardless of the background or "spontaneous" mutation rate in the 

exposed population. The estimate for total genetic effects, or the 

equilibrium estimate, is based on the doubling-dose concept. For 

these estimates, the background mutation rate is important: it is 

the background rate that is being "doubled." 

If there is long-lived environmental contamination, such that 30 

generations or more are exposed (N1000 years), the background 

mutation rate will change and come into equilibrium with the new 

level of radiation background. There will be an accumulation of new 

radiation-induced mutations until the background mutation rate has 

reached equilibrium with this continued insult. 

While predicting 1,000 years in the future is chancy at best, if 

it is assumed that there are no medical advances, and no changes in 

man or his environment, then an estimate can be made. In Table 6- 

- 21, it is estimated that exposure to 1 rad per generation of lowdose-rate, low-LET radiation will induce 260 cases of serious 

heritable disorders per lo6 live births in all generations. This is 

for a background mutation rate leading to 29,120 cases of serious 
heritable disorders per 106 Live births. The "all generations'" 

estimate in Table 6-21 is equal to the "equilibrium" estimate in 

Table 6-19. The "all generations" estimate is used for exposures to 

a single generation; the same number is employed as the 

"equilibrium" estimate for multigeneration exposures (see NAS80, 

p. 126, note 16). Thus, the risk estimate can be re-expressed as an 

estimate of the effects expected for a given change in the level of 

background radiation (Table 6-22). Since these calculations are 

based both on the background level mutations and the doubling dose, 

changes in either must be reflected in new calculations. 

Table 6-22. Increase in background level of genetic effects 

after 30 generations or more. 

Increase in Increase in serious heritable 

background disorders per lob live births 

radiation Low-dose-rate, High-LET 

(mrad/y) low-LET radiation radiation 

6.6.7 Uncertainties in Risk Estimates for Radiogenic Genetic 

Effects 

As noted throughout the preceding sections, there are sources 

of uncertainty in the genetic risk estimates. The overall 

uncertainty can be addressed only in a semi-quantitative manner. 

The identified sources of uncertainty are listed in Table 6-23. 

Uncertainties listed in this table are likely to be independent of 

each other and therefore unlikely to be correlated in sign. 

Although the root mean square sum of the numerical uncertainties 

suggests the true risk could be a factor of 4 higher or lower 

(x/t) by a factor of 4, it is unlikely, in light of the Japanese Abomb survivor data, that the upper bound is correct. 

6.6.8 Teratogenic Effects 

Although human teratogenesis (congenital abnormalities or 

defects) associated with x-ray exposure has a long history, the 

early literature deals mostly with case reports. (St21, Mu29, Go29). 

However, the irradiation exposures were high. 
Table 6-23. Causes of uncertainty in the genetic 

risk estimates. 

Degree of Uncertainty 

Source of Uncertainty in Risk Estimates 

Selection of species to use in 

developing a direct estimate x/+ factor of 4 

Selection of species and loci to 

use in developing a doubling dose -100% to estimate 

+indeterminate ''I 

Use of - division by a factor of 3 - to convert acute, high dose, low-LET 

estimates to chronic, low-LET estimates x/+ factor of 3 

Sensitivity of oogonia compared to 

spermatogonia as described in BEIR-I11 -44%Ib' + 56%"' 

Background rate selected for use 

with a doubling dose x/+, indeterminate 

Selection of RBE for high-LET 

radiation compared to an RBE of 20 x/+ a factor of 5 

Underestimate of the doubling dose 

required in man x/+ a factor of 2Id' 

la) The risk estimate cannot go below zero, -100%; but it may not 

be possible to determine the upper bound, indeterminate. 

(bl In reference to low estimate in Table 6-15 

(El In reference to high estimate in Table 6-15. 

Id) If the most recent analysis of the Japanese A-bomb survivors 

is correct, the lower bound for an estimate of the doubling 

dose in man is at least 2 times greater than the doubling dose 

estimate derived from the mouse. 
In 1930, Murphy exposed rats to x-rays at doses of 200 R to 

1,600 R. Of 120 exposed females, 34 had litters, and five of the 

litters had animals with developmental defects (Mu30). He felt that 

this study confirmed his clinical observations and earlier reports 

of animal studies. Although there were additional studies of 

radiation-induced mammalian teratogenesis before 1950, the majority 

of the studies were done after that time (see Ru53 for a review), 

perhaps reflecting concerns about radiation hazards caused by the 

explosion of nuclear weapons in 1945 (Ja70). 

Much of the work done after World War I1 used mice (Ru50, 

Ru54, Ru56) or rats (Wi54, Hi54). Early studies, at relatively high 

radiation exposures, 25 R and above, established some dose-response 

relationships. More important, they established the timetable of 

sensitivity of the developing rodent embryo and fetus to radiation 

effects (Ru54, Hi53, Se69, Hi66). 

Rugh, in his review of radiation teratogenesis (Ru70), listed 

the reported mammalian anomalies and the exposures causing them. 

The lowest reported exposure was 12.5 R for structural defects and 1 

R for functional defects. He also suggested human exposure between 

ovulation and about 7 weeks gestational age could lead to structural 

defects, and exposures from about 6 weeks gestational age until 

birth could lead to functional defects. In a later review (Ru71), 

Rugh suggested structural defects in the skeleton might be induced 

as late as the 10th week of gestation and functional defects as 

early as the 4th week. It should be noted that the gestation period 

in mice is much shorter than that in humans and that weeks of 

gestation referred to above are in terms of equivalent stages of 

mouse-human development. However, estimates of equivalent 

gestational age are not very accurate. 

Rugh (Ru71) suggested there may be no threshold for radiationinduced congenital effects in the early human fetus. In the case of 

human microcephaly (small head size) and mental retardation, at 

least, some data support this theory (Ot83, 84). 

However, for most teratogenic effects, the dose response at low 

doses is not known. In 1978, Michel and Fritz-Niggli (Mi78) 

reported induction of a significant increase in growth retardation, 

eye and nervous system abnormalities, and post-implantation losses 

in mice exposed to 1 R. The increase was still greater if there was 

concurrent exposure to radiosensitizing chemicals such as 

iodoacetimide or tetracycline (Mi78). 

In other reports of animal studies, it appeared as if 

teratologic effects, other than perhaps growth retardation, had a 

threshold for induction of effects (Ru54, Ru53, Wi54). However, 

Ohzu (Oh65) showed that doses as low as 5 R to preimplantation mouse - embryos caused increased resorption of implanted embryos and 

structural abnormalities in survivors. Then in 1970, Jacobsen 

(Ja70) reported a study in which mice were exposed to 5, 20, or 100 

R on the eighth day of pregnancy. He concluded that the dose 
response function for induction of skeletal effects was Linear, or 

nearly linear, with no observable threshold. This appears 

consistent with a report by Russell (Ru57), which suggested a 

threshold for some effects whereas others appeared to be linearly 

proportional to dose. 

One of the problems with the teratologic studies in animals is 

the difficulty of determining how dose response data should be 

interpreted. Russell (Ru54) pointed out some aspects of the 

problem: (1) although radiation is absorbed throughout the embryo, 

it causes selective damage that is consistently dependent on the 

stage of embryonic development at the time of irradiation, and 

(2) the damaged parts respond, in a consistent manner, within a 

narrow time range. However, while low-dose irradiation at a certain 

stage of development produces changes only in those tissues and 

systems that are most sensitive at that time, higher doses may 

induce additional abnormalities in components that are most 

sensitive at other stages of development, and may further modify 

expression of the changes induced in parts of the embryo at maximum 

sensitivity during the time of irradiation. In the first case, 

damage may be to primordial cells themselves, while in the second, 

the damage may lead indirectly to the same or different endpoints. 

The human embryo/fetus starts as a single, fertilized egg and 

divides and differentiates to produce the normal infant at term. 

(The embryonic period, when organs develop, is the period from 

conception through 7 weeks gestational age. The fetal period, a 

time of -- in utero growth, is the period from 8 weeks gestational age 

to birth.) The different organ and tissue primordia develop 

independently and at different rates. However, they are in contact 

through chemical induction or evocation (Ar54). These chemical 

messages between cells are important in bringing about orderly 

development and the correct timing and fitting together of parts of 

organs or organisms. While radiation can disrupt this pattern, 

interpretation of the response may be difficult. Since the cells in 

the embryo/fetus differentiate, divide, and proliferate at different 

times during gestation and at different rates, gestational times 

when cells of specific organs or tissues reach maximum sensitivity 

to radiation are different. Each embryo/fetus has a different 

timetable. In fact, each half (left/right) of an embryo/fetus may 

have a slightly different timetable. 

In addition, there is a continuum of variation from the 

hypothetical normal to the extreme deviant which is obviously 

recognizable. There is no logical place to draw a line of 

separation between normal and abnormal. The distinction between 

minor variations of normal and frank malformation, therefore, is an 

arbitrary one, and each investigator must establish his or her own 

criteria and apply them to spontaneous and induced abnormalities 

alike (HWC73). 

The limitations of the human data available make the use of 

animals in both descriptive and experimental studies inevitable. 

However, this gives rise to speculation about the possible relevance 
of such studies to man. There are species differences in 

development attributable partly to the differj-ng complexity of the 

adu3.t organs, but especially to differences in growth :rates and 

timing of birth in relation to the developmental events. For 

example, the histological structure of the brain is, in general, 

surprisingly similar, both in composition and in function, from one 

mammalian species to another, and the sequence of events is also 

similar (Do73). However, the processes of brain development that 

occur from conception to about the second year of life in man are 

qualitatively similar to those seen in the rat during the first six 

weeks after conception (Do79, 81) . 

For example, a major landmark, the transition from the 

principal phase of multiplication of the neuronal precursors to that 

of glial multiplication, occurs shortly before mid-gestation in man, 

but at about the time of birth in the rat (Do73). In thls respect, 

then, the rat is much less neurologically mature at birth than the 

newborn human infant. Many other species are more mature at birth; 

the spectrum ranges from the late-maturing mouse and rat to the 

early-maturing guinea pig, with non-human primates much closer to 

the guinea pig than to man (Do79, 81). As a consequence, it is 

unreasonable to compare a newborn rat's brain, which has not begun 

to myelinate, with that of a newborn human which has, or with that 

of a newborn guinea pig in which myelination has been completed 

(Do79, 81). 

Nevertheless, in the study of teratogenic effects of prenatal 

exposure to ionizing radiation, in which the timing of the exposure 

in relation to the program of developmental events dictates the 

consequences of that insult, it is necessary only to apply the 

experimental exposure at the appropriate stage (rather than at a 

similar age) of embryonic or fetal development in any species to 

produce similar results in all (Do79, 81). The duration of exposure 

must, however, match the different time scales in the different 

species. Unless these elementary rules 

of cross-species adjustments are followed, extrapolation of even 

qualitative estimates of effects will be of dubious relevance and 

worth. 

Because of the problems in interpretation listed above, a 

pragmatic approach to evaluation of studies is useful. The dose 

response should be given as the simplest function that fits the data 

(often linear or linear with a threshold). No attempt should be 

made to develop complex dose response models unless the evidence is 

unequivocal. 

6.6.8.1 Teratologic Effects: Mental Retardation in Humans 

The first report of congenital abnormalities in children exposed 

-- in utero to radiation from atomic bombs was that of Plumer (Pl52). 

Twelve children with microcephaiy, of which ten also had mental 

retardation, had been identified in Hiroshima in a small set of the 

in .- utero exposed survivors. They were found as part of a program 
started in I.950 i.o st.udy chi]-dren exposed in the first. trimester of 

gestati.on. However, not all of the in utero exposed survivors were 

examined. In 1955, the program was expanded to j.ncl.ude all 

survivors exposed -- in utero. 

Studies initiated during the program have shown radiationrelated (I) growth retardation; (2) increased microcephaly; 

(3) increased mortality, especially i-nfant mortality; (4) temporary 

suppression of antibody production against influenza; and 

(5) increased frequency of chromosomal aberrations in peripheral 

lymphocytes (Ka.13). 

Although there have been a number of studies of Japanese 

A-bomb survivors, including one showing a dose- and gestational agerelated increase in postnatal mortality (Ka73), only the incidences 

of microcephaly and mental. retardation have been investigated to any 

great extent. In the most recent report, Otake and Schull (Ot83, 

84) showed that mental retardation was particularly associated with 

exposure between 8 and 15 weeks of gestation (10 to 17 weeks of 

gestation if counted from the last menstrual period). They further 

found the data suggested little, if any, non-linearity and were 

consistent with a linear dose-response relationship for induction of 

mental retardation that yielded a probability of occurrence of 

severe mental retardation of 4.16t0.4 cases per 1,000 live births 

per rad of exposure (Ot84). A child was classified as severely 

mentally retarded if he or she was "unable to perform simple 

calculations, to make simple conversation, to care for himself or 

herself, or if he or she was completely unmanageable or had been 

insti.tutionalized" (Ot83, 84) . There was, however, no evidence of 

an effect in those exposed at 0 to 7 weeks of gestation (Ot83). 

Exposure at 16 weeks or more of gestation was about a factor of 4 

less effective, with only a weak relationship between exposure and 

risk, and with few cases below 50 rads exposure (Ot84). 

Mental retardation can be classified as mild (IQ 50-70), 

moderate (IQ 35-49), severe (IQ 20-341, and profound (IQ < 20) 

(WH075). However, some investigators use only mild mental 

retardation (IQ 50-70) and severe mental retardation (IQ < 50) as 

classes (Gu77b, HaBla, St84). Mental retardation is not usual1.y 

diagnosed at birth but at some later time, often at school age. 

Since the mental retardation may have been caused before or during 

gestation, at the time of birth, or at some time after birth, that 

fraction caused before or during gestation must be estimated. In 

like manner, since mental retardation caused before birth may be due 

to genetic conditions, infections, physiologic conditions, etc., the 

fraction related to unknown causes during gestation must be 

estimated. Thi-s is the fraction that might possibly be related to 

radiation exposure. 

Estimates of the risk of mental retardation for a rad of 

embryo/fetus exposure in the U.S. population can be derived using 

the absolute risk calculated by Otake and Schull for the Japanese 

survivors (0t84). Otake and Schull (0t84) gave an estiinate foi: one 
case entitled, "The Relationship of Mental Retardation to Absorbed 

Fetal Exposure in the Tensitive' Period When A11 'Contro1s"re 

Combined." This estimate oE frequency of mental retardation, 0.416 

per LOO rads, could be directly applicable to a U.S. population. In 

this case, the risk estimate would be about four cases of severe 

mental retardation per 1,000 live births per rad of exposure during 

the 8th and 15th week of gestation. 

The ICRP published an excellent review of biology and possible 

mechanisms of occurrence of radiation-induced brain damage, -- in utero 

(ICRP86). ICRP estimates: (1) for exposures from the 8th through 

the 15th week after conception, the risk of severe mental 

retardation is 4 x per rad, with a confidence interval of 2.5 x 

to 5.5 x lom3, and (2) for exposures from the 16th through the 

25th week after conception, the risk of severe mental retardation is 

1 x per rad. However, a threshold below 50 rad cannot be 

excluded (ICRP86). 

If the estimate is applicable, the low-LET background radiation . 

- (about 15 mrads) delivered during the 8- to 15-week gestational agesensitive period could induce a risk of 6 x 10" cases of severe 

mental retardation per live birth. This can be compared to an 

estimate of a spontaneous occurrence of 0.6 x 10" to 3.1 x cases - 

of idiopathic severe mental retardation per live birth (EPA84). 

- 6.6.8.2 Teratologic Effects: Microcephaly in Humans 

Plumrner (P152) reported microcephaly associated with mental 

retardation in Japanese A-bomb survivors exposed -- in utero. Wood 

(Wo65, 66) reported both were increased. The diagnosis of reduced 

head circumference was based on "normal distribution" statistical 

theory (Wo66); i.e., in a population, the probability of having a 

given head circumference is expected to be normally distributed 

around the mean head circumference for that population. 

For example, in a population of live-born children, 2.275 

percent will have a head circumference 2 standard deviations or more 

smaller than the mean, 0.621 percent will have a head circumference 

2.5 standard deviations or more smaller than the mean, and 0.135 

percent will have a head circumference 3 standard 

deviations or more smaller than the mean (statistical estimates 

based on a normal distribution). 

For most of the studies of the Japanese A-bomb survivors exposed 

-- in utero, if the head circumference was two or more standard 

deviations smaller than the mean for the appropriate controls in the 

unexposed population, the case was classified as having reduced head 

circumference even if the data had not been adjusted for differences 

in stature (Ta67, Mi72, Wo65). While a definitive relationship 

between reduced head circumference and mental retardation has not 

been established, there is evidence that they are related. 
Studies of the Japanese survivors show a re1ationshj.p between 

reduced head size and mental retardation, but all these studies are 

based on subsets of the total -- in utero population. The fraction of 

mentally retarded with reduced head circumference has been reported 

as 50 percent (RERF78) to 70 percent (Wo66), while the fraction of 

those selected for reduced head circumference who had mental 

retardation has been reported as 11 percent (Wo66) to 22 percent 

(Mi72). Thus, while the relationship appears to exist, it has not 

been quantified. 

The majority of the cases of reduced head size are observed in 

those exposed in the first trimester of gestation, particularly the 

6th or 7th to 15th weeks of gestation (Mi59, Wo66, Mi72, Wo65, 

Ta67). Most recently, it has been shown that reduction in head 

circumference was a linear function of dose (1~84). However, the 

authors noted that the analysis was based on T65 dosimetry, and the 

data should be reanalyzed after completion of the dosimetry 

reassessment currently in progress. 

These findings of reduction in head circumference, with a window 

of effect in the same time period of gestation as mental 

retardation, help support the observations on mental retardation. 

Although the exact dose response functions are still uncertain, data 

on both types of effects have so far been consistent with a linear, 

no-threshold dose response during the critical period. 

6.6.8.3 Other Teratologic Effects 

Japanese A-bomb survivors exposed -- in utero also showed a number 

of structural abnormalities and, particularly in those who were 

microcephalic, retarded growth (Wo65). No estimate has been made of 

the radiation-related incidence or dose-response relationships for 

these abnormalities. However, UNSCEAR (UNSCEAR77) made a very 

tentative estimate based on animal studies that the increased 

incidence of structural abnormalities in animals may be 0.005 cases 

per R per live born, but stated that projection to humans was 

unwarranted. In 1986, UNSCEAR assumed the risk of an absolute 

increase of malformed fetuses of the order of 5E-3 per rad seen in 

animals might apply to the human species as well, for exposure over 

the period from 2 to 8 weeks post-conception (UNSCEAR86). In any 

event, the available human data cannot show whether the risk 

estimates derived from high-dose animal data overestimate the risk 

in humans or if a threshold can be excluded. 

It should be noted that all of the above estimates are based on 

high-dose-rate, low-LET exposure. In 1977, UNSCEAR also 

investigated the dose rate question and stated: 

"In conclusion, the majority of the data available for most 

species indicate a decrease of the cellular and malformature 

effects by lowering the dose rate or by fractionating the 

dose. However, deviations from this trend have been well 

documented in a few instances and are not inconsistent with 
the knowledge about mechanisms of the teratogenic effects. 

It is therefore impossible to assume that dose rate and 

fractionation factors have the same influence on all 

teratological effects." (UNSCEAR77). 

6.6.9 Nonstochastic Effects 

Nonstochastic effects, those effects that increase in severity 

with increasing dose and have a threshold, have been reviewed in the 

1982 UNSCEAR report (UNSCEAR82). Nonstochastic effects following in 

utero exposure were reviewed in the 1986 UNSCEAR report (UNSCEAR~~~ 

In general, acute doses of 10 rads low-LET radiation and higher are 

required to induce these effects in animals. It is possible that 

some of the observed effects of in utero exposure are nonstochastic: 

e.g., the risk of embryonic loss, estimated to be 10.' per R 

(UNSCEAR77) or per rad (UNSCEAR86) following radiation exposure soon 

after fertilization. However, there are no data to address the 

question of similar effects in humans. Usually, nonstochastic 

effects are not expected at environmental levels of radiation 

exposure. 

In 1986, the United Nations Scientific Committee on the Effects 

of Atomic Radiation also reviewed the question of mental retardation 

as a part of the overall review of the biological effects of 

prenatal radiation exposure (UNSCEAR86). UNSCEAR, like the ICRP, 

concluded there was a risk of severe mental retardation of 4 x lo-' 

per rad over the period of 8 to 15 weeks after conception and of 1 x 

per rad over the period 16-25 weeks after conception 

(UNSCEAR86). UNSCEAR also estimated (1) a pre-implantation loss of 

1 x per rad during the first two weeks after conception, (2) a 

malformation risk of 5 x lo-' per rad during weeks 2 to 8 after 

conception, and (3) a risk of leukemia and solid tumors expressed 

during the first 10 years of life of 2 x per rad (UNSCEAR86). 

EPA has adopted similar conclusions for estimating prenatal 

carcinogenic, teratologic, and nonstochastic effects in man (Table 

6-24) . 

6.7 Summary of EPA's Radiation Risk Factors - A Perspective 

Table 6-25 summarizes EPAts estimate of risk from lifetime 

whole-body exposures to high- and low-LET radiation and to radon 

decay products. The nominal risk factors reflect EPA's best 

judgment as to the relationship between dose and risk based on 

review of all relevant information available to the Agency. 

Likewise the cited ranges reflect EPA's current best judgment as to 

the uncertainties in these risk factors. 
* * * DRAFT * * * 

Table 6-24. Possible effects of in utero radiation exposure 

Type of Risk Risk per Rad Risk per Event in a 

to Conceptus 100 mrad per Year 

Background 

Cancer 5.3 lo4 4.8 x lo-' 

Mental Retardation 4 

(exposure at 8 - 15 weeks) 

Mental Retardation 1 

(exposure at 16 - 25 weeks) 

Malformation 5 x 5.8 x 10.' 

(exposure at 2 - 8 weeks) 

Pre-implantation 1 x 10 -2 3.8 x 

Loss (exposure at 

0 - 2 weeks) 

To provide a perspective on the risk of fatal radiogenic cancers 

and the hereditary damage due to radiation, EPA has calculated the 

risk from background radiation to the U.S. population using the risk 

factors summarized in Table 6-24. The risk from background 

radiation provides a useful perspective for the risks caused by 

emissions of radionuclides. Unlike cigarette smoking, auto 

accidents, and other measures of common risks, the risks resulting 

from background radiation are neither voluntary nor the result of 

self-induced damage. The risk caused by background radiation is 

largely unavoidable; therefore, it is a good benchmark for judging 

the estimated risks from radionuclide emissions. Moreover, to the 

degree that the estimated risk of radionuclides is biased, the same 

bias is present in the risk estimates for background radiation. 

The absorbed dose rate from low-LET background radiation has 

three major components: cosmic radiation, which averages about 

28 mrad/yr in the United States; terrestrial sources, such as radium 

in soil, which contribute an average of 28 mrad/yr (NCRP87); and the 

low-LET dose resulting from internal emitters. The last differs 

among organs, to some extent, but for soft tissues it is about 

24 mrad/yr (NCRP87). Other minor radiation sources such as fallout 

from nuclear weapons tests, cosmogenic radionuclides, naturally 

occurring radioactive materials in buildings, airline travel, and 

consumer products, contribute about another 7 mrad for a total lowLET whole-body dose of about 87 mrad/yr. The lung and bone receive 

somewhat larger doses, not included in the 87 mrad/yr estimate, due 

to high-LET radiations (see below). Although extremes do occur, the 

distribution of this background annual dose to the U.S. population 

is relatively narrow. A population-weighted analysis indicates that 
KO percent of the U,S, popuiati.on wou1.d receive annual. doses that 

are between 75 rnrad/gr and 115 rnrad/yr (EPA8ll. 

As outlined in Section 6.2, the BEIR 111 linear, relative risk 

models yield, for 1-ifetirne exposure to low-LET radiation, an average 

lifetime risk of fatal radiogenic cancer of 3.9~10.~ per rad. Note 

that this average is for a group having the age- and sex-specific 

mortality rates of the 1.970 U.S. population. This risk estimate can 

be used to calculate the average lifetime risk due to low-LET 

background radiation as follows. The average duration of exposure 

in this group is 70.7 yr, and at 90 mrad/yr, the average lifetime 

dose is 6.4 rads. The risk of fatal cancer per person in this group 

is: 

or about 0.24 percent of all deaths. The vital statistics used in 

EPArs radiation risk analyses indicate that the probability of dying 

from cancer in the United States from all causes is about 0.16, 

i.e., 16 percent. Thus, the 0.24 percent result for the BEIR I11 

linear dose response model. indicates that about 1.5 

percent of all U.S. cancer is due to low-LET background radiation. 

The BEIR 111 linear-quadratic model indicates that about 0.1 percent 

of all deaths are due to low-LET background radiation or about 0.6 

percent of all cancer deaths. 

Table 6-11 indicates a risk of 5.6xlO-'rad-' for alpha emitters 

in lung tissue. UNSCEAR estimated that in "normal" areas the annual 

absorbed dose in the lungs from alpha emitters other than radon 

decay products would be about 0.51 mrad (UNSCEAR77). The individual 

lifetime cancer risk from this exposure is: 

which is about 1/100 of the risk due to low-LET background radiation 

calculated by means of the BEIR I11 linear model. 

The 1982 UNSCEAK report indicates that the average annual 

absorbed dose to the endosteal surfaces of bone due to naturally 

occurring, high-LET alpha radiation is about 6 mrad/yr, based on a 

quality factor of 20 and an absorbed dose equivalent of 120 mrem/yr 

(UNSCEAR82). Table 6-11 indicates that the individual lifetime risk 

of fatal bone cancer due to this portion of the naturally occurring 

radiation background is: 

The exposure due to naturally occurring background radon-222 

progeny in the indoor environment is not well known. The 1982 

UNSCEAR report lists for the United States an indoor concentration 

of about 0.004 worlcing levels (15 Bq/m3) (UNSCEAR82) . This estimate 

is not based on a national survey and is known to be exceeded by as 

much as a factor of 10 or more in some houses. However, as pointed 
Table 6-25. Summary of EPAQ radiation risk factors 

Risk Significant Risk. Factor 

Exposure Period Nominal Range 

Low LET (10-"ad-'i 

~eratolo~ical:~ 

Severe mental Weeks 8 to 15 4,000 

retardation of gestation 

Genetic : 

Severe hereditary 30 year 260 

defects, all reproductive 

generations generation 

Somatic : 

Fatal cancers Lifetime 

All cancers Lifetime 

Hiqh LET (10.' rad-') 

Genetic : 

Severe hereditary 30 year 690 

defects, a11 reproductive 

generations generation 

Somatic : 

Fatal cancers Lifetime 

All cancers Lifetime 

Radon decay products (10-"LM-') 

Fatal lung cancer Lifetime 360 160 - 720 

a The range assumes a linear, non-threshold dose response. However, it is 

plausible that a threshold may exist for this effect. 
out in UNSCEAR82, the national collective exposure may not be too 

dependent on exceptions to the mean concentration. The UNSCEAR 

estimate for the United States now appears low (Ne86); the average 

residential exposure is probably 0.2-0.3 WLM/yr (in standard 

exposure units). 

Assuming 0.25 WLM/yr is a reasonable estimate for indoor 

exposure to radon-222 progeny in this country, the mean lifetime 

exposure, indoors, is about 18 WLM. Based on the geometric mean 

lifetime risk coefficient from Section 6.4.5, 360 cases/106 WLM, a 

lifetime risk of 0.64 percent is estimated. For comparison, roughly 

5 percent of all deaths in 1980 were due to lung cancer. Based on 

these assumptions, therefore, about one of eight lung cancer deaths 

may be attributable to background radon exposure. This would 

correspond to about 4 percent of all cancer deaths. This is 2.5 

times the 1.61 percent of all cancer fatalities estimated above for 

low-LET background radiation. The reader is cautioned, however, 

that this risk estimate applies only to the United States population 

taken as a whole, i.e., men and women, smokers and nonsmokers. 

Since the vast majority of the 1980 lung cancer mortality occurred 

in male smokers, this risk estimate cannot be applied 

indiscriminately to women or nonsmokers (see Section 6.4). 

The spontaneous incidence of serious congenital and genetic 

abnormalities has been estimated to be about 105,000 per lo6 live 

births, about 10.5 percent of live births (NAS80, UNSCEAR82). The 

low-LET background radiation dose of about 87 mrad/year in soft 

tissue results in a genetically significant dose of 2.6 rads during 

the 30-year reproductive generation. Since this dose would have 

occurred in a large number of generations, the genetic effects of 

the radiation exposure are thought to be at an equilibrium level of 

expression. Since genetic risk estimates vary by a factor of 20 or 

more, EPA uses a log mean of this range to obtain an average value 

for estimating genetic risk. Based on this average value, the 

background radiation causes about 680 genetic effects per 10' live 

births (see Section 6.6). This result indicates that about 0.6 

percent of the current spontaneous incidence of serious congenital 

and genetic abnormalities may be due to the low-LET background 

radiation. 
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7. SUMMARY OF UNCERTAINTIES IN DOSES AND RISKS 

7.1 INTRODUCTION 

The purpose of this chapter is to provide insight into the 

uncertainty associated with the dose and risk estimates for the 

nearby individuals and the U.S. population for each of the twelve 

emission source categories. For each of these categories and their 

subcategories the calculation of doses and risks involves the 

following steps: 

o calculation of the radionuclide release rate to the 

environment, 

o calculation of atmospheric transport and depletion, 

o calculation of the radionuclide concentration 

in the environment accessible to the general 

population in the vicinity of the release, 

o calculation of the doses to the exposed individuals 

from the various external, ingestion, and inhalation 

pathways, and 

o calculation of the health risks due to the doses 

received. 

Accompanying the calculational methods described in Chapters 4, 

5, and 6 of this volume is a discussion of the uncertainties 

associated with each of these calculational steps: 

o The uncertainty in the average annual atmospheric 

dispersion factors for any given location can range from 

about a factor of 2 to 10, depending on distance from the 

release point and complexity of the release and terrain. 

For population exposures, uncertainty in the atmospheric 

dispersion factors is smaller due to the effects of 

averaging (see Section 4.2). Since dispersion modeling 

is applicable to all categories, radionuclides, and 

pathways of exposure, this parameter is an important 

source of uncertainty. 

o For particulate emissions, the rate of deposition of 

particulates at a given location can vary by over a 

factor of 10 (see Section 4.3). However, the total 

deposition over an extended time period or over a large 

area is less uncertain due to the effects of time and 

spacial averaging. In addition, since the principal risk 

from many source categories is due to inhalation, this 

source of uncertainty is not highly significant. 
o The uncertainty in the food chain transfer factors, 

which describe the uptake of radionuclides by 

vegetation from soil and subsequent transfer to meat, 

milk, and produce, is large and varies substantially 

from site to site, by pathway, and by radionuclide (see 

Section 4.4). However, direct deposition onto plant 

surfaces generally contributes more to the radionuclide 

concentrations in foods than does uptake from soil, which 

tends to reduce the importance of this source of 

uncertainty. 

0 The uncertainty in the dose conversion factors (DCFs), 

which relate exposure to dose, is small for external 

exposures but variable for internal exposures, depending 

on the radionuclide. Key sources of uncertainty in 

internal DCFs are (1) individual variability, (2) 

uncertainty in particle size and clearance class, and (3) 

uncertainties in the model formulation and the values of 

the parameters for a given particle size and clearance 

class (see Chapter 5). 

o The uncertainty in risk conversion factors, which 

relate dose or exposure to risk, is estimated to be a 

factor of about 3 for the average member of the 

population. For any one individual, the uncertainties 

cannot be meaningfully defined (see Chapter 6). 

The important contributors to uncertainty, and therefore the 

overall uncertainty in calculated doses and risks, can differ 

significantly among the release categories because of differences in 

the important radionuclides, pathways, and environmental settings 

among the different categories. 

A number of approaches may be used to describe and quantify 

uncertainty (Mo78, He79, Cr88, Sc80). One method to assess 

uncertainty is to assign probability distributions to each of the 

parameters and then to propagate the uncertainties through the dose 

and risk calculations. These approaches are currently being 

evaluated by EPA with the intent to apply them to the risk 

assessments; the results to be published as a supplement to this 

report. Appendix C to this volume presents a summary of the 

uncertainty analysis techniques under consideration. 

For the purpose of this report, a semiquantitative 

characterization of uncertainty is provided, whereby the uncertainty 

in the key calculational parameters for each category is described. 

No attempt is made to establish probability distributions for the 

parameters or to propagate rigorously the uncertainties. Instead, 

approximate ranges of doses and risks are provided. The ranges are 
based on inspection of the important contributors to risk and the 

discussion of uncertainties provided in Chapters 4, 5, and 6. This 

approach is appropriate when deterministic calculations are 

employed, and when explicit distributions for the input parameters 

cannot be readily developed. 

Throughout the following discussion, a distinction is made 

between the uncertainties in the individual and population doses and 

risks. This distinction is significant. By their very nature, 

population doses and risks have much less inherent uncertainty 

because the uncertainty is only in the average value of a given 

parameter, as opposed to the range of possible values for the 

parameter. For example, the soil to plant transfer factor (B,,, see 

Chapter 4) for any particular radionuclide can vary by several 

orders of magnitude depending on the soil conditions and the type of 

plant. These characteristics can vary widely within the vicinity of 

a facility, and, therefore, the Biv values could also vary widely. 

However, population doses are the sum of the individual doses at a 

site. Accordingly, it is the average value of the B, at the site 

which is needed to calculate population doses, and the individual 

differences among Biv values at specific locations at the site 

average out of the calculation. 

7.2 INACTIVE URANIUM MILLS 

The radiation exposures attributable to air emissions from each 

of the 24 inactive uranium mill tailing sites are due to the 

windblown resuspension of uranium series radionuclides, the emission 

of radon from the piles, and their subsequent transport to nearby 

members of the population. Numerous evaluations of these sources of 

exposure have demonstrated that the offsite exposures are dominated 

by radon and its progeny (DOE86). Accordingly, the uncertainty 

analysis is limited to radon exposures. 

Each of the 24 sites differs to varying degrees in the size 

and characteristics of the pile, meteorological conditions, and 

population density and distribution. Accordingly, the exposures and 

risks to the maximum individual and to the 80-km population differ 

substantially among the 24 sites. Due to this variability and the 

availability of site-specific information, the radon exposures were 

evaluated for each site. 

The uncertainty in the individual and population risks for each 

site arises from the uncertainties in (1) the average annual radon 

emanation rate from the pile, (2) the average annual atmospheric 

dispersion factors to the nearby residents, (3) the average resident 

occupancy time, and (4) the assumed risk factor for radon exposure. 

The following discusses the uncertainties in each of these 

calculational parameters as they apply to the maximum individual and 

population doses and risks. 
'7.2.1 Radon Emanation Rates 

The radon emanation rates from the piles are expressed in units 

of Ci/yr. This is a calculated value based on the area of the pile, 

the average Ka-226 concentration in the pile, and the assumption 

that the radon emanation rate is 1 pCi/m2-sec per pCi/g of Ra-226 in 

the tailings. The concentrations of Ra-226 in the piles have been 

extensively studied, and their uncertainty is not an important 

contributor to the overall uncertainty in the source term. 

The assumed emanation rate per unit area is based on empirical 

data which vary substantially as a function of time and location on 

the pile and numerous physical and chemical parameters (La87, Mo87, 

Ha86). However, since the quantity of interest is average annual 

release rate over the entire pile, the uncertainty is reduced due to 

averaging. 

For the five piles that have been stabilized, the radon 

emanation rate is assumed to be at the regulatory limit of 20 

pCi/m2-sec. This is a conservative assumption, since remedial 

designs typically exceed their design objectives. Post-remediation 

radon flux measurements have not been routinely made on the piles, 

and the actual flux for these piles may be comparable to natural 

background, i.e., 1-5 pCi/m2-sec. 

In conclusion, the average annual radon emanation rate for the 

limiting pile used in the calculations for the maximally exposed 

individual is realistic. For the population doses, there is a small 

uncertainty in the effective area of the piles. A great deal of the 

uncertainty typically associated with predicting radon emanation is 

eliminated because the parameters of interest are average or 

integrated values. For the stabilized piles, the radon emanation 

rate may be overestimated if design objectives are exceeded. 

7.2.2 Atmospheric Transport 

The product of the average annual source term with appropriate 

average annual atmospheric dispersion factors yields the average 

annual radionuclide concentrations at the locations of interest. 

The dispersion factors used in the risk assessments were obtained 

using site-specific meteorological data. These data are used as 

input to the standard Gaussian dispersion model for area sources. 

Gaussian dispersion modeling has been found to be reliable for 

near-field, average annual conditions in relatively flat terrain 

(Mu85). For the maximum individual doses, these conditions hold. 

For the 80-km population doses, these conditions do not hold due to 

the distances involved. However, since the population doses are 

integrated over space and time, the uncertainties average out, 

resulting in an acceptable estimate of integrated dose. 
In conclusion, though the uncertainties in calculating an 

atmospheric dispersion factor at a specific location and point in 

time may be large (Mi88), when applied to average annual conditions, 

as is the case for the maximum individual doses, or when integrated 

over time and space, as is the case for the population doses, the 

dispersion factors are more reliable. As discussed in Chapter 4, 

validation studies for Gaussian dispersion models indicate an 

uncertainty of approximately a factor of 2 for average annual 

concentrations for locations within 10 km of the release and a 

factor of 4-10 for locations between 30-140 km of a release. 

7.2.3 Location of the Maximum Individual 

One of the largest sources of uncertainty in calculating the 

maximum individual doses and risks is uncertainty in the location of 

the maximum individual and the duration of exposure. The location 

of the maximum individual was determined through site units. The 

Agency's ground-rule when estimating risk is that individuals remain 

at these locations throughout their lives. There is a certain 

degree of inherent uncertainty associated with this assumption; 

however, it ensures that those individuals that may reside at one 

location for their entire lives are given appropriate consideration. 

7.2.4 Risk Factor 

The risk factor for exposure to radon daughters used in this 

risk assessment is 360 lung cancer fatalities per 1.OE+6 WLM. The 

uncertainties associated with the risk factor for exposure to radon 

have been thoroughly reviewed and discussed (BEIR88, ICRP87, 

NCRP85) . Based on the ICRP 50 (ICRP87) report findings, an 

uncertainty range of 160-720 fatal lung cancers per 1.OE+6 WLM is 

arrived at in Chapter 6. These values represent estimates for the 

average individual for the case of constant lifetime exposure. The 

question of sensitive subgroups is not addressed. 

Major sources of uncertainty in radon risk factors for the 

general public include: (1) uncertainties in the epidemiological 

data on miners (especially with respect to exposure estimation); (2) 

uncertainties in the sex, age and time dependencies of the risk; (3) 

uncertainties in how differences in exposure conditions between 

mines and homes, particularly with respect to unattached fraction, 

affect risk; and (4) uncertainties regarding the possible 

synergistic effects of radon and cigarette smoke in causing lung 

cancer. 

7.3 DOE RADON 

Five sites are included in the DOE radon category: a mill 

tailings site in Monticello, Utah, the Middlesex Sampling Plant site 

in Middlesex, N.J., the Weldon Spring site in Weldon Spring, MO, the 

Niagara Falls Storage Site in Niagara Falls, NY, and the Feed 
Materials Production Center in Fernald, Ohio, As is the case for 

the uranium mill tailings, the risks associated with these 

facilities are due to exposure to radon and radon progeny. 

~ccordingly, the uncertainties are similar to those described for 

mill tailings. The following discussion is therefore limited to 

uncertainties unique to each facility. 

The tailings at the Monticello site are covered with a thin 

layer of earth to control the resuspension of particulates. The 

degree to which this retards the emanation of radon from the pile is 

not addressed. A site visit confirmed the location of the maximum 

individual, which reduces this source of uncertainty. 

The Weldon Springs site contains an old stone quarry and a 

former chemical processing plant that are sources of radon emission. 

DOE has measured the radon emission rates from these sources. 

Accordingly, the source terms are well characterized. One source of 

additional uncertainty is dispersion of radon from the quarry. The 

geometry of the release from a quarry may tend to increase 

turbulence and reduce exposures. The effect may be likened to 

building wake effects which can reduce near-field concentrations by 

about a factor of two. 

7.4 LICENSED URANIUM MILL TAILINGS FACILITIES 

Site-specific risk assessments have been performed for each of 

the 26 uranium mills licensed as of August 1988. The uncertainties 

in this assessment are similar to those described for inactive 

facilities. However, large portions of active mill tailings are 

moistened or are under water. These areas do not contribute 

significantly to the radon source term, but uncertainties in the 

assumptions regarding the portions of the piles that are dry 

contribute significantly to the uncertainty in the source term. In 

addition, there is evidence that the dry areas contain fewer slimes 

which tends to reduce the Ra-226 concentration in the dry areas. 

7.5 UNDERGROUND URANIUM MINES 

The doses and risks associated with airborne emissions from 

underground uranium mines are due to the high volume of mine exhaust 

air and the high concentrations of radon in the air. Though the 

radionuclide of concern is radon, there are many important 

differences in the uncertainties between mines and tailings. These 

differences are primarily related to the source term and the 

atmospheric dispersion factors. 

For three of the six mines, the radon release rate was 

calculated based on monitored levels of radon daughters. The radon 

concentration was estimated assuming a daughter equilibrium 

fraction of 0.07 observed at other mines. As applied to any one 

mine, this approach introduces a large uncertainty since the degree 
of equilibrium is mine specific, based on the residence time of air 

in the mine and the rate of deposition of the daughters in transit 

in the mine. For the three mines where the radon concentrations 

were measured directly, the uncertainty in the source term is 

negligible. 

The second important source of uncertainty is the near-field 

atmospheric dispersion factors for mines with rugged terrain. For 

these mines, the Industrial Source Complex Dispersion (ICS) model 

was used. Though this model is widely used for complex terrain and 

release configurations, the results may be uncertain by a factor of 

2 or more. 

7.6 PHOSPHOGYPSUM STACKS 

Like uranium mill tailings, the exposures from phosphogypsum 

stacks are due to radon emanation and resuspended particulates. 

Accordingly, the uncertainties in exposures are similar. However, 

unlike the uranium mill tailings, the radon flux from each 

individual stack was not directly estimated. Instead, the measured 

flux from one stack was used as the basis for calculating the flux 

from the other stacks in the region by applying corrections for 

dimensions, design, and use of the stack. Flux measurements for all 

the piles would have been preferable, but the large number and size 

of the piles made this impractical. 

Particulate emissions from the stacks were assumed to be 

entirely from vehicular fugitive emissions. These were calculated 

using an empirical model which considers vehicle speed, rainfall, 

and particle size. Each of these parameters varies in time and 

space. As a result, such models must be used discriminantly and, 

when possible, should be supplemented with field measurements (Be76, 

Mi75). 

Dispersion modeling of radon and particulates from 

phosphogypsum stacks is difficult due to the aerial extent and 

height of the stacks and the variability of the emission rates as a 

function of location. For a maximum individual close to the pile, 

the calculated airborne concentrations are highly uncertain because 

of these factors, and it is preferable to perform field measurements 

at the maximum individual location to confirm predicted values. For 

population exposures, the uncertainties are smaller because the pile 

behaves effectively as a point source. 

For particulate radionuclides, the uncertainties are further 

increased due to uncertainties in the dose conversion factors (DCFs) 

(Ga80, Me82). As discussed in Chapter 5, these uncertainties are 

due to uncertainty in the physical (i.e., particle size) and 

chemical (i.e., solubility) properties of the radionuclide, as well 

as uncertainties in dosimetric model formulation and the values of 

the model input parameters. 
7.7 ELEMENTAL PHOSPHORUS PLANTS 

Site-specific risk assessments were performed for each of the 

eight phosphate plants. Because site-specific meteorology and 

demography were employed and site visits were made to identify the 

locations of the maximum individual, these sources of uncertainty 

are small. 

The radionuclide emission rates from three of the eight 

operating phosphate plants have been extensively characterized in 

terms of radionuclide content and physical and chemical form. 

Accordingly, the uncertainty in the source term, which is dominated 

by Po-210 and Pb-210, is not an important source of uncertainty. 

Dispersion of the emissions is modeled for a buoyant, elevated 

release using Gaussian modeling. The uncertainty in the dispersion 

factor for near-field and far-field dispersion, described above, is 

an important contributor to uncertainty. 

Because the physical and chemical form of the release is well 

characterized, the uncertainty in the DCFs is limited to the 

inherent uncertainty of the dose conversion factors (DCFs) and the 

effects of individual variability. Both these apply primarily to 

the maximum individual dose. 

A major source of uncertainty is the risk factor used to 

convert dose to risk. As discussed in Chapter 6, the risk factor 

has an estimated uncertainty of about a factor of 3 for population 

exposures. Risk estimates reported for the maximum individual are 

based on the assumption that the maximum individual is similar to 

the average members of the population with regard to sensitivity to 

the potential carcinogenic effects of exposure to ionizing 

radiation. 

7.8 INACTIVE SURFACE URANIUM MINES 

Due to the large number of surface mines, it was necessary to 

establish categories of mines which were used to identify reference 

mines for use in the risk assessment. One reference mine in each of 

five regions was selected for detailed characterization. The 

principal measurements made were radon flux and Ra-226 

concentrations in soil. The source terms obtained for the reference 

mines were then scaled for use at other mines. 

In a similar manner, regional land use and demographic data 

were reviewed to define a representative site for each region. 

These data, along with the reference and scaled source terms, were 

used to calculate maximum individual and population doses. As 

applied to population doses, the uncertainties are small because 

individual site variabilities are averaged out. With regard to the 
maximum individual doses, the site and source terms were selected in 

order to ensure that the doses were not significantly 

underestimated. As a result, the basic approach used to 

characterize sites and land use is believed to contribute little to 

the overall uncertainty in dose and risk. 

The important sources of uncertainty are the methods used to 

model source terms and the inherent uncertainties in pathway, dose, 

and risk modeling. For surface mines, the critical radionuclides 

are radon and Ra-226, and the critical exposure 

pathways are the inhalation of radon daughters and Ra-226. The 

uncertainties are therefore similar to those described above. 

The uncertainty in the radon source terms is small since radon 

flux was measured directly. The uncertainty in the particulate 

source term is large because it was based on resuspension models 

which are subject to large uncertainties (Mi76, Be76). The 

variability in resuspension factors can be several orders of 

magnitude, depending on soil properties and weather. Because of 

this uncertainty, conservative assumptions were used to model 

resuspension which likely resulted in an overestimate of exposures 

from this pathway. 

7.9 URANIUM FUEL CYCLE FACILITIES 

The uranium fuel cycle facilities evaluated include uranium 

mills, conversion facilities, fuel fabrication facilities, and power 

reactors. 

7.9.1 Uranium Mills 

Site-specific impacts for each currently licensed uranium mill 

were evaluated. Site-specific source terms, meteorology, and 

demography were used. Because of the site specific nature of the 

analysis, uncertainties are reduced. 

The radionuclides and pathways contributing to dose are 

inhalation of Pb-210, Th-230, Ra-226, and U-238. Accordingly, 

uncertainty in particle size and solubility, along with the inherent 

uncertainty in the inhalation DCFs for these radionuclides, are 

important contributors to uncertainty (DOE75; Me82, Wr86). 

7.9.2 Uranium Conversion Facilities 

Separate assessments were performed for the two uranium 

conversion facilities. Site-specific source terms, meteorology, 

demography, and land use were used. Accordingly, the uncertainties 

associated with using model or reference facilities were avoided. 

The limiting pathways of exposure are the inhalation of Ra-226, Th230, and U-238. The important sources of uncertainty are as 

described for uranium mills. 
-1.9.3 Fuel Fabrication Facilities 

The five operating fuel fabrication facilities each have wellcharacterized source terms and site characteristics. The dose and 

risk assessments were performed for the most limiting facility. 

Accordingly, the maximum individual and population doses and risks 

are well-represented for the industry as a whole, but may 

overestimate risks for some specific facilities. 

The limiting pathway is inhalation of uranium. The important 

contributors to uncertainty are as discussed above for the 

inhalation of particulates (Fi86). 

7.9.4 Nuclear Power Plants 

There are over 100 commercial nuclear power plants in the 

United States. The radionuclide emissions and site characteristics 

of each facility are well defined and maintained current in 

accordance with regulatory requirements. In addition, each facility 

is required to demonstrate semiannually that its offsite impacts are 

maintained within the guidelines set forth in Appendix I to 10 CFR 

50. In brief, these guidelines are 5 mrem/yr to the whole body or 

any organ and 25 mrem/yr to the thyroid. As a result, this places 

an upper limit on the maximum individual doses and risks. A review 

of the reported impacts reveals that the doses are typically 

maintained at a fraction of a mrem/yr to a few mrem/yr. The 

populati.on impacts are reported to be about 1 to 10 person rems/yr 

per plant. 

For the purpose of this assessment, average annual source terms 

for BWRs and PWRs were used. A reference site with a generally 

high popul.ation density was selected so that the calculated doses 

would represent typical values for the higher populated sites. The 

upper bound doses for the maximally exposed individuals, based on 

values reported by each utility, were less than Appendix I 

guidelines. Because of the reference site selected, the population 

doses calculated are higher than the actual doses at most sites. 

The calculations revealed that for PWRs the limiting exposure 

pathway is external exposure to noble gases. Except for uncertainty 

in atmospheric dispersion, the uncertainties in these exposures are 

small. (It should be emphasized that, not withstanding these 

uncertainties, the maximum individual doses are all below the 

Appendix I limits.) The uncertainties in risk are due to the 

uncertainties in the risk factor, as described above. 

For BWRs, in addition to external exposures to noble gases, 

internal exposures from the ingestion of Sr-90 also contribute to 

dose. Thus, uncertainties in the environmental transfer factors and 
DCF for Sr-90 also contribute to uncertainty. Numerous studies have 

been published describing the uncertainties in ingestion pathway 

exposures (Jo88, Tr83, NUREG81, Sc80, Ng82, Ng77, NUREG79, Du81, 

Sh85, Ib88, Wh88, Li88). For example, the soil to plant 

concentration factor for Sr-90 was reported in NUREG/CR-2612 

(NUREG81) to have a geometric mean of 0.33 and a geometric standard 

deviation of 3.3. The basic data used to develop this distribution 

comes from the work of Ng (1982). The distribution incorporates 

radioactive and stable data for crops grown in both laboratory and 

field conditions and in soil of virtually every texture 

classification. 

7.10 NRC-LICENSED FACILITIES AND NON-DOE FEDERAL FACILITIES 

7.10.1 NRC-Licensed Facilities - Hospitals 

The number and types of NRC and NRC Agreement State licensees 

are large and highly varied. For the purpose of evaluating the 

uncertainties in the risks of airborne emissions from these 

facilities, hospitals are selected as conservatively representative 

of this source category. Hospitals handle relatively large 

quantities of volatile radionuclides and are usually located in 

populated areas. Accordingly, their potential for offsite exposure 

from airborne radionuclide emissions is relatively large. 

Section 3.2 of Volume 11 characterizes the number and types of 

sources of airborne emissions of radionuclides from hospitals. 

Because of the large numbers of facilities and varied practices, 

along with minimal routine reporting requirements, the offsite doses 

and risks attributable to any one hospital or to the entire category 

are highly uncertain. 

Because of the complexity of this category, and the limited 

data available, a set of model facilities was defined in Section 

3.2.2, which is used to extrapolate to the category as a whole. 

This is the first source of significant uncertainty, especially in 

terms of the representativeness of the model facilities for any one 

hospital. The representativeness issue is significant when 

attempting to characterize the doses and risks to the offsite 

maximum individual at any one facility, and, as is discussed below, 

the approach selected is likely to overestimate the doses and risks 

at most facilities. 

The representativeness issue is of less concern when 

considering the population doses and risks associated with the 

entire category because individual differences among hospitals tend 

to average out. However, as discussed below, the model facilities 

are fairly conservative representatives. 

The first and perhaps the most important source of uncertainty 

is the radionuclide emission rates. Source term data were obtained 
from surveys of 100 hospitals (performed by tlie Conference of 

Radiation Control Program Directors [CRCPD] and using data available 

from the DOE/EG&G database). These data are used to represent 3,680 

hospitals in the United States that handle diagnostic or therapeutic 

amounts of radiopharmaceuticals. Based on these data, the largest 

reported emissions were used to represent large hospitals, and the 

average of the reported emissions was used as representative of 

typical hospitals. 

The uncertainties in these emissions are difficult to estimate 

since they are based on periodic grab samples taken from effluent 

release points and analyses designed to ensure that the discharges 

do not exceed the offsite maximum permissible concentrations (MPC) 

set forth in Appendix B to 10 CFR 20. As a result, the precision of 

the estimate is not required to be high, and non-monitored release 

points are not addressed. Another source of uncertainty is the 

absence of C-14 and H-3 in the estimate of release. Hospitals, 

especially research hospitals, handle large quantities of these 

radionuclides which can be readily discharged in the gaseous 

effluents. It is difficult to address the uncertainties in the dose 

and risk estimates due to these considerations. 

The second source of uncertainty is the assumed location of the 

maximum individual. The analysis is based on the assumption that 

the nearest individual is 100 meters away. This is a reasonable 

assumption for most facilities; however, for some facilities this 

assumption may be highly conservative. 

For populations, the assessment divided the category into urban 

and rural locations, each with markedly different demographic and 

land use characteristics. In the aggregate, the dose estimates are 

likely to be reasonable. However, not 

including C-14 and H-3 releases could result in an underestimate in 

the population doses because these are long-lived radionuclides. 

Given the uncertainty in the location of the nearby individual 

and land use patterns, the uncertainties associated with atmospheric 

dispersion, the exposure pathways (including food), and the dose 

conversion factors are not likely to contribute significantly to the 

uncertainty in risk. The external and internal dosimetry and 

environmental behavior of the radionuclides considered e, Xe-133 

and 1-131) are well understood. Finally, the uncertainty in risk, 

due to uncertainty in the risk factor, is about a factor of three. 

In summary, the uncertainties in the doses and risks to 

individuals and the populations in the vicinity of hospitals is 

dominated by uncertainty in the source term. In addition, 

uncertainties in demographic characteristics and the risk factor are 

important contributors to uncertainty. 

These observations also apply to most other NRC-licensed 
facilities and Agreement State licensees, since the types of 

radionuclides emitted in the gaseous effluents (i.e., 

1-131, Xe-133, H-3, and C-14) are similar for these licensees. In 

addition, the regulatory requirements, including reporting and 

record-keeping requirements, are similar. The types of facilities 

include radiopharmaceutical manufacturers, laboratories, and 

universities. 

7.10.2 NRC-Licensed Facilities - Research and Test Reactors 

Research and test reactors are NRC-licensed facilities that 

differ significantly from hospitals and require separate 

consideration. The source terms from these facilities are better 

characterized and are limited to radionuclides that are relatively 

easy to model; i.e. Ar-41 and H-3. The approach used in Section 3.5 

of Volume I1 to assess risk is to select the four largest emitters 

of the 70 facilities, and estimate doses and risks for these 

facilities. The potential for the doses and risks being greater 

than the indicated values is small. 

The population risk was estimated by prorating by the total 

source term for the entire category. Accordingly, the source term 

is likely to be fairly accurate. However, demographic patterns at 

different sites may be quite different than the four reference 

sites. Since the reference sites are in highly populated urban 

areas, the population risks are appropriate for the reference sites 

but are conservative as applies to the category as a whole. 

As a final point, since the key radionuclide emissions in this 

category are H-3 and Ar-41, dose conversion factors do not 

contribute significantly to uncertainty. In addition, the food 

ingestion pathway is not significant. As a result, uncertainties 

due to these factors are small. 

7.10.3 NRC Licensed Facilities-Source Material Licensees 

Sections 3.7 and 3.8 of Volume 11 discuss emissions from NRC 

licensed facilities that fabricate fuel for research reactors or are 

licensed to handle source material. These facilities primarily emit 

uranium and thorium. Accordingly, the discussion provided in 

Section 7.9 on fuel fabrication facilities for the uranium fuel 

cycle is applicable to this category also. 

These facilities maintain effluent records and, therefore, the 

source terms are well defined. Because the number of such 

facilities is limited, the analyses were performed for individual 

facilities or a conservatively selected number of reference 

facilities. Accordingly, representativeness is not an important 

source of uncertainty. However, the environmental behavior and 

internal dosimetry of uranium and thorium is uncertain because the 

particle size and chemical composition of the emissions are not well 
defined. Nevertheless, conservative assumptions were used which 

provides assurance that the doses and risks were not underestimated. 

7.10.4. Non-DOE Federal Facilities 

As discussed in Section 3.10 of Volume 11, this category 

consists primarily of shipyards and naval bases and DOD research 

reactors. The research reactors are similar to the research and 

test reactor category discussed above. Accordingly, this discussion 

emphasizes the shipyards and naval bases. 

The airborne emissions from the shipyards are upper bound 

estimates based on many years of monitoring. Using these source 

terms, the doses are risks at a reference site were calculated. 

Because of the conservative source terms and the highly populated 

west coast reference site used in the analysis, the doses and risks 

represent upper bound estimates. 

7.11 COAL-FIRED BOILERS 

The dose and risk assessment provided in Section 7 of Volume I1 

attempted to account for all the important variables in selecting a 

set of reference facilities and sites for both utility and 

industrial boilers. In order to calculate the risks to the maximum 

individual, utility boilers were grouped into four categories 

according to demography. For each category, the largest boilers 

were identified to calculate the doses to the maximum individuals. 

For the population doses, the facilities with the average emissions 

for each category were used as reference facilities. This approach 

places an upper bound on the individual risks and may overestimate 

individual risks for some facilities. 

Another important source of uncertainty for individual risk is 

the dose conversion factor. The particle size and the chemical 

composition of the inhaled radionuclides are variable among sites 

and as a function of time. The doses to individuals from the food 

ingestion pathways are also highly uncertain due to uncertainty in 

the environmental transport parameters for long lived naturally 

occurring radionuclides. For population doses, these uncertainties 

are markedly reduced due to the effects of averaging. 

The uncertainties in the doses and risks associated with 

airborne radionuclide emissions from industrial boilers are similar 

to those for utility boilers. For the purpose of the 

risk assessment, the industrial boiler with the largest particulate 

release was selected as the reference plant. The radionuclide 

concentration in the particulate emissions from industrial boilers 

was estimated indirectly based on the utility boiler experience. 

Since the largest industrial boiler was used as the reference 

facility, and since assumptions applicable to utility boiler 

emissions are conservative as applied to industrial boilers, the 
source term is a generally conservative vaiue for most fncil~ities. 

The actual site data at the reference plants were used. As a 

result, some plants have higher or lower population densities and 

land use characteristics. It is not possible to judge the degree of 

representatives associated with the reference sites. However, given 

the conservative source term assumptions, it is likely that the 

population doses and risks are reasonable values, if not soniewhat 

conservative. 

7.12 DOE FACILITIES 

As indicated in Chapter 2 of Volume 11, there are a large 

number of diverse DOE facilities with highly varied site 

characteristics. It is not possible to discuss the important 

sources of uncertainty for this category in the aggregate. Instead, 

a site specific assessment of each facility is required. In order 

to provide insight into the nature and extent of the uncertainties, 

a discussion of two of the facilities is provided. 

7.12.1 Reactive Metals, Inc. 

As indicated in Section 2.2 of Volume 11, Reactive Metals, Inc. 

processes uranium ingots and releases particulate uranium, probably 

uranium oxide, in its multiple stack effluents at the site. The 

emissions are monitored; however, the particle size and solubility 

is not monitored. As a result, the environmental behavior of the 

uranium and its inhalation dose conversion factors are uncertain. 

Another source of uncertainty is due to the method used to model 

dispersion. The uranium is discharged at three stacks, but is 

modeled as a single source. 

Due to the lack of onsite meteorological data, data from the 

nearest airport were used. Nevertheless, as discussed above in 

Chapter 4, uncertainty in the average annual atmosphere dispersion 

factors do not contribute significantly to overall uncertainty as 

long as the location of this maximum individual is well defined. In 

this case, site specific information was available which reduces 

this source of uncertainty. 

Since the facility is located in an urban area, the food 

pathway is not an important contribution to dose or risk. 

7.12.2 Hanford Reservation 

As discussed in Section 2.3 of Volume 11, the airborne 

emissions from the Hanford Reservation originate from several 

different sources located in each of four separate production areas. 

The risk assessment, however, assumed that the releases were all 

from the fuel processing area of the reservation, which is the 

largest source of release. This is a somewhat conservative 
assumption since the cioses from one point source is usually larger 

than the dose from multiple sources with the same total release 

quantity. Offsetting this conservatism is the possibility that some 

receptor locations many be closer to one of the release sources than 

the 15 Km assumed in the analyses. Without a more refined analysis 

of individual sources and receptors, it is difficult to assess 

whether overall these factors contribute to an overestimate or 

underestimate of the dose. However, as indicated below, assuming 

the entire release is from the reprocessing area tends to place an 

upper bound on the individual doses due to its location relative to 

the nearby populations. 

The emissions from the facility are well characterized but 

default particle size and solubility assumptions were used for 

particulates, resulting in a possible overestimate of the inhalation 

dose. However, since less than 50 percent of the calculated dose 

was due to particulates (most of the dose was due to Ar-41), any 

overestimate in dose due to these assumptions is less than a factor 

of 2. 

Since the doses are dominated by external exposure to Ar-41, 

the uncertainty in the dose conversion factors is not an important 

contribution to the uncertainty in dose. In addition, Ar-41 is 

primarily emitted from the N-Reactor, and not from the reprocessing 

area. This assumption tends to overestimate dose because the 

reprocessing area is closer to the populated areas and in the more 

predominant wind direction. 

7.13 HIGH-LEVEL WASTE DISPOSAL FACILITIES 

No operational high-level waste (HLW) disposal facilities are 

operating in the United States, nor are any routine air emissions 

anticipated from planned facilities. Nevertheless, estimates of 

radionuclide emissions were made for the planned Monitored 

Retrievable Storage (MRS) facility, the Waste Isolation 

Pilot Program (WIPP), and the commercial HLW disposal facility 

planned for Yucca Mountain. The source terms were estimated 

assuming a waste or spent fuel handling accident which releases 

radionuclides in accord with a defined set of conservative 

assumptions. 

The inhalation pathway was found to be the critical pathway. 

The critical radionuclides are H-3, Kr-85 and 1-129 for the MRS and 

Yucca, and Pu-239 and Am-241 for WIPP. The most important source of 

uncertainty in these calculations is the assumption made with 

respect to the likelihood of accidents occurring and releasing the 

defined types and quantities of radionuclides. 
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APPENDIX A 

ASSESSMENT METHODOLOGY 

A.1 INTRODUCTION 

This appendix to Volume I provides a brief overview of some of 

the key calculational assumptions used by the Environmental 

Protection Agency (EPA) to assess the doses and health risk from 

radiation exposures. 

A.2 ENVIRONMENTAL PATHWAY MODELING 

A.2.1 Individual Assessment 

The nearby individuals were assessed on the following basis: 

(1) The nearby individuals for each source category are 

intended to represent an average of individuals living 

near each facility within the source category. The 

location of one or more persons on the assessment grid 

which provides the greatest lifetime risk (all pathways 

considered) was chosen for the nearby individuals. 

(2) The organ dose-equivalent rates in the tables are based 

on the calculated environmental concentrations by AIRDOSEPA (Mo79). For inhaled or ingested radionuclides, the 

conversion factors are 50-year committed dose 

equivalents. 

(3) The individual is assumed to home-grow a portion of his 

or her diet consistent with the type of site. Individuals 

living in urban areas were assumed to consume much less 

home-produced food than an individual living in a rural 

area. It was assumed that in an agriculturally 

unproductive location, people would home-produce a 

portion of their food comparable to residents of an urban 

area, and so the urban fraction is used for such nonurban 

locations. The fractions of home- produced food consumed 

by individuals for the generic sites are shown in Table 

A-1 . 

Table A-1. Presumed sources of food for urban and rural sites. 

Food Urban/Low productivity Rural 

Vegetables 

Meat 

Milk 
F1 and F2 are the home-produced fractions at the i.ndividualsn 

location and within the 80 km assessment area, respectively. The 

balance of the diet, F3, is considered to be imported from outside 

the assessment area, with negligible radionuclide concentrations due 

to the assessed source. If there is insufficient production of a 

food category within the assessment area to provide the non houseproduced fraction for the population, F2 is reduced and F3 is 

increased accordingly. Fractions are based on an analysis of 

household data from the USDA 1965-1966 National Food Consumption 

Survey (USDA72) . 

A.2.2 Collective Assessment 

The collective assessment to the population within an 80 km radius 

of the facility under consideration was performed as follows: 

(1) The population distribution around the generic site was 

based on the 1980 census. The population was assumed to 

remain stationary in time. 

(2) Average agricultural production data for the state in 

which the generic site is located were assumed for all 

distances greater than 500 meters from the source. For 

distance less than 500 meters, no agricultural production 

is calculated. 

(3) The population in the assessment area consumes food from 

the assessment area to the extent that the calculated 

production allows. Any additional food required is 

assumed to be imported without contamination by the 

assessment source. Any surplus is not considered in the 

assessment. 

(4) The collective organ dose-equivalent rates are based on 

the calculated environmental concentrations. Fifty-year 

dose commitment factors (as for the individual case) are 

used for ingestion and inhalation. The collective dose 

equivalent rates in the tables can be considered to be 

either the dose commitment rates after 100 years of 

plant operation, or equivalently, the incurred doses that 

will be for up to 100 years from the time of release. 

Tables A-2 and A-3 summarizes AIRDOS-EPA parameters used 

for the assessments (Sj84). 

Table A-2 summarizes agricultural model parameters, usage 

factors, and other AIRDOS-EPA parameters which are independent of 

the released radionuclides. Table A-3 tabulates element dependent 

data. These include the default inhalation clearance class and, the 

fraction of the stable element reacting body fluids after ingestion. 

Inhaled clearance classes D, W and Y correspond to those materials 

which clear from the lung over periods of days, weeks, and years 

respectively. Class * is for gases. Biv, and Biv, are the soil to 

pasture and soil to produce concentration factors respectively. 
Both factors are for soil concentration on a dry weight basis. The 

pasture and produce concentrations are on dry and fresh weight bases 

respectively. 

Fm and Ff relate the stable element intake rate to the 

concentration in milk and meat, respectively. The values for the 

factors in this table are maintained in the PREPAR file ACCRAD 

(S j84) . 

A.2.3 Dairy and Beef Cattle 

Dairy and beef cattle distributions are part of the AIRDOS-EPA 

input. A constant cattle density is assumed except for the area 

closest to the source or stack in the case of a point source, i.e., 

no cattle within 500 m of the source. These densities were derived 

from data developed by NRC (NRC75). Milk production density in 

units of liters/day-square mile was converted to number of dairy 

cattle/square kilometer by assuming a milk production rate of 11.0 

liters/day per dairy cow. Meat production density in units of 

kilograms/day-square mile was changed to an equivalent number of 

beef cattle/square kilometer by assuming a slaughter rate of .00381 

day-1 and 200 kilograms of beef/animal slaughtered. A 180-day 

grazing period was assumed for dairy and beef cattle. 

A.2.4 Vegetable Crop Area 

A certain fraction of the land within 80 km of the source is 

used for vegetable crop production and is assumed to be uniformly 

distributed throughout the entire assessment area with the exception 

of the first 500 meters from the source. Information on the 

vegetable production density in terms of kilograms (fresh 

weight)/day-square mile was obtained from NRC data (NRC75). The 

vegetable crop fractions by state were obtained from the production 

densities by assuming a production rate of 2 kilograms (fresh 

weight)/year-square meter (NRC77). 

A.2.5 Population 

The population data for each generic site were generated by a 

computer program, SECPOP (At74). which utilizes an edited and 

compressed version of the 1980 United States Census Bureau's MARF 

data containing housing and population counts for each census 

enumeration district (CED) and the geographic coordinates of the 

population centroid for the district. In the Standard Metropolitan 

Statistical Areas (SMSA), the CED is usually a "block group" which 

consists of a physical city block. Outside the SMSAs, the CED is an 

"enumeration district," which may cover several square miles or more 

in a rural area. 

There are over 250,000 CEDs in the United States with a typical 

population of about 800 persons. The position of the population 

centroid for each CED was marked on the district maps by the 

individual census official responsible for each district and is 
based only on personal judgrnerlt from inspection of the population 

distribution 011 a map. The CED entries are sorted i.s ascending 

order by longitude on the final. data tape. 

The resolution of a calculated population distribution cannot 

be better than the distribution of the CEDs. Hence, in a 

metropolitan area the resolution is often as small as one block, but 

in rural areas it may be on the order of a mile or more. 

A.2.6 Risk Conversion Factors 

Table A-5 summarizes the average lifetime risk per unit intake 

or exposure for most of the radionuclides considered in the 

assessments. Note that the external exposure factors do not include 

the contribution from any decay products. For example, the external 

risk factors for cesium-137 have values of 0. since there is no 

photon released in its decay. Hence, the exposure due to the 

cesium-137 decay product barium-137m must be considered in assessing 

cesium-137. The clearance class and gut-to-blood transfer factor, 

f,, values are shown in Table A-3. 
Tab1.e A-2. A:illDOS-E:PA parameters used for gener:lc site assessments 

Symbo1.i~ 

variable Description Value 

BRTHRT 

T 

DDI 

Breathing Rate (cm3/h) 

Surface buildup time (days) 

Activity fraction after washing 

TSUBHI 

TSUBHZ 

TSUBH3 

Time delay-pasture grass (h) 

Time delay-stored food (h) 

Time delay-leafy vegetables (h) 

Weathering removal rate 

factor (h") 

Exposure period-pasture (h) 

Exposure period-crops or leafy 

vegetables (h) 

Productivity-pasture (dry 

weight) (kg/m2) 

Productivity-crops and leafy 

vegetables (kg/m2) 

FSUBP 

FSUBS 

Time fraction-pasture grazing 

Pasture feed f raction-while 

pasture grazing 

QSUBF Feed or forage consumption 

rate (kg-dry/day) 

TSUBF Consumption delay time-milk (d) 

Vegetable utilization rate (kg/y) 

Milk utilization rate (kg/y) 

Meat utilization rate (kg/y) 

Leafy vegetable utilization 

rate ikg/y) 

TSUBS Consumption time delay-meat (days) 
Table A-2. AIRDOS-EPA parameters used for generic site assessments 

(Continued) 

Symbolic 

variable Description Value 

FSUBG 

FSUBL 

TSUBB 

P 

TAUBEF 

MSUBB 

VSUBM 

Produce fraction (garden of 

interest) 

Leafy veg fraction (garden of 

interest) 

Soil buildup time (y) 

Effective surface density of 

soil (kg/m2) 

Meat herd-slaughter rate 

factor (d") 

Mass of meat of slaughter (kg) 

Milk production rate of cow (L/d) 

Deposition interception fractionpasture 

Deposition interception fractionleafy vegetables 
Table? A--3. nt!faii It vallii?s iieed for elemerit dependent factors 

mer\t Class fl 

W 1 . OE- 3 

Y 3.0E-1 

Y I. ok:-1 

n 9.5~-1 

Y 5. OR- I 

0.0 0.0 

1.OE-2 7.OE-4 

1.OE-3 5.5E-4 

2. OE- 5 7.5E-4 

0.0 0.0 

1.. OE-2 7.OE-3 

I.. OE-4 8.OE-3 

2.OE--6 1.5E-3 

7. OE- 3 2.OE-2 

0.0 0.0 
at. A- i,~:t,~iiii v~iiric!:; iisiiii fox. e?i<inteiiL dependen?. factors. 

((?ani:irrt~ed) 

Nb Y I.. OE-7 

Nd Y 3 .OK4 

N i. W 5. OF-? 

NP W I..OE-Z 

0 I 9.5E-1. 

P 0 W 1 .OR-1 

P r Y ?.OR-4 

PI1 Y I. OR-?* 

RH W 7. OE-I 

K h n 9.5~- 1 

R (? W R .OF:-1 

Hh Y !i . ()I?-2 

Hn X 0.0 

HI1 Y 5. or-2 

S r) 8. OR-I. 

S h W 1 .O1!-1 

SC Y 1 .O4-4 

Sf? W R .OR-] 

s n~ w ?.RE-4 

S tl W ?.OR-2 

s r n 3. OTC- I 

1" h W 3 . OK-4 

7'C W 8 .OX-] 

Tf? W 2. OE-I 

T h Y 2.0ii:- 4 

3.5 1. . 5 l.5E-Z 

2.5K-3 1 .1E-4 5.OE-6 

4.5E-2 3.90-3 2.5E-4 

I .5E-I 1.7E-2 1.. OE- 2 

1.. OE-Z 1 .7E-3 2.OE-5 

Z. SE-Z 1.7E-3 3.5E-4 

1.0E-2 1.7E-3 2. OE- 5 

4.5R-4 1 .9E-5 1.. OE-7 

I .5E-2 6.4E-4 4.5E-4 

1 .5E-1. 3.OE-2 I.. OE-2 

Z.OR--l 1.3E-2 1. . OE- 4 

6.0E-3 4.3R-4 5. OE-6 

2. SE-Z 1 .I E-2 4.OE-3 

1 . OK-2 1 .7E- 3 2.OE-5 

3.0E-2 2.6E-3 I.OE--3 

2.5 1 . IR-I. 1.5E-3 

1 . OE-Z 1 .7E-3 2.OE-5 

9. !iE 6.41:- I L.0E-Z 

2. 5A-Z I.. 71?- 3 2.0E-4 

R. 5E-4 3.6E-5 S.OE--6 

1.5 3.9E-1 I.. OE-2 

2.OE-3 2.10-4 3.OE-5 

3. OE- 4 

4.5E-3 

8.5E-3 

1.5E-2 

6. OE--6 

1. OE-I 

5.5E-3 
Table A-4. Cattle densities and vegetable crop 

distributions for use with AIKDOS-EPA 

Dairy cattle 

State density 

iil km2 

Alabama 7.02E-1 

Arizona 2.80E-1 

Arkansas 5.90E-1 

California 2.85 

Colorado 3.50E-1 

Connecticut 2.50E-1 

Delaware 2.72 

Florida 1.37 

Georgia 8.63E-1 

Idaho 8.56E-1 

Illinois 2.16 

Indiana 2.80 

Iowa 3.14 

Kansas 8.00E-1 

Kentucky 2.57 

Louisiana 9.62E-1 

Maine 8.07E-1 

Mary land 6.11 

Massachusetts 3.13 

Michigan 3.51 

Minnesota 4.88 

Hississippi 8.70E-1 

Missouri 1.89 

Montana 9.27E-2 

Nebraska 8.78E-1 

Nevada 5.65E-2 

New Hampshire 1.58 

New Jersey 3.29 

New Mexico 1.14E-1 

New York 8.56 

Beef cattle Vegetable 

density crop fraction 

iilkm2 kin2 /km2 
Table A-4. Cattle densities and vegetable crop 

distributions for use with ALRDOS-EPA--continued 

Dairy cattle Beef cattle Vegetable 

State density density crop fraction 

i//km2 #lkd km2jkm2 

North Carolina 1.26 1.02E+1 6.32E-3 

North Dakota 6.25E-1 1.18E+1 6.29E-2 

Ohio 4.56 2.03E+1 1.70E -2 

Oklahoma 7.13E-1 2.68E+l 2.8OE-2 

Oregon 4.53E-1 4.56 1.59E-2 

Pennsylvania 6.46 9.63 1.32E-2 

Rhode Island 2.30 2.50 4.54E-2 

South Carolina 7.02E-1 8.87 1.84E-3 

South Dakota 8.85E-1 2.32E+1 1.2OE-2 

Tennessee 2.OOE-1 2.11E+1 2.72E-3 

Texas 5.30E-1 1.90E+1 5.77E-3 

Utah 4.46E-1 2.84 1.83E-3 

Vermont 8.88 4.71 l.08E-3 

Virginia 1.84 1.31E+1 8.7OE-3 

Washington 1.50 5.62 5.2OE-2 

West Virgina 6.00E-1 6.23 1.16E-3 

Wisconsin 1.43E+1 1.81E+1 1.7BE-2 

Wyoming. 5.79E-2 5.12 1.59E-3 
* * * DRAFT * * * 

Table A-5. Risk factors for selected radionuclides (see Table A-3 for 

default inhalation class and ingestion f, values). 

Nuclide Inhal . Ingest. Immer . Surface 

(PC~.') (rCi-') (m3/rci yr) (mz/rCi yr) 
Risk factors for selected radionuclides (see Table 

A-3 for default inhalation class and ingestion f, 

values). (Continued) 

Table A-5. 

Nuclide Inhal. Ingest. Immer . Surface 

(rCi-') (r~i-') (m3/@Ci yr) (m2/rci yr) 
Table A-5. Risk factors for selected radionuclides (see Table 

A-3 for default inhalation class and ingestion f, 

values) . (Continued) 

Nuclide Inhal. Ingest. Immer . Surface 

(pCim') (~ci-') (m3/pci yr) (rn2/u-~i yr) 

O.OE+OO 

0. OEtOO 

3.43-03 

3.63-03 

6.43-04 

O.OE+OO 

0. OEtOO 

6.03-05 

6.63-05 

1.4E-05 
REFERENCES 

At74 Athey T.W., R.A. Tell, and D.E. Janes, 1974, The Use of 

an Automated Data Base in Population Exposure Calculations, 

from Population Exposures, Health Physics Society, 

CONF-74018, October 1974. 

Mo79 Moore R.E., C.F. Baes, 111, L.M. McDowell-Boyer, A.P. 

Watson, F.O. Hoffman, J.C. Pleasant, C.W. Miller, 1979, 

AIRDOS-EPA: A Computerized Methodology for Estimating 

Environmental Concentrations and Dose to Man from Airborne 

Releases of Radionuclides, EPA 520/1-79-009, EPA Office of 

Radiation Programs, Washington, D.C. 20460, December 1979. 

NRC7 5 Memo from K. Eckerman, N. Dayem, R. Emch, Radiological 

Assessment Branch, Division of Technical Review, Nuclear 

Regulatory Commission, Code Input Data for Man-Rem Estimates 

(Washington, DC, October 15, 1975). 

NRC77 U.S. Nuclear Regulatory Commission, Calculation of Annual 

Doses to Man from Routine Releases of Reactor Effluents for 

the Purpose of Evaluating Compliance with 10 CFR Part 50 

Appendix I (Revision I), Regulatory Guide 1.109, Office of 

Standards Development, Washington, D.C., October 1977. 

S j84 Sjoreen, A.L., and Miller, C.W., PRERPAR - A User Friendly 

Preprocessor to Create AIRDOS-EPA Input Data Sets, ORNL-5952, 

August 1984. 

USDA72 United States Department of Agriculture, 1972, Food 

Consumption of Households in the United States (Seasons and 

Year 1965-1966), Household Food Consumption Survey 

1965-1966, Report No. 12, Agricultural Reseach Service, 

USDA, Washington, DC (March 1972). 
APPENDIX B 

MECHANICS OF THE LIFE TABLE IMPLEMENTATION OF THE RISK ESTIMATES 

B.1 INTRODUCTION 

This appendix describes the mechanics of the life table 

implementation of the risk estimates derived in Chapter 6. 

B.2 LIFE TABLE ANALYSIS TO ESTIMATE THE RISK OF EXCESS CANCER 

Radiation effects can be classified as stochastic or 

nonstochastic (NAS80, ICRP77). For stochastic effects, the 

probability of occurrence of the effect, as opposed to the severity, 

is a function of dose; induction of cancer, for example, is 

considered a stochastic effect. Nonstochastic effects are those 

health effects for which the severity of the effect is a function of 

dose; examples of nonstochastic effects include cell killing, 

suppression of cell division, cataracts, and nonmalignant skin 

damage. At the low levels of radiation exposure attributed to 

radionuclides in the environment, the principal health detriment is 

the induction of cancers (solid tumors and leukemia) and the 

expression, in later generations, of genetic effects. In order to 

estimate these effects, instantaneous dose rates for each organ at 

specified times are sent to a subroutine adaptation of CAIRD (Co78) 

contained in the RADRISK code. This subroutine uses annual doses 

derived from the transmitted dose rates to estimate the number of 

incremental fatalities in the cohort due to radiation induced cancer 

in the reference organ. The calculation of incremental fatalities 

is based on estimated annual incremental risks, computed from annual 

doses to the organ, together with radiation risk factors, such as 

those given in tha 1980 NAS report BEIR-3 (NASBO). Derivation of 

the risk factors in current use is discussed in Chapter 6. 

An important feature of this methodology is the use of 

actuarial life tables to account for the time dependence of the 

radiation insult and to allow for competing risks of death in the 

estimation of risk due to radiation exposure. A life table consists 

of data describing age-specific mortality rates from all causes of 

death for a given population. This information is derived from data 

obtained on actual mortality rates in a real population. Mortality 

data for the U.S. population during the years 1969-1971 (HEW75) are 

used throughout this study. 

The use of life tables in studies of risk due to low-level 

radiation exposure is important because of the time delay inherent 

in radiation risk. After a radiation dose is received, there is a 

minimum induction period of several years (latency period) before a 

cancer is clinically observed. Following the latency period, the 

probability of occurrence of a cancer during a given year is assumed 
to be constant for a specified period, called a plateau period. The 

length of both the latency and plateau periods depends upon the type 

of cancer. During or after radiation exposure, a potential cancer 

victim may experience years of life in which he is continually 

exposed to risk of death from causes other than incremental 

radiation exposure. Hence, some individuals will be lost from the 

population due to competing causes of death and are not potential 

victims of incremental radiation-induced cancer. 

It is assumed that each member of the hypothetical cohort is 

exposed to a specified activity of a given radionuclide. In this 

analysis, each member of the cohort annually inhales or ingests 1 

pCi of the nuclide, or is exposed to a constant external 

concentration of 1 pCi/cm3 in air or 1 pCi/cm2 on ground surfaces. 

Since the models used in RADRISK are linear, these results may be 

scaled to evaluate other exposure conditions. The cohort consists 

of an initial population of 100,000 persons, all of whom are 

simultaneously liveborn. In the scenario employed here, the 

radiation exposure is assumed to begin at birth and continue 

throughout the entire lifetime of each individual. No member of the 

cohort lives more than 110 years. The span from 0 to 110 years is 

divided into nine age intervals, and dose rates to specified organs 

at the midpoints of the age intervals are used as estimates of the 

annual dose during the age interval. For a given organ, the 

incremental probability of death due to radiation-induced cancer is 

estimated for each year using radiation risk factors and the 

calculated doses during that year and relevant preceding years. 

The incremental probabilities of death are used in conjunction 

with the actuarial life tables to estimate the incremental number of 

radiation-induced deaths each year. The estimation of the number of 

premature deaths proceeds in the following manner. At the beginning 

of each year, m, there is a proba- bility, P, of dying during that 

year from nonradiological causes, as calculated from the life table 

data, and an estimated incremental probability PR of dying during 

that year due to radiation-induced cancer of the given organ. In 

general, for the m-th year, the calculations are: 

M (m) = total number of deaths in cohort during year m, 

= [PN (m) + PR (m) x N (m) Q (m) 

= incremental number of deaths during year m due to 

radiation-induced cancer of a given organ 
N(m+l) = number of survivors at the beginning of year m i l 

= N(rn) - M (m) (N(l)=100,000) . 

PR is assumed to be small relative to PN, an assumption which 

is reasonable only for low-level exposures (BuEl), such as those 

considered here. The total number of incremental deaths for the 

cohort is then obtained by summing Q(m) over all organs for 110 

years. 

In addition to providing an estimate of the incremental number of 

deaths, the life table methodology can be used to estimate the total 

number of years of life lost to those dying of radiation-induced 

cancer, the average number of years of life lost per incremental 

mortality, and the decrease in the population's life expectancy. 

The total number of years of life lost to those dying of 

radiation-induced cancer is computed as the difference between the 

total number of years of life lived by the cohort assuming no 

incremental radiation risk, and the total number of years of life 

lived by the same cohort assuming the incremental risk from 

radiation. The decrease in the population's life expectancy can be 

calculated as the total years of life lost divided by tha original 

cohort size (N(1)=100,000), 

Either absolute or relative risk factors can be used. Absolute 

risk factors, given in terms of deaths per unit dose, are based on 

the assumption that there is some absolute number of deaths in a 

population exposed at a given age per unit of dose. Relative risk 

factors, the percentage increase in the ambient cancer death rate 

per unit dose, are based on the assumption that tha annual rate of 

radiation-induced excess cancer deaths, due to a. specific type of 

cancer, is proportional to the ambient rate of occurrence of fatal 

cancers of that type. EitBer the absolute or the relative risk 

factor is assumed to apply uniformly during a plateau period, 

beginning at the end of the latent period. 

The estimates of incremental deaths in the cohort from chronic 

exposure are identical to those obtained if a corresponding 

stationary population (i.e., a population in which equal numbers of 

persons are born and die in each year) is subjected to an acute 

radiation dose of the same magnitude. Since the total person-years 

lived by the cohort in this study is approximately 7.07 million, the 

estimates of incremental mortality in the cohort from chronic 

irradiation also apply to a one-year dose of the same magnitude to a 
popillati.on of this size, age distribution, and age-specific 

mortality rates. More precise life table estimates for a specific 

popu:l.ation can be obtained by altering the structure of the cohort 

to reflect the age distribution of a particular population at risk. 

In addition, since the stationary population is formed by 

superposition of all age groups in the cohort, each age group 

corresponds to a segment of the stationary population with the total 

population equal to the sum of all the age groups. Therefore, the 

number of excess fatal cancers calculated for lifetime exposure of 

the cohort at a constant dose rate would be numerically equal to 

that calculated for the stationary population exposed to an annual 

dose of the same magnitude. Thus, the risk estimates may be 

reported as a lifetime risk (the cohort interpretation) or as the 

risk ensuing from an annual exposure to the stationary population. 

This equivalence is particularly useful in analyzing acute 

population exposures. For example, estimates for a stationary 

population exposed to annual doses that vary from year to year may 

be obtained by summing the results of a series of cohort 

calculations at various annual dose rates. 
REFERENCES 

Bunger, B.M., Cook, J.R., and Barrick, M.K.,"Life Table 

Methodoloav - for Evaluatina > Radiation Risk: An 

Application Based on Occupational Exposures," Health 

40, 439-455. 

Cook, J.R., Bunger, B., and Barrick, M.K., A Computer 

Code for Cohort Analysis of Increased Risks of Death 

(CAIRD), EPA 520/4-78-012, 1978. 

U.S. Department of Health Education and Welfare, "U.S. 

Decennial Life Tables for 1969-1971," Vol. l., No. I., 

DHEW Publication No. (HRA) 75-1150, Public Health & 

Service, Health Resources Administration, National Center 

for Health Statistics, Rockville, Maryland, 1975. 

ICRP77 International Commission on Radiological Protection, 

"Recommendations of the International Commission on 

Radiological Protection," Ann. ICRP, Vol. 1, No. 1, 

Pergamon Press, 1977. 

National Academv - of Sciences - National Research Council, 

The Effects on Population of Exposure to Low Levels of 

Ionizing Radiation, Committee on the Biological Effects 

of Ionizing Radiation, Washington, D.C., 1980. 

APPENDIX C 

9VERVIEW C? '?k'cl!Xl(~L'Es USED TO QUANTIFY 

UNCEKTAINTY IN EKVIR3NMENTAL RISK ASSESSI<EN'I'S 

C.l INTRODUCTION 

The doses and risks attributable to airborne emissions from the 

various facilities and categories of facilities addressed in Volume 

I1 have been estimated using the models and assumptions described in 

this volume. The calculational methods use monitored data 

characterizing airborne emissions and then apply mathematical models 

to estimate the radionuclide concentrations and radiation fields in 

the environment. These calculated values are then used to derive 

radiation doses to individuals exposed to these radionuclides. The 

final products of this exercise are the doses to individuals and 

populations, expressed in units of mrem/yr and person-rem/yr, 

respectively. In addition, cancer risks, expressed in terms of the 

additional lifetime risk to individuals and the number of additional 

cancer fatalities in the exposed populations, are also estimated. 

Rather than using mathematical models to assess impacts, it 

would be preferable to measure the actual impacts directly; i.e., 

radionuclide concentrations and radiation fields in the environment, 

radionuclide concentrations in the various organs of the exposed 

populations, and the increased incidence of cancer, if any, due to 

the exposures. However, this is not possible because the 

radionuclide releases do not generally result in detectable levels 

in the environment or in the exposed members of the population. 

Accordingly, the actual or potential impacts of the emissions must 

be predicted using calculational models. 

The dose and risk estimates provided in this BID for each 

facility or release category should be considered a reasonable 

assessment which does not significantly underestimate or grossly 

overestimate impacts and is of sufficient accuracy to support 

decisionmaking. Since each facility is unique, the models used to 

calculate doses and risks are generalizations and simplifications of 

the processes which result in exposure and risk. In addition, our 

ability to model the processes is also limited to a degree by the 

availability of data characterizing each site and our understanding 

of the processes. 

In Volume 11, doses and risks for each category are presented 

as discrete values; i.e., mrem/yr; person-rem/yr; individual 

probability of a fatal cancer, and number of cancer fatalities per 

year in a population. Each of these calculated values is an 

expression of impact on an individual or small group of individuals 

or on a population as a whole. The values presented, however, are 
of more use to decision-makers when there is some characterization 

of their uricerta-inty. For exampl-e, a srnall impact may be 

calculated; i.e., 1.OE-6 Lifetime risk of cancer for an individual.. 

However, if the uncertainty in this number is several orders of 

magnitude, the real risk of this source of emission may in fact be 

higher than another source of emission which has a calculated risk 

of 1.OE-5 lifetime risk of cancer but has a small degree of 

uncertainty. Alternatively, an upper bound risk of 1.OE-2 lifetime 

risk may be calculated and appear to represent an unacceptable risk. 

However, the actual risk may be an order of magnitude smaller. This 

situation often occurs when, due to limited information and 

uncertainty in the calculational parameters, conservative 

assumptions are used throughout the calculation in order to ensure 

that the risks are not underestimated. 

The Office of Radiation Programs has initiated a quantitative 

uncertainty analysis to supplement the semiquantitative analysis 

provided in Volume I of the BID. 

This appendix summarizes the quantitative uncertainty analysis 

techniques currently under review by the Office. 

C.2 QUANTITATIVE UNCERTAINTY ANALYSIS 

The use of quantitative uncertainty analysis to address 

environmental risks became widespread following the Reactor Safety 

Study (NRC751, and was recommended by the Agency in support of 

environmental risk assessments in 1984 (EPA84). The technique 

results in a range of values of impact rather than a discrete value 

by using a range of values for the calculational input parameters. 

In this way,, the impacts of a given technological activity can be 

bounded and different technologies can be intercompared. In cases 

where probability distributions can be assigned to the value of a 

given set of calculational parameters, the results are expressed as 

probability distributions. Risks can thereby be expressed as "best 

estimate" values, 90 percentile values or 99 percentile values, etc. 

Figure C-l presents an example of the output of such an analysis. 

The results are expressed as a cumulative probability distribution. 

Inspection of the distribution reveals that, in this case, there is 

about a 90 percent level of confidence that the technological 

activity will result in less than 1 mortality per 10,000 years, and 

that the best estimate (i.e., the 50 percentile value) is less than 

0.1 fatality per 10,000 years. 

Though the concept is simple, the implementation and 

interpretation of uncertainty analyses performed in support of 

environmental risk assessment has evolved into an area of 

specialization founded in work performed at Carnegie Mellon 

University (Mo78). The use of quantitative uncertainty analyses in 

support of environmental radiological risk assessment has been 

steadily increasing since its use in the Reactor Safety Study 

(NRC75). Selected uncertainty analyses, which are especially 
* . 

8 

8 

I 

I 

1 

a 

. 

* 0 

I 

0 ,081 .01 .1 1 10 

MORTALITY EFFECTS110.000 YEARS 

Figure C-1. Example of the output of a risk assessment using 

quantitative uncertainty analyses (from Ri83). 
relevant to this Background Information Document, include work 

performed by Hoffman (NUKEG79, NUREG81), Rish (Ri03), and Crick 

(Cr88). 

These applications of uncertainty analysis are currently 

undergoing review to identify the approach most appropriately 

applied to the analyses presented in Volume I1 of this BID. Each 

application uses a somewhat different calculational approach and set 

of input data. The appropriateness of the approaches depends on 

types of risks being calculated and on the level of analysis 

required to support rulemaking. The following describes the 

different approaches being considered and the data requirements. 

C.3 LEVEL OF ANALYSIS 

The results of any risk assessment are uncertain due to the 

following three sources of uncertainty (Cr88): 

(1) Modeling uncertainties 

(2) Completeness uncertainties 

(3) Parameter uncertainties 

Modeling uncertainties pertain to the formulation of 

mathematical models used to predict risk and the degree to which 

they accurately represent reality. One way to address this source 

of uncertainty is to perform the analysis using a set of feasible 

alternative model structures. 

In general, modeling uncertainty is the most difficult 

component to assess since it is often impossible to justify a set of 

plausible alternative models in light of the available data and to 

assign probabilities to these alternatives. To an extent, modeling 

uncertainty is incorporated into the estimates of uncertainty, e.g., 

the uncertainty in risk factors for low-LET radiation includes a 

consideration of the uncertainty in the form of the dose-response 

and risk projection models. On the other hand, as noted in Chapter 

5, the uncertainty in formulation of metabolic models is a serious 

problem in estimating dose conversion factors for many 

radionuclides. Modeling uncertainty for dispersion and pathway 

calculations pose similar problems. As a result, the Agency's 

estimates of uncertainty in radiological risk do not fully reflect 

the contribution of modeling uncertainty. 

Completeness uncertainties are applicable to this BID, as they 

are to all risk assessments. The issue has to do with whether all 

significant radionuclides and pathways of exposure have been 

addressed. For most facilities addressed in this BID, the source 

terms are well characterized and there is little likelihood that a 

significant undetected radionuclide release is occurring. With 

regard to pathways of exposure, the analyses assume that all the 

major pathways of exposure are present at all sites, and it is more 
likely that a pathway has been assumed to be present which in fact 

is not. Accordingly, except for some specific categories of 

emissions, such as C-14 and H-3 emissions from research hospitals, 

this source of uncertainty is not expected to be an important 

contributor to overall uncertainty. 

Uncertainties in the values of the calculational input 

y in input parameters, such as dose and risk 

ion of both parameter and modeling 

of a quantitative uncertainty analysis, 

ined and will be treated in subsequent 

t parameter uncertainty. 

UE TO PARAMETER UNCERTAINTY 

ranges or a set of discrete 

are characterized, their 

e models using a simulation 

ibution representing 

ode1 results. To describe 

est calculated 

the vicinity of 

o, Florida. The 

wer is what is the 
(2) the atmospheric dispersion factor, which is used to 

calculate the average annual airborne radon concentration 

at the receptor location, 

(3) the radon daughter conversion factor, which converts the 

calculated airborne radon concentration to radon 

daughter concentration in working levels (WL), which 

is the parameter that is directly related to risk, 

(4) exposure duration in hours per year, and 

(5) the risk conversion factor, which converts risk expressed 

in WL to probability of cancer. 

The product of each of these parameters, along with appropriate 

unit conversions, results in an estimate of lifetime cancer risk due 

to exposure. Each of the five parameters has some degree of 

uncertainty, which contributes to the uncertainty in the calculated 

risk. 

The source term (Ci/yr) is itself an estimated value which 

varies as a function of time. However, since this is a lifetime 

risk, it is necessary to estimate the uncertainty in the average 

annual release rate over many years. This distinction is important 

because it virtually eliminates the need to explicitly consider 

uncertainties associated with the time-varying nature of the source 

term. If the concern was with the maximum risk to an individual in 

any one year, the time-varying nature of the source term would need 

to be expiicitly addressed. 

Ideally, based on extensive measurements made over the area of 

the stack over prolonged periods of time, the source term could be 

accurately defined. However, the source term has been approximated 

using a limited number of samples and a conservative set of 

assumptions which provides assurance that the real source term has 

not been underestimated. In a quantitative uncertainty analysis, a 

source term probability distribution would be constructed based on a 

close inspection of the measurements and assumptions used in the 

analysis. 

The second calculational parameter is the atmospheric 

dispersion factor, which is used to derive the average annual radon 

concentration at the receptor location. The dispersion factor is 

expressed in units of sec/m3, so that when it is multiplied by the 

release rate in Ci/yr, along with the appropriate unit conversion, 

the result is the average annual radon concentration at the receptor 

location. Uncertainty in the actual location of the nearest 

resident is an important source of uncertainty. 
A second important, and less obvious source of uncertainty, is 

the method used to estimate dispersion. The accuracy of this method 

is provided in Chapter 4. As applied to this particular problem, 

the uncertainties increase due to the non-uniformity of the area 

source term. This could either increase or decrease the risk 

estimate, depending on the location of the receptor relative to 

areas of the pile that are the major contributors to the source 

term. Note that the magnitude of this source of uncertainty is much 

smaller when performing population doses since, as the distance from 

the receptor to the pile increases, the source term behaves more and 

more as a point source relative to the receptor. 

Considering all of these factors, an uncertainty distribution 

is developed for the atmospheric dispersion factor. Note that the 

distribution of the atmospheric dispersion factors for the maximum 

individual and the population risk assessments will differ. 

The third parameter converts radon concentration to radon 

daughter concentration, which is the parameter of interest. The 

uncertainty in this value is well characterized, and constructing a 

reasonable probability distribution for this parameter will be a 

relatively straight forward exercise. 

The fourth parameter, occupancy time, is the fraction of the 

time the individual is located at the receptor location. For 

purposes of this BID, the individual at maximum risk is presumed to 

be a lifetime resident at the presently occupied location that 

results in the greatest lifetime risk. HenEe the value of this 

factor is the average fraction of each day that a resident is 

expected to be within his or her home. The presumption of lifetime 

residence does not have any uncertainty; it is a given condition for 

the assessment. 

The last parameter, the risk factor, relates exposure to risk. 

As discussed in Chapter 6, values for this parameter are based on 

epidemiological data and only apply to large populations. It is 

assumed that the maximum individual has the average 

radiosensitivity, and a risk factor probability distribution is 

developed based on uncertainty in the average risk factor. 

It is apparent from this discussion that in order to perform an 

uncertainty analysis, it is necessary to clearly define the risk 

that is being estimated. Is the risk for a real or hypothetical 

person, is it the maximum or the average risk, and is it the current 

or possible future risk that is of concern? The individuals 

constructing the distributions must clearly understand the 

objectives of the analysis or the resulting distributions will be 

incompatible. 

Upon completion of this exercise, each of the calculational 

parameters will have been assigned probability distributions. These 
distributions are used as input to models that propagat.e the 

uncertainties. 

C.5 TECHNIQUES FOR PROPAGATING UNCERTAINTIES 

The basic approaches used to propagate uncertainties are method 

of moments techniques, or Monte Carlo techniques. Method of moments 

is the standard method for propagating error described in 

fundamental texts on statistics. This method propagates errors by 

calculating a linear combination of the moments for each model 

factor. Since these coefficients depend on the values of the 

parameters, the method is only useful when the range of each 

parameter is small enough that it will not significantly perturb the 

coefficients. Even if these conditions are not met, 

it is possible to establish reasonable estimates of uncertainty 

using this technique. 

The alternative to the method of moments is the use of a Monte 

Carlo, or Monte Carlo type, analysis. This approach can consume 

considerable computer resources but has the potential to yield more 

satisfying results. The technique calculates risk in the same 

manner as described above, except it performs the calculation many 

times, each time randomly selecting an input value from each of the 

probability distributions representing each parameter. The output 

is a risk distribution. The more times the calculation is 

performed, the more complete the results. The number of repetitions 

will determine the precision of the output. The more repetitions 

and the larger the number of calculational parameters treated as 

distributions in the model, the greater the computer resource 

requirements. 

By controlling how the values are sampled from each 

distribution, parameters that are directly or indirectly correlated 

can also be modeled. In addition, by selectively fixing the value 

of individual parameters, the parameters that are important 

contributors to uncertainty can be identified. 

A number of computerized techniques are available to perform 

quantitative uncertainty analysis. Descriptions of these methods, 

provided by Crick (Cr88) and Hofer (Ho85), are being reviewed in 

order to determine which methods are most appropriate for 

quantifying the uncertainty in the risk estimates provided in the 

BID. In addition, a comprehensive guide on uncertainty analysis is 

scheduled for publication in the spring of 1989 (Mo89). The 

publication will be the first comprehensive treatment of this 

subject. 

C.6 PARAMETER DISTRIBUTIONS 

The final and by far the most important issue pertinent to the 

implementation of a quantitative uncertainty analysis is the 
completeness and reliability of the data characterizing the 

distributions of each of the calculational parameters. The number 

of radionuclides, pathways and parameters used in the risk 

assessments (see the AIRDOS input sheets in the Appendix to Volume 

11) is very large. However, through a screening process, the number 

of radionuclides and pathways that require explicit analysis can be 

sharply reduced. 

Once the parameters of interest are identified, it is necessary 

to evaluate how each parameter is used in the risk calculations; 

that is, is it used to calculate risks to a population or an 

individual; and is it used to calculate annual or lifetime risk? 

Once this is determined, probability distributions for each 

parameter, as it is used in the risk calculations, are constructed. 

A number of such distributions have been constructed in the past 

which will facilitate this process (NUREG79, NUREG81, Ri83). In 

addition, it will likely be necessary to elicit subjective 

probability distributions for specific parameters by interviewing 

researchers specializing in each parameter. In order to obtain 

unbiased distributions, formal elicitation techniques, as described 

by Hogarth (Ho75), may be required. 
REFERENCES 

Crick, M.J., et al., Uncertainty Analysis of the 

Foodchain and Atmospheric Dispersion Modules of 

- MARC, NRPB-R184, May 1988 

U.S. Environmental Protection Agency, Proposed 

Guidelines for Exposure Assessment, Request 

for Comments, 49 FR 46304, November 

23, 1984. 

Hogarth, R.M., "Cognitive Processes and the 

Assessment of Subjective Probability Distributions", 

J. of the Am. Statistical Assoc., 70(350), 271, June 

1975. 

Hofer, E., and Krzykacz, B., CEC Study Contract: 

Uncertainty analysis of the computational 

assessment of the radiological consequences of 

nuclear accidents. Final Report to Part I (October 

1984) and to Part I1 (July 1985). 

Morgan, M. G., et al., "Sulfur control 

in coal-fired power plants: A probabilistic 

approach to policy analysis," APCA Journal, 28(10) 

993, 1978. 

Moraan, d. M.G., and Henrion, M., Uncertainty: A 

Guide to Dealing with Uncertainty in Quantitative 

Risk and Policy Analysis, Cambridge University Press 

(scheduled for publication in the summer of 1989). 

U.S. Nuclear Regulatory Commission, Reactor Safety 

Study: An Assessment of Accident Risks in United 

States Commercial Nuclear Power Plants, WASH-1400, 

October 1975. 

NUREG79 u.S. Nuclear Regulatory Commission, A statistical 

analysis of selected parameters for predicting food 

chain transport and internal dose of radionuclides, 

Hoffman, F .O., and Baes, C. F., I11 (eds. ) , 

NUREG81 U.S. Nuclear Regulatory Commission, Variability in 

dose estimate associated with food chain transport 

and ingestion of selected radionuclides, by F.O. 

Hoffman, R.H. Gardner, and K.F. Eckerman, 
Rish, W.R., Mauro, J.M., and Schaffer, S.A., 

"Analyses of Uncertainties in the EPA Ore Body 

Release and River Mode Exposure Pathway Models Used 

as the Bases for Proposed Geologic Repository 

Release Limits", Final Report to Battelle Project 

Manager Division (BPMD) for the Department of 

Energy, June 10, 1983.