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# Risk

# Assessment 

# Methodology

Draft

# Environmental

# Impact Statementfor 

# Proposed ESHAPS

# for 

# Radionuclides

Volume 

# 1 BackgroundInformation 

# 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 20460The Environmental Protection Agency isproposing National Emission Standards for Hazardous 

### Air Pollutants (NESHAPs) for Radionuclides. A Draft Environmental Impact Statement (DEIS) has been prepared insupport of the proposed rulemaking. The DEIS consistsof 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 tolow levels of ionizing radiation, and a summary of theuncertainties incalculations 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

### allthe computer runs used togenerate 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 theDEIS inwhole or inpart are avaiable 

### to all interested persons; an announcement of the availability 

### appears inthe ~edeial Register. Comments on theDEIS should

### be sent in duplicate to:

Central Docket Section 

### (LE131) 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 20460LIST 

### 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 PROGRAMSAND 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 INDOSESAND 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-1LIST OF TABLES

Quality Factor for Various Types of Radiation



## ... Weighting FactorsRecommended by the ICRP



## ............... for Stochastic Risks
Comparison of Customary and SI SpecialUnits forRadiation Quantities



# ............ Target Organs and Tissues Used for Calculatingthe 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 IncidenceRisk 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 Incidenceof (and Mortality from) Leukemia and Bone Cancer

(absolute risk model)



# .............. Site SpecificMortality Risk Per Unit Dose per rad) for Combined Leukemia-Bone and



## .......... Constrained Relative Risk Model
Site Specific IncidenceRisk Per Unit Dose (lo-$ per rad) for Combined Leukemia-bone and



## .......... Constrained Relative Risk Model
Site SpecificMortality Risk Per Unit Dose from Low-LET Radiation, Average for Both Sexes and All Ages


## .......................Site-Specific Mortality Risk Per Unit Dose from Internally DepositedAlpha EmittersAveraged for Both Sexes andAllAges



## .............. 6-33



# .... Risk Estimate forExposures toRadon Progeny 6-38

Lifetime Risk From Radon Daughter Exposure Risk of Lung Cancer Death



# ................. 6-45

Lifetime Risk From Excess Radon Daughter Exposure (adjustedfor a Background Exposure of 0.25WLM/yr) 6-47

Lifetime Risk forExposure of Varying Duration



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

Lifetime Risk forVarying Age atFirstExposure 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 (Chronicx-ray orgamma radiation exposure)


# ...................6-62
Summary of GeneticRisk Estimates Per 106Liveborn of Low Dose Rate, Low-LET Radiation in

### a 30 Year Generation



# .................. 6-63

Genetic Risk Estimates per 106 Liveborn foran Average Population Exposure of 1Rad of High-LET Radiation ina 30-year Generation



## .... 6-64

Radiation Induced Reciprocal Translocations in Several Species



# ............. 6-65

Estimated Frequency of Genetic Disorders ina Birth Cohort Due toExposure of theParentsto 1 Rad per Generation


# ............. 6-68

Increase inBackground Level of Genetic Effects After 30 Generations orMore



# .......... 6-70

Causes of Uncertainty inthe Genetic Risk Estimates 6-71



## ... Possible Effects of InUtero Radiation Exposure 6-79



# ....... Summary of EPAfsRadiation Risk Factors 6-81

Presumed Sources of Food forUrban andAIRDOS-EPA Parameters Used Eor Generic Site Assessments



## ....................A-5

Default Values Used forElement Dependent



# ....................... Factors 

### A-7

Cattle Densities andVegetable Crop Distributions foruse with AIRDOS-EPA



## . A-9

Risk Factor for SelectedRadionuclides 

# . A-11LIST OF FIGURES
Pathways ofAirborne 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,



## ...... GastrointestinalTract, and Blood 5-13
Typical Pattern of Decline ofActivity of
a Radionuclide in an Organ, Assuming an InitialActivity in the Organ and No Additional Uptake of Radionuclideby the Organ


# .................... 5-15

The ICRP Task Group LungModel for Particulates.



# ............. 5-20
Example of the Output of a Risk Assessment



## ... Using Quantitative UncertaintyAnalysis C-3

### 1. 

### INTRODUCTION

1.1 HISTORY OF STANDARDS DEVELOPMENT

In 1977, Congress amended the Clean Air Act (the Act) to address emissions of radioactivematerials. Before 1977, these emissions were either regulated under theAtomic EnergyAct or unregulated. Section 122 of theAct required the Administrator of the U.S. Environmental ProtectionAgency (EPA),after providing public notice and opportunity forpublic hearings (44FR 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 airpollutants under Section 112 of theAct (44FR 76738, December 27, 1979). To support this determination, EPA published a report entitled "Radiological Impact Caused by Emissions of Radionuclides intoAir in the United States, Preliminary Report"

(EPA520/7-79-006, 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 theNorthern District of California pursuant to the citizens' suit provision of theAct (SierraClub v Gorsuch,No. 81-2436WTS). The suit alleged that EPA had a nondiscretionaryduty to propose standards for radionuclides under Section 112 of theAct 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 uraniummines, 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 categorieswere (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 (48FR 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 theseproposed 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). Followingpublication of the proposed standards,EPA held an informalpublic hearing inWashington, D.C.,onApril 28 and 29, 1983. The commentperiod 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 (48FR 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 (SierraClub v


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

- nondiscretionary duty to issue final emissions standards for radionuclidesor 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 (EPA520/1-84- 022-1).

On February 6, 1985, National Emission Standards for Hazardous Air Pollutants (NESHAPS) were promulgated for radionuclide emissions fromDOE facilities,NRC-licensed and non-DOE Federal facilities, and elemental phosphorusplants (50FR 5190). Two additional radionuclideNESHAPS, covering radon-222 emissions from underground uranium mines and licensed uranium mill tailings, were promulgated
onApril 17, 1985 (50FR 15386) and September 24, 1986 (51FR 34056), respectively.

The EPA's basis for the radionuclide NESHAPS was challenged in lawsuits filedby the Sierra Club and the National Resources Defense Council (NRDC).While these suits were under adjudication,the 

### U.S. Court ofAppeals for the District of Columbia issued a decision finding that the EPA's NESHAP forvinyl chloridewas 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 inmany rulemakings on radionuclide emissions.On December 9, 1987, the Court accepted the EPA's proposal to leave the existing radionuclideNESHAPS in place while

### theAgency reconsidered the standards. in 

### the interim, the suits filedby the Sierra Club and theNRDC 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 theAgency's procedures and methods for estimating radiation dose and risk due to radionuclide emissions to the air. Thismaterial 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 internationaladvisory bodies and state and Federal agencies in regard to radiation.



### o Chapter 

### 3

## - A 

### description of what makes radiation hazardous, the evidence that proves thehazard, 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 thebody 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 froma source outside of thebody.
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 theprocedures and information in theprevious 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 factorsused in the dose assessment models. Appendix B describes the mechanics of the life table analysis used to estimate risk. Appendix 

### Cpresents an overview of the quantitativeuncertainty analysis techniques currently under review for use as a method for expanding the semiquantitative uncertainty analysisprovided inVolume 

### I.

Volume I1 contains detailed risk estimates for each source of emissions, which wereperformed according to the procedures given in Volume I. Each chapter inVolume 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 siteparameters relevant to the dose assessment, 

### (3)the results of the dose and risk calculation, along

. with an extrapolationto 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 inVolume 11.Appendix A presents the detailedAIRDOS input sheets used to calculate individual and population doses and risks associatedwith each category. Appendix B presents themethodology used to evaluate the

. costs and effectivenessof 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 uraniummills are addressedbecause they are coveredby NESHAPS. Uranium fuel cycle facilities,high-level waste disposal facilities, coal-fired boilers, and inactive uraniummill tailings sites are addressed because of challenges to previous determinations that they were adequately coveredby other laws. Surface uranium mines, DOE radon, and phosphogypsum stacks are addressed because of challenges to the EPAfslack 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 InactiveUranium Mill Tailings Licensed UraniumMill Tailingso DOE Itadon
o Underground Uranium Mines
o Surface Uranium Mines
o Phosphogypsum Stacks

For each category,Volume I1presents 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 ControlAct (UMTRCA),theAgency 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 otherNESHAP 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, nuclearphysics, and nuclear medicine. By the 192OPs,the use ofx-rays in diagnostic medicine and industrial applicationswas widespread, and radiumwas being used by industry for luminescent dials and by doctors in therapeutic procedures. By the 19301s,biomedical and genetic researcherswere 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 reactionwas 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 (andtheir supporting fuel-cycle facilities) generate electricity,power ships and submarines, produce radioisotopes for research, space, defense, andmedical 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 recognizedmedical specialty in which radioisotopes are used in the diagnosis and treatment of numerous diseases.

o X-rays arewidely used as a diagnostic tool inmedicine 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 PROTECTIONAND THE NATIONAL COUNCIL ON RADIATION PROTECTIONAND MEASUREMENTS Initially, the dangers and risksposed by x-rays and radioactivitywere little understood. By 1896, however, "x-ray burns" were being reported in themedical literature, and by 1910, it was understood that such "burns" could also be causedby radioactive materials. By the 1920rs,sufficient direct evidence (from experiences of radium dialpainters, medical radiologists, and

- 

### miners) and indirect evidence (from biomedical and genetic experiments with animals) had been accumulated to persuade the scientific community that an officialbody shouldbe 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 commissionwas created. Reflecting the use of radiation and

- 

### radioactive materials at the time, thebody was named the InternationalX-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 Commissionwas 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 theirviewpoints before the ICRP, and to act in concert with the Commission in developing and disseminating recommendations on radiationprotection. This suggestion led to the formation, in 1929, 

### of theAdvisory Group. After a series of reorganizations and name changes, this committee emerged in 1964 in itspresent form as the congressionally chartered National Council on Radiation Protection andMeasurements (NCRP).The congressional charter provides for the NCRP to: o Collect, analyze, develop, and disseminate in thepublic interest information and recommendations about radiation protection and radiation quantities, units, and measurements.



### o Developbasic concepts about radiation protection and radiation quantities, units, andmeasurements, and the application of these concepts.

### o Provide ameans by which organizations concerned with radiation protection and radiation quantities, units, and measurementsmay cooperate to use their combined resources effectively and to stimulate thework 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 theNCRP 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 theNCRP function as non-government advisory bodies. Their recommendationsare not binding on any government or user of radiation or radioactive materials.

The first exposure limits adopted by the ICRP and theNCRP (ICRP34,ICRP38, andNCRP36) established 0.2 roentgen/daylas 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 theblood, 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 ofmoderate or low doses. Therefore,the aim of radiation protectionwas 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 0.2R/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 nomention of exposure of the general populace.

BY 

### the end of World War 11, thewidespread use of radioactive materials and scientific evidence of genetic and somatic effects at lower doses and dose rates suggested that the radiation protection recommendationsof theNCRP and the ICRPwould have to be revised downward.

The NCRPfsrecommendation 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 itsposition on appropriate
new limits.These limitswere largely accepted by the ICRP in its recommendations of 1950 and formally issuedby theNCRP in 1954

(ICRP51,NCRP54). Whereas the immediate effectwas 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 theweighting factors that shouldbe applied to radiations of differing types and energies. TheNCRP advocated the use of the "rem" to express the equivalence in biological effect between radiations of differing types and energy.2Although 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 shouldbe 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 conceptwas 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 organswere considered critical and were limited to the same exposure as thewhole body.
Third, the NCRP recommendations included the suggestion that individualsunder 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 tomatter 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. Forbeta 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 thisparticular recommendation is interesting, as it reflects clearly the lim'ted knowledge of the times. The scientific evidence indicateda clear relationship between accumulated dose and genetic effect. However, this evidencewas obtained exclusively from animal studies that had been conductedwith doses ranging from 25 to thousands of rads. There was no evidence from exposure less than 25 rads accumulated dose, and the interpretationof 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 forprolonged periods to the permissible exposure limit (1.0R/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 radiationworkers. 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 replacedby the concept of a maximum permissible dose. The change in terminology reflected the increasing awareness that any radiation exposuremight involve some risk and that repairmechanismsmight be less effective than previously believed. Therefore, the concept of a maximum permissible dose (expressedas dose per unit of time) was adopted because it better reflected the uncertainty in our knowledge than did the concept of tolerancedose. 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 exposuremight 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 asALARA (aslow 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 survivorsof the atomic bombs dropped on Japan. This evidence suggested that genetic effects and long-term somatic effectsweremore important thanpreviously 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 occupationaldosefor who1.ebody exposure and themost 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 organswas set at 30 rems/year; internal exposures were controlledby a comprehensive set ofmaximum permissible concentrationsof radionuclides in air and water based on the most restrictive case of a youngworker; and recommendations were included for some nonoccupational groups and for the general population (for the first time).
The lowering of the maximum permissiblewhole-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 asmuch as 750 rads (15 rads/year times 50years) was toomuch. Lowering themaximum 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 inmost instances, particularly with the additional operational flexibilityprovided 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 recommendationthat accumulated gonadal dose to age 30be limited to 

### 5 rems from sources other than natural background and medical exposure. As an operational guide, the NCRP recommendedthat themaximum dose to any individual be limited to 0.5 rem/year,with maximum permissiblebody 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 reasonablypossible, 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 inconclusivewith 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 ICRPmade the following similar recommendations:o Limit the dose to the whole-body, redbonemarrow, 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 0.17remper year.
The scientific evidence and the protection philosophy onwhich the above recommendationswere 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 limitswas complicatedby 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 individualswho 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 ICRPmade a fundamental change in its recommendationswhen 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 forworkers, but included a recommendationthat chronic exposures of the generalpublic from all controllable
sourcesbe limited to 0.1redyear.

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 (approximately0.5 redyear) are equated with risks in safe industries, given as 1.0 

### E-4 annually.At themaximum 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, theNCRP revised its recommendations to be comparable with those of the ICRP (NCRP87). TheNCRP adopted the effective dose equivalent concept and its related recommendations regarding occupational and nonoccupational exposures to acceptable levels of risk. However, theNCRP 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 tobe elaborated.

TheNCRP recommendations in (NCRP87) for occupational exposures 

### - correspond to the ICRP recommendations. In addition, the relevant nonoccupational exposure guidelines are:

o 0.5rem/year effective whole-body dose equivalent, not including background ormedical radiation, for individuals in thepopulation when the exposure is not continuous.

- 

### o 0.1rem/year effective whole-body dose equivalent, not including background ormedical radiation, for individuals in thepopulation when the exposure is continuous.

o Continuous use of a total dose limitation system based on justificationof every exposure and application of the "as low as reasonably achievable" philosophy.

TheNCRP equates continuous exposure at a level of 0.1rem/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 thepermissible limitswill necessarily be higher than thewhole-body limit in inverse ratio for a particular organ to the total risk forwhole-body exposure.

In response to EPh's proposed national emission standards for radionuclides, theNCRE suggested that since the 0.1 rem/year limit is the limit for all exposures from all sources (excluding natural background andmedical radiation), the operator of any site responsible for more than 25 percent of the annual limit be required to assure that the exposure of themaximally 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 Presidentto 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 concernedwith radiation protection and acted as a coordinatingbody 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 radiationmatters, directly or indirectly affecting health, including guidance for all Federal Agencies in the formulationof 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 agencieswere approvedby the President in 1960. Based largely on the work and recommendationsof 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 yearsbeyond age 

### 18.

o Skin of whole body and thyroid--not to exceed 10 rems in 13weeks 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 

### 13weeks or 15
rems per year.
Although these levels differ slightly from those recommendedby 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 recommendedby 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 significantbenefit 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 compliancewith the limit when doses to all individualsare unknown. Exposure of thispopulation samplewas 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 considerationof 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 shouldbe no radiation exposurewithout an expectation of benefit and that "every effort shouldbe made to encourage the maintenance of radiation doses as far below this guide aspracticable." 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 responsewas 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 radiationhazards in uranium mining (FRC67).The need for such guidancewas clearly indicated by the epidemiological evidence that showed a higher incidence of lung cancer in adult males who worked in uranium mines comparedwith the incidence in adult males from the same locationswho had not worked in themines. The guidance established specific exposure limits and recommendedthat all exposuresbe kept as far below the guide limits as possible. The limits chosen represented a tradeoffbetween the risks incurredat 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 theAdministrator of the 

### U.S. Environmental ProtectionAgency. 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 exposurewas 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 routinex-ray examinations. It also establishes the requirement that all x-ray exposuresbe based
on clinical indication and diagnostic need, and that all exposure of patients shouldbe 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 guidancewas adopted as Federal policy in 1987 (EPA87).

The Environmental ProtectionAgency has various statutory authorities and responsibilitiesregarding regulation of exposure to radiation in addition to the statutory responsibility to provide Federal guidance on radiation protection. EPA's standardsand regulations for controlling radiation exposures are summarizedhere.
Reorganization PlanNo. 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 operationsof 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

(excludingradon and its daughters) to 25mrems to the whole body, 75mrems to the thyroid, and 25mrems 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 combinedwith 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 standardtook 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, theAgency 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, operationsmust 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 limitswere derivedby summing, overlong timeperiods, the estimated risks to all persons exposed to radioactive materials released into the environment. The uncertainties involved in estimating theperformance 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 theAtomic Energy Act of 1954, as amended, and the Toxic SubstanceControl Act, the EPA is developing proposed environmental standards for the land disposal of low-level radioactivewastes and certain naturally occurring and acceleratorproduced radioactive wastes. Theproposed standards will establish

(1)exposure limits forpre-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) groundwaterprotection requirements. Theproposed regulations are scheduled to be published in the Federal Register in late 1988 (Gr88).

Under the authority of the Safe DrinkingWater Act, the EPA has issued interim regulations covering thepermissible levels of radium, gross alpha andman-made beta, and photon-emitting contaminants in community water systems (EPA76).The limits are expressed in picocuries/liter. The limits chosen forman-made beta and photon emitters equate to approximately 4mrems/year whole-body
or organ dose to themost exposed individual.

Section 122 of the Clean Air Act amendments of 1977 (Public Law 95-95) directed theAdministrator of the EPA to review all relevant information and determine if emissions of hazardous pollutants into air will cause or contribute to airpollution that may reasonablybe expected to endanger public health. In December 1979, EPA designated radionuclides as hazardous airpollutants 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 HazardousAir 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 uraniummill tailings, were promulgated on April 17, 1985 and September 24, 1986, respectively (FR85b,FR86).

2.5 NUCLEAR REGULATORY COMMISSION

Under the authority of theAtomic 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 protectspublic health and safety. The Federal guidance on radiation protection applies to theNRC; therefore, the NRC must assure that
none of the operations of its licensees exposes a member of the public tomore than 0.5 rem/year. The dose limits imposedby the EPA's standard for uranium fuel cycle facilities also apply to the fuel cycle facilities licensedby theNRC. 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.

TheNRC 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. TheNRC 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 theNRC's large source and special nuclear material and production and utilization facilities. TheNRC's regulations require an analysis of probable radioactive effluents and their effects on the population
near fuel cycle facilities. TheNRC 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. Forpractical purposes, the NRC adopted themaximum permissible concentrations developed by the NCRP to relate effluent concentrations to exposure.

In the 19701s,theNRC formalized the implementation of as low as reasonably achievable exposure levelsby 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 

### 5mrem limit applies to themost exposed individual actually living in the vicinity of the reactor and refers to whole-body doses from external radiationby air pathway (NRC77).

2.5.2 Byproduct Material Licenses

TheNRC's licensing and inspection procedure for byproduct material users is less uniform than that imposed on major fuel cycle licensees for two reasons: (1)themuch larger number of byproduct material licensees, and (2)their much smallerpotential for releasing significant quantities of radioactive materials into theenvironment. Theprelicensing assurance procedures of imposing design reviews, operating practices, and license condLtbons prior to construction and operation are similar.

Theprotection afforded thepublic from releases of radioactive materials from these facilities can vary considerably because of three factors. First, the requirements that theNRC imposes for monitoring effluents and environmental radioactivity are much less stringent for these licensees. If the quantity of materials handled is small enough, theNRC 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 theparticular licensee, this area has been defined as the nearest inhabited structure, as the boundary of the user's property line, as the roof of thebuilding where the effluents are vented, or as themouth of the stack of vent. Finally, not all users are allowed to reach 100 percent of themaximum permissible concentration in their effluents. In fact, the NRC has placed as low as reasonably achievable requirements onmany of their licenseesby limiting them to 10 percent of themaximum 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 licensedby 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 itspredecessor, 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 0.5rem/yearwhole-body and 1.5 rems/year to any organ are being observed. The DOE also has a requirement that all dosesbe 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 guidelinesby 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 (similarto 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 acorresponding reductyion in the effluents (including airborne and liquid radioactive materials) released to the environment.

As 

### a continuation of thisprogram, 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 FEDERALAGENCIES

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 forMedical Devices and Radiological Health

Under the Radiation Control Act of 1968, the major responsibility of the Center forMedical Devices and Radiological Health in the area of radiation protection is the specificationof performance criteria for electronic products, includingx-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 ofminers to radon and its decay products and other (nonradiological) pollutants in mines. The MSHA has adopted the Federal guidance for exposure of uraniumminers (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 forworkers 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 notimposed the requirement to keep aLL doses as ,lowas 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. Forpractical purposes, the DOT has implemented its authority through the specificationof 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 compliancewith 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 STATEAGENCIES

States have important authority for protecting the public from the hazards associatedwith ionizing radiation. 

### In 26 states, the states have assumedNRC'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 regulatemore than 11,500 users of radiation and radioactive materials, are bound by formal agreements to adopt requirements consistent with those imposedby 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 ProtectionAgency, "Radiation Protection Guidance for FederalAgencies: Underground Mining of Uranium Ore," Federal Register 36(132), July 9, 1971.

U.S. Environmental ProtectionAgency, National Interim Primary DrinkingWater 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 ProtectionAgency, "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 InactiveUranium Processing Sites," Federal Register 48(590), January 5, 1983.

U.S. Environmental ProtectionAgency, "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, 1986FRCGO 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., "Overviewof EPA1sEnvironmental Standards for the Land Disposal of LLW and NARM Waste-1988," ORP/EPA, for presentation at DOE'S 10thAnnual LLW Management Conference, Denver, Colorado, August 30 

### - September 1, 1988.

ICRP34 InternationalX-Ray and Radium Protection Commission,



### - "International Recommendations for X-Ray and Radium Protection," British Journal of Radiology 7, 695-699, 1934.

ICRP38 InternationalX-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.

NCRP5 4 National Committee on Radiation Protection, Permissible Dose From External Sources of Ionizing Radiation, National Bureau of Standards Handbook 59, 1954.

NCRP

### 59 National Committee on Radiation Protection, Maximum PermissibleBody Burdens and Maximum Permissible

### Concentrations of Radionuclides inAir and in Water for Occupational Exposure,National Bureau of Standards Handbook 

### 69, 1959.



### NCRP71 

### National Council on Radiation Protection and Measurements,Basic Radiation Protection Criteria, NCRP Report No. 39, 1971.

NCRP8

### 4 National Council on Radiation Protection and Measurements, Control ofAir Emissions of Radionuclides, September 18, 1984.

NCRP87 National Council on Radiation Protection and Measurements, Recommendations on Limits for Exposure to IonizingRadiation, 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 toproduce 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~lesinto 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 (suchas 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 andmutagenicity 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 thebody. 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 thisphenomenon 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 literaturehas 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 tobe 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 ismore 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, themetabolism 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 andNagasaki, Japan, were exposed to whole-body external radiation doses of 0 

### to more than 

### 200 rads.' An internationalgroup has been observing the population since 1950. Themost recent reports published by this group (Ka82,Wa83) indicate that an increase in cancer mortality has been shown formany 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 investigatorshave 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.eorgans irradiated (e. g., breast, brain, thyroid, and probably salivary glands, skin, bone, and pelvic organs) following doses ranging from less than 10 tomore 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 someminers 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 formedical 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 theperson'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.likelyto 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 radionuclidesmay 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 tobase 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-raysor gamma rays).

A number of researchers have induced transformations in mammalian tissue culture, including the embryo cells ofmice and hamsters (Bo84,Ke84, Ha84, Gu84). Chromosome aberrations in cultured human peripheral lymphocytes have been demonstrated at Rn-222 alpha doses of about 48mrads/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. Genemutations 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 chromosomehas 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 somild they are not noticeable, and
somemutagenic 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 (becausethe 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 ofmutations 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 thesemutations 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/ingestedRn-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 (cellsof 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 ismuch 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 aremost 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, thismay 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 themechanisms 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 ofboth 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., thepercentage 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 

### SUWYOF 

### EVIDENCE THAT RADIATION ISA 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-cproperties 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 (between8 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.

### REFERENCES

Awa, A.A., et al., Relationship Between Dose and Chromosome Aberrations in Atomic Bomb Survivors, Hiroshima and Nagasaki, RERF TR 12-77, Radiation Effects Research Foundation, Japan, 1978.

Beebe, G.W., Kato, H., 

### and Land, C.E., Mortality Experience of Atomic Bomb Survivors, 1950-1974, Life Span Study Report 8, RERF TR 1-77, Radiation Effects Research Foundation, Japan, 1977.
Borek, C., et al., "Inhibition of X-ray and Ultraviolet Light-Induced Transformation in Vitro by Modifiers Poly (ADP-ribose) Synthesis," Radiation Research 99:219-227, 1984.
Brandom, W.F., et al., "Somatic Cell Chromosome Changes in Humans Exposed to 2j9Plutonium and 222Radon," Contract No. E (29-2) -3639, Progress Report, July 1, 1976, through September 

### 30, 1977, Department of Energy, Washington, D.C., 1977.
Carter, C.O., "Some Rough Estimates of the Load From Spontaneously Arising Genetic Disorders,"paper submitted to the U.N. 

### Scientific Committee on the Effects of Atomic Radiation, September 1980.

Guerrero, 

### I.,Villasante, A., Corces, V., 

### and Pellicer A., 

### "Activation of a c-K-ras Oncogene by 

### Somatic Mutation in Mouse Lymphomas Induced by Gamma Radiati~n,~ Science 225, 1159-1162, September 14, 1984.

Han, A., Hill, C.K., and Elkind, M.M., "Repair Processes and Radiation Quality in Neoplastic Transformation of Mammalian Cells," Radiation Research 99, 249-261, 1984.
Hoffman, F.L., "Radium (Mesothorium) Necrosis," 

### J.A.M.A. 85, 961-965, 1925.
International Meeting on the Toxicity of Thorotrast and Other Alpha-Emitting Heavy Elements, Lisbon, June 1977, Environmental Research 18, 1-255, 1979.

### Kato, 

### H., and Schull, 

### W.J., '"Studies 

### of the Mortality of A-Bomb Survivors," Report 7 Part 1, "Cancer Mortality Among Atomic Bomb Survivors, 1950-78," Radiation Research 90, 395-432, 1982.
Kennedy, A.R., and Little, J.B., "Evidence That a Second Event in X-ray Induced Oncogenic Transformation in Vitro Occurs During Cellular Proliferation," Radiation Research 99, 228-248, 1984.
King, R.C., Genetics, Oxford University Press, New York, 1962.
Ludewig, P., and Lorenser, E., "Untersuchung der Grubenluft in den Schneeberger Gruben auf den Gebalt an Radiumemanation," Zschr. f. Phys. 22, 178-185, 1924.
Marie, P., Clunet, J., and Raulot-Lapointe, G., "Contribution a L1etudedu Developpement des Tumeurs Malignes sur les Ulceres de Roentgen," Bull. Assoc. Franc. Etude Cancer, 3, 404, 1910, cited in UNSCEAR77
Marie, P., Clunet, 

### J., and Raulot-Lapointe, G., "Nouveau cas de Tumeur Maliqne Provoqueepar une Radiodermite Experimentale ~hez let Rat ~ianc," Assoc. Franc. Etude Cancer 5, 125, 1912, cited in UNSCEAR77.

Ma29 Martland, H.S., and Humphries, E.E., "Osteogenic Sarcoma in Dial PaintersUsing Luminous Paint," Arch. Pathol. 

### 7, 406-417, 1929.
Mo67 Morgan, K.Z., and Turner, J.E., Principles of Radiation Protection, John Wiley and Sons, Inc., New York, 1967.

Mu30 Murphy, D.P


## ., and DeRenyi,

### M., "Post-ConceptionPelvic Irradiation of the Albino Rat (Musnorvesicus): Its 

## - Effect on the Offspring,'' Surgery, Gynecology and Obstetrics 50, 861-863, 1930.

NAS72 National Academy of Sciences 

### - National Research Council, The Effects on Populations of Exposures to Low Levels of IonizinqRadiation, Report of the Committee
on the Biological Effects of Ionizing Radiations (BEIR Report), Washington, D.C., 1972.

### NationalAcademy of Sciences 

### - National Research Council, The Effects on Populations of Exposures to Low Levels of lonizingRadiation, Committee on the Biological Effects of IonizingRadiation, Washington, D.C., 1980.
Otake, M., and Schull W., Mental Retardation in Children Exposed in Utero to theAtomic Bombs: A Reassessment, Technical Report REW 

### TR 1-83, Radiation Effects Research Foundation, Japan,April 1984,

Pohl-Ruling, J., Fischer, P., and Pohl, E., "Einfluss Erhohter Umweltradio-Aktivitatand Beruflicher Strahlen-Belastungauf die 

### Chromosomen-Aberrationenin den Lymphocyten des Peripheren Blutes," Tagungsaber, 0sterr.-Ungar, Tagung uber biomedizin, Forschung, Seibersdorf, September 1977.
Pohl-Ruling, J., Fischer, P., and Pohl, E., "The Low-Level Shape of Dose Response for Chromosome Aberrations," IAEA-SM-224/403, presented at Internatinal Symposium on the Latent Biological Effects of Ionizing Radiation, IAEA,Vienna, 1978.
Rugh, R., "VertebrateRadiobiology: Embryology,"

# a. Rev. Nucl. Sci. 3, 

### 271-302, 1953.
Schull,W.J., Otake, M., and Neel, J.V., "Genetic Effects of theAtomic Bombs: A Reappraisal," Science 213, 1220-1227, September 1981.
Stettner, E., "Ein Weiterer Fall einer Schadingung einer Menschichen Frucht durch Roentgen Bestrahlung, Jb. Kinderheitk," Phys. Erzieh. 95, 43-51, 1921.

Storer, J.B., "Radiation Carcinogenesis," Cancer 1, pp. 453-483. F.F. Becker, editor, Plenum Press,New York, 1975.
Trimble, B.K., and Smith,M.E., "The Incidence of Genetic Disease and the Im~act on Man of an Altered

~~ ~-~

~ * 

### Mutation Rate," Canadian Journal of Genetic Cytology, 19, 375-385, 1977.

Tuscany, R., 

### and Klener,V., "PoklesEuploidie v Bunkach Kostni Drene osob sVnitrni Kontaminaci Nekterymi Radioisotopy," Cisk. Fysiol. 12, 391, 1963.

### Uh21 

### Uhlig, 

### M., "Uberden Schneeberger Lungenkrebs, Virchows," Arch. Pathol.Anat. 230, 76-98, 1921.

UNSCEAR58 United Nations Scientific Committee Report on the Effects of Atomic Radiation, Official Records: Thirteenth Session, Supplement No. 17(A/3838),United Nations, New York, 1958.

UNSCEAR77 United Nations Scientific Committee on the Effects of Atomic Radiation, Sources and Effects of Ionizinq Radiation, United Nations, New York, 1977.

UNSCEAR82 United Nations Scientific Committee on the Effects of Atomic Radiation, IonizingRadiation: Sources and Biological Effects, United Nations, New York, 1982.
Up75 Upton, A.C., "Physical Carcinogenesis: Radiation-- History and Sources,"Cancer 1, 387-403, F.F. Becker, editor, Plenum Press, New York, 1975.

V002 Von Frieben,A., "Demonstration Linas Cancroids des Rechten Handruckens, das sich nach Langdauernder Einwirkung von Rontgen-strahlen Entwickelt Hatte," Fortschr. Geb. Rontgenstr. 6, 106, 1902, cited in Up75

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 ofparticulates or gases, form a plume that disperses down wind (Section4.2). These radionuclides in the air can directly affect people in two ways: through external dose causedby photon exposure from the plume, or through internal dose resulting from radionuclide inhalation. As the airborne radionuclidesmove from the point of release, they

(especiallythose inparticulate form) deposit on ground surfaces and vegetation as a result of dry deposition and precipitation scavenging (Section4.3). 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 asmilk and meat (Section4.4). 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 calculatedusing 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 referencesHa82, Ti83, and NCRP84 for a more detaileddescription 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 theirpoint 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, ormixed layer, is created that controls the

# IAIRDISPERSION

EXTERNALDOSE INTERNALDOSE

Figure 4-1. Pathways of airborne radionuclides into the environment.dispersion of the released radionuclides. The depth and dispersion properties of themixed layer, which are highly variable over short periods of time, are controlledby two sources of turbulent effects: mechanical drag of the ground surface and heat transfer into or from the boundary layer. Themechanical drag of the ground surface on the atmosphere creates a shear zone that can produce significant mechanical mixing. Themechanical 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 (dimensionor thickness) of themechanicalmixing
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 themixed layer's turbulent structure and thickness. Solar heating createshuge rising bubbles
or thermals near the ground. These large bubbles produce turbulent eddies of a much larger scale than those from themechanical drag of the ground surface. With strong solarheating 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 otherpollutants) 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, themathematical complexities and difficulties 

### of
a direct solution to the turbulent dispersion problem are profound andbeyond thepractical 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 quantitiesgovern dispersion: wind direction, wind speed, and stability. Wind direction determines which way a plume will be carriedby 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 otherproperties 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 thirdmeteorological. quantity, categorizes the behavicr of a parcel of air when it is adiabaticall-y (without heat transfer) displaced in a vertical direction. If the displacedparcel would be expected to return toward its original position, the category is stable; if it would continue tomove away from its original position, the category is unstable. Under conditions of neutral stability, the parcel would be expected to remain at its new elevationwithout 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 surfaceat night when the sky is clear and wind speeds are low. Dilution due to vertical mixing occursmore rapidly with increasing distance under unstable conditions than under stable
ones. Stability categories range from A (veryunstable) to D

(neutral)to G (verystable)



## .
A table of joint frequencies (fractionsof time) for each combination of stability, wind direction, and wind speed is the startingpoint for any assessment of long-term atmospheric dispersion. These data are usually obtainedby the analysis of long-term observations from weather stations or from site- specific meteorological facilities.

4.2.2 Air DispersionModels

EPA uses an empirical Gaussianmodel formost radionuclide dispersion calculations. The model also considers such processes as plume rise, depletion due to deposition, and radionuclide ingrowth and decay.

Gaussian PlumeModel

The basic workhorse of EPA dispersion calculations is the Gaussianmodel. Several reasons why the Gaussianmodel is one of the most commonly used are quoted below (Ha82):

"(1) Itproduces results that agree with experimental data
as well as any model.

"(2) It is fairly easy to performmathematical 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 constantsK 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 itsway 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 ismaximum at ground level and decreaseswith 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, characteristicof a full Gaussian distribution. Actually, the vertical dispersion is limitedby the ground surface below and any inversion lid above the release (seeFigure 4-2). An inversion lid is defined by the altitude in the atmospherewhere the potential temperature begins to increase with increasingheight, thus limiting the volume of air available for diluting releases.

At largedistances 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 concentrationdecreases monotonically with distance from the release point. For an elevated release, the ground-level concentration increases, reaches amaximum value, and then decreases with increasing distance from the releasepoint.
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 distancex(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

# -----

~~MIDIATIomm ~nos mnAycs IORWNOANDYlXlWO IUMI~ORYLV YlXRDI L10AIIICTPIOCILTI I I t

Figure 4-2. Vertical concentration profiles forplume 

### vs downwind distance from releasewhere h (m) isthe lidheight (meters), and the other quantities are 

### L the same asbefore.

Plume RiseModel

Verticalmomentum orbuoyancy can cause 

### a plume torise toan effectiveheight that is several 

### times thephysical height of the release. The momentum fluxof 

### a release isproportional tothe product of thevolume flow rate andthe vertical exitvelocity, while thebuoyancy flux isproportional totheproduct of thevolume flow rate and the difference between thetemperatures of the release gases and the ambient air. Momentum rise is initiallydominant for most plumes, even though buoyant rise may become themore important process atlargerdistances. In any case, plume rise increases with distance from therelease point; the effective height of theplume may not reach 

### a limitingvalue until theplume is several kilometers from thepoint of release.

Plume Depletion Model

As radionuclides intheplume aredispersed, their activity is depletedby dry deposition andprecipitation scavenging. The rate of plume depletion due todry deposition andprecipitation scavenging isproportional tothe deposition rate (see Section 4.3). EPA's Office of Radiation Programs uses a source depletionmodel which considers the shape of the vertical concentrationprofileto be unchanged by depletion. Depletion due todeposition generally does not cause more than half of the released activitytobe removed at

### a distance of 80 km. Depletion by precipitation scavenging occurs only during periods of precipitation.
Radiological Decay and Ingrowth
Radiological decay can alsoreduce the radionuclide concentration intheplume. A typical elapsed timefortraverse between thepoint of release and a receptor located 80 km away is about 5 hours. Thus, only nuclides with short half-lives would be appreciably depletedby radiological decay. For example, Ar-41, which has a 1.8hour half-life, decays toabout 15 percent of its original activity in5 hours. When 

### a released radionuclide is a parent for other radionuclides ina chain, those decay products will become part of theplume's activity even though they were not released by the source. For example, Cs-137 

### istheparent of Ba-137m, which has a half-life of about 2.6 minutes. The Ba-137m activitywould reach 90 percent of that of the Cs-137 inabout 8.5 minutes, the

### time required ata typical wind speed of 5 m/s forthe releasetotravel about 2.5 km. For many nuclides, the radiological effects associated with exposure todecay products areatleastas important asthose from exposure totheparent. For example, theexternal photon dose from a release of Cs-137 is entirely due t.o photons from its decay product Ba-13711.

4.2.3 Uncertainties inAtmospheric Dispersion Modeling

EPA must deal with several uncertainties in itsmodeling of atmospheric dispersion. Two basic considerations

### contribute to these uncertainties. The first involves the parameters that enter into themodel and how well they are known or can be determined for
a particular situation. The presumption is that the basic assumptions forwhich themodel was developed are satisfied and that the uncertainty of predicted concentrations dependsprimarily 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 forwhich the model was developed. Such usemay 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 dominatedby hills and valleys.

In regard to the first consideration, the authors of NCRP84 concluded that the appropriatebasic parameters, such as wind speed and direction,,can be determined accurately enough so that they are not major contributorstomodel uncertainty. However, the uncertainties associatedwith derived parameters (such as stability class) or lumped parameters (suchas 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 risemodel to within a factor of about 1.4 (NCRP84). From equations 4-1 and 4-2, it is clear that when 0,ismuch 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 hehas little effect on the calculated concentration.

Little and Miller (Li79and Mi82) have surveyed a number of validation studies of atmospheric dispersionmodels. Although these studiesprovide limited data, they indicatean uncertainty of approximately a factor of 2 for annual average concentrations for locationswithin 10 km of the release and approximately a factor of 4 (77percent of their samples) to 10 (92percent 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 andvalleys, 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 contactwith it and "wet" when the transfer is first from theplume to precipitation and then from theprecipitation to the ground or vegetation surfaces.

4.3.2 Dry Deposition Model

Dry depositionmodels generally relate the surface deposition flux to the air concentration at some reference height, typically 1meter 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 fixedvalue, it actually represents the result of a diverse combination of effects.

4.3.3 Wet DepositionModel

Wet deposition models relate the flux due to precipitation scavenging to the concentration in theplume. Since the activity scavenged from theplume by an element of precipitation ispresumed to remainwith theprecipitation element until reaching the ground surface, the deposition flux is proportional to the total.wetted activity in a vertical segment of theplume (ci/m2). The resulting equation can be expressed as



### - 

## w = 

### h,,

# xL (4-4)



### - where W is the surface flux (Ci/mzs),Xisthe 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 interactionsbetween precipitation and theplume. Because the deposition flux is proportional to the vertically integrated concentration (i.e.,the total activity in a column of unit ground surfacearea), it is independent of the effectiveheight of the release. Raising the effective height of a releasemay lower the dry deposition flux but leaves the flux resulting from precipitation scavenging unchanged.

4.3.4 Soil ConcentrationModel
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 canbe 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 concentrationsat 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 canbe expressed as

where C, is the soil concentration (Ci/kg)and P is the areal density of dry soil (kg/m2)for the plowed ormixed soil layer.

The value of t,, the deposition accumulation time, is typically in the range of 20 to 100years. For nearby individual assessments, t, is chosen to correspond to the expected operational life of the facility. If 

### EPA considers it likely that the facilitywould 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 affectedby the choice of t,. For collective (population) assessments, a value of 100years is used for t,. This value corresponds to establishing a 100-year cutoff for the time following a releasewhen any significant intake or external exposure associated with deposition on soilmight 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 radionuclidesfrom 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, (theequilibrium distribution coefficient relating 

### the ratio of the radionuclide concentration in soil water to that on soil particles) for Cs is from 36.5 to 30,000ml/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 itsvalidation do not exist. Since the choice of appropriate values for 

### h, is so uncertain, EPA has used 0.2y-'as a genetal nominal value (thegeometricmean of 

### h, for Put,I-,Csi,and Sr2+ions is 1.2x10-'y-'using Hoffman and Baes median data values) and a value of 0.1yT1for 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 forwet removal modeling is probably best regarded as an order of magnitude estimation procedure" (Ha82). Actually, much of the wide range of values reflectsmeasurement 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 radionuclidesmay 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 canbe transferred to foods such asmilk and meat.

4.4.2 Concentration inVegetation

The radionuclide concentrations inplants 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 representedby the expression

where

### y was found to range between 

### 2.3 and 

### 3.3m2/kgwhen 

### Y, is expressed in kg/m2,dry. Since the product 'yY, is generally less than 1.0,formany practical purposes equation 4-8 can be approximated as 

### -

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

### y which showsmuch less environmental variation than f, and Y, do separately. Note that Y, is the total vegetative yield which can be several times the edibleportion 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. (Avalue of 1.0 is used for all crops in AIRDOS-EPA.)

The value for 

### h, is the sum of

### h,the radionuclidedecay 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 thisprocess 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 fromboth processes isGenerally, 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 concentrationC, (Ci/kg)in animal feed, the concentration inmeat 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). 

### Inpractice,
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 tomilk and the otherparameters 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 integrationand 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 sourceuse of foods grown in proximity to the release location, the fraction of an individual's food that is home produced and other factorsthat can strongly influencethe significance of the foodpathway. Unfortunately, generally useful validation studies to quantify the substantialuncertainties in the food chain have not been made. References such as NCRP84, Ti83, Mi82, and Li79 cite ranges for someparameters and make limited model uncertainty estimates but do not make quantitative evaluations of the uncertainties for the ingestionpathway taken as a whole.

EPA has chosen a factor of 10 as a reasonableupper 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 sigmavalues for a log normal distribution, thecombined uncertainty for thepathway (depositionand intake of radionuclides from food) is a factor of 

### 26.' EPA has rounded this value to 

### 30 as an estimate of the overall foodpathway uncertainty factor.

4.5 CALCULATING THE ENVIRONMENTAL CONCENTRATION OF RADIONUCLIDES: THE AIRDOS-EPA CODE

4.5.1 Introduction

Environmental concentrationsof radionuclides calculatedby EPA may be site specific, meaning that available data relevant to the site are incorporated into the assessment. Or an assessmentmay be generic; that is, an assessment of a hypothetical facility at a location considered an appropriatepossibility 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, EPAmakes both individual and collective (population) assessments. Thepurpose of the individual assessment is to assess doses and lifetime risk to individuals living near a facility. EPAfsassumption is that these individuals reside at the same locationmuch 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 longperiod of time under the assessed conditions. EPA's collective (orpopulation) assessments evaluate doses and risks to a population that may be regional (typicallyup 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 thepopulation per year of facility operation.

4.5.2 AIRDOS-EPA

EPA has used theAIRDOS-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 radioactivityvia 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 itspreparation can be facilitated by using the preprocessor PREPAR (Sj84). Appendix A of this document summarizesmany of the default values and assumptionsused in EPA's assessments.

AIRDOS-EPA calculates atmosphericdispersion for radionuclides released from one to six stacks or area sources. Radionuclide concentrations in meat, milk, and freshproduce are estimatedby coupling the deposition rate output of the atmospheric dispersion models with the Regulatory Guide 1.109 (NRC77)terrestrial food chainmodels. Radionuclide concentrations for specified distances and directions are calculated for the following exposurepathways:

(1)immersion in air containing radionuclides, (2)exposure to ground surfaces contaminatedby deposited radionuclides, 

### (3) inhalation of radionuclides in air, and (4) ingestion of food in the
area. The codemay be used to calculate either annual individual exposures or annual population exposures at each grid location. For either option,AIRDOS-EPA output tables summarizeair 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 directionsproceeding 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 foodproduction 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 coveredby the population and agricultural data. Choosing a suitable set of grid distancesmay 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 (2shown) and 16 directions 

### (5shown)
Figure 4-3. Circular grid system used by AIRDOS-EPA.EnvironmentalP,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 thosepathways 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 customarilyused.

Source Considerations

Point sources are characterizedby their physical height and, when desired, the parameters to calculate buoyant ormomentum plume rise using Brigg's (Br69)or Rupp's (Ru48)formulations respectively. Alternatively, a fixed plume risemay be specified for each Pasquill-Gifford atmospheric stability class A through G.

The area sourcemodel 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-EPAmodels. 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 consideredtobe co-located at the origin of the assessment grid.

Radionuclide Releases

Releases for up to 36 radionuclidesmay be specified for AIRDOS-EPA. Each release is characterizedby the radionuclidename, 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 radionuclidemust 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 assessmentsas described in Chapters 

### 5 and 

### 6.

### The approach ORP 

### has used for calculating 

### aprecipitation scavenging coefficient isbased on Slinn's (S177)equation 32:


### h,, = Po 

### c J 

### E(a,%)

where 

### h., is the scavenging coefficient, c is a constant (Slinnuses 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 interpretsDana 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 (0.3mm) and selecting a long-term average value of 1,000mm/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 theprecipitation scavenging coefficient and then scaled according

- 

### to the annualprecipitation 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 ofmagnitude. The 

### EPA scaling procedure reflects thepremise that thevariation of rainfall from one location to another dependsmore 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. Thevertical 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,

### whereE, 

### andl,are respectively themorning and afternoon mixing depths and h is theirharmonic 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 inAIRDOS-EPA. The dry deposition rate is theproduct of thedeposition velocity and the near ground-level air concentration, while thewet deposition rate is theproduct 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 sourceheight can 

### -

. be 

### significant fordry deposition. For locationsclose to an elevated source, wet deposition can provide theprincipal 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 (dryplus wet) deposition rate. The soil concentration is calculated by dividing thisvalueby the effective agricultural soil surfacedensity (kg/m2). Both concentrationsare 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 theparent radionuclide. For example, the ingrowth factor forPb-210 as a parent of Po-210 would be calculatedby 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 factorsmust be calculated in advance of runningAIRDOS-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 samemodel as inNRC Regulatory Guide 1.109 (NRC77). Changes from that model include consideration of environmental removal from the root zone, and separate values for food and pasturecrops of the interception fraction, areal yield, and soil-to-plant transfer values. Concentration calculations formeat and milk use the same models as the Regulatory Guidemodel. There are numerous parameters in the terrestrial pathwaysmodel. 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 locationmust be provided. EPA uses the 1970 census enumeration district data to calculatepopulation 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 itemsproduced are consumed.

Food Utilization Factors

In addition to the consumptionrate for different food categories (leafyvegetables, 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 categoryproduced 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 dominatedby the radionuclide intake from home-grown food, and hence there is generally no significant differencebetween 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/gof H,O) is calculated for an absolutehumidity of 

### 8 mg/m3

(NRC77). Etnier (Et8O)has calculated average absolute humidities for over 200 U.S. locations. The 8mg/m3value would be within a factor of 2 formost of them. The specific activity of atmospheric C-14 (pCi/gof carbon) is calculated for a CO, concentration of 330 ppm by volume (Ki78). Concentrationsof these nuclides invegetation are cal..cul.atecion the assuntptioil thaL the water and carbon content invegetation are front the atmosphere and have the
same specific activity as inthe atmosphere. The Rn-222 concentration in airis replaced by its short-lived decay product concentration inworking level units using a fixed equilibrium fraction (typically0.5 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 foodproducts," 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 ProtectionAgency, 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 datawith 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 InformationService, Springfield,Va.,November 1969.
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Dana, M.T. and Wolf, M.A., "ExperimentalStudies in Precipitation Scavenging," in PacificNorthwest 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 PacificNorthwest 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., "TurbulentDiffusion-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 UrbanAir Pollution Throughout the Contiguous United States, PublicationNo. AP-101, U.S. Environmental ProtectionAgency, Office ofAir 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 InternalDose 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 14creleased into the atmosphere as carbon dioxide,"Health Phys., 34, 141-159, 1978.
Little, C.A. andMiller, 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 EstimatesAssociated 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 ComputerizedMethodology for Estimating Environmental Concentrations and Dose to Man fromAirborne Releases of Radionuclides, EPA 520/1-79-009 (reprintof ORNL-5532), U.S. Environmental Protection Agency, Office of Radiation Programs, Washington, D.C., December 1979.NCRP84 Nat:l.onalCouncil 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-atoryCommission, "Calculation of Annual Doses to Man from Routine Releases of Reactor Effluents for the Purpose of Evaluating Compliancewith 10 CFR Part 50Appendix 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 DevelopmentAdministration, Washington, D.C.,June 1977.

Sj84 Sjoreen,A.L., and Miller, C.W.,PREPAR-A User Friendly Preprocessor to CreateAIRDOS-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.?'., '"ExperimentalStudies in Precipitation Scavenging," in PacificNorthwest 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 receivedby 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 thephoton 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 spontaneouslydecay. 

### 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, themillicurie

(mCi),the microcurie (pCi),and thepicocurie (pCi),are also often used. The curie, which is defined as 3.7x10L0 disintegrationsper 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 thatA,(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 (theactivity 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 beproduced 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(28.6yr) ~?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 thebody, 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 themost 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 themean energy impartedA&,by ionizing radiation to a small finitemass of organ tissue with a mass, Am, and is expressed as



### - - D = 

### de/dm = 

### lirn 

### (A EJ

### A

### m)



## . Am4 

### Internal and external exposures from radiation sources are not usually instantaneousbut are distributed over extended periods of time. The resulting time rate of change of the absorbed dose to a smallvolume ofmass is referred to as the
absorbed dose rate, D:

D = 

### dD/dt = 

### lirn (AD/A1).
AL-40



### The customary unit of absorbed dose rate is any quotient of the rad (oritsmultiple 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 lengthby charged particles in an absorbingmedium. It represents the increment of themean energy lost, AE,to tissue by a charged particle of specified energy in traversing a distance, AX:

L, = 

### dE/dX = 

### lim 

## (AEE/X) (keVpm-l) (5-8)
Ax4



### For photons, L-represents the energy imparted by the secondary electrons (electronsthat are knocked out of their orbitals by primary radiation) resulting from secondary interactions between the photons and tissuematerial. 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 theproduct 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 othermodifying 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 

### (AH/At)



## . (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 20

### 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 introducedby 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 fordose equivalentsdistributed nonuniformly in thebody. 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 tableprocedure in conjunction with age- and organ-specific riskmodels adapted from the BEIR I11 

### report (NAS80).

These models (seeChapter 6) assume a linear dose-response relationship and a lifetime relative risk projection for cancers

### Table 5-2. 

### Weighting factors recommendedby 

### 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 risksper unit dose equivalent (mrem)will generallybe 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 1.3x lo5MeV of alpha energy. An activity concentration of 100pCi/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 levelmonths (WLM). One WLM correspondsto an exposure to a concentration of 1 WL for the commonly used reference period of 170hours.
Customary and SI Units
The relationshipbetween 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 DOSIMETRICMODELS

The EPA dosimetric models, to be discussed in the following sections,have been described in detail in previous publications

(Du80,Su81). Informationon 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 EPArsdosimetric models. Inmost cases, the EPA models are similar or identical to those recommendedby the ICRP

(ICRP79,ICRP80, ICRP81). However, differences inmodel 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 specialunits 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



## . radper second J kg-' s-I grayper 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 remper second J kg-' 

### s-' sievertper J kg-' 

### s-'

- Rate 

### (H) (rem 

### s-') second (Sv s-')

Linear Energy kiloelectron 1.602~10-'~ ~m-' Transfer 

### (L,) voltsper micrometer (keV

## m-')5.3.1 Internal DoseModels

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
aremade 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 compartmentmay 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 reportmay represent only portions of a structure usually considered to be an organ, while some compartmentsmay represent portions of the body usually not associated with organs. Examples of compartmentsused in this report are the stomach, thepulmonary 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 (thetracheobronchial, 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 intestinemay be absorbed through the wall into the bloodstream, from which it may be taken inparallel into any of several compartments within the skeleton, liver, kidney, thyroid, and other organs and tissues.

The list of organs or regions forwhich dose rates are calculated is found in Table 

### 5-4. Activity in the lungmay reach thebloodstream either directly or indirectly through the stomach or lymphatic system. The respiratory system and gastrointestinal tract models are discussed further in later sections. Figure 5-1Table 5-4. Target organs and tissues used for calculatingthe ICRP effectivedose 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 intestinewall Upper large intestinewall Lower large intestinewall 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 tobreast isassumed toequal dose tomuscle.
b) The ICRP estimate is

### a mass-weighted mean of the values forthe tracheo-bronchial region, thepulmonary region, and the respiratory lymph nodes.

C) The EPA averages thevalues forthe small, upper large, and lower large intestineusing weights of 0.2, 0.4, and 0.4 respectively for calculating the risk of bowel cancer. d) The pancreas is alsoused as a surrogateorgan for calculating the cancer risk for allother organs andtissues.i.ll.ustratesthe 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" indexedby 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 ithspecies 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 organmay 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 describedby 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,, canbe 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 absorbedby 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 estimatedby the following equation:

where: Dl(X+Y,;t) = 

### S,(X+Y ,) 

### A,,(t) ;

### andA,,(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 (numberper 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 toproduce 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 themodifying factor, N,(Y,) :



### Hi

### (X;t) =

## CCAik(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 crossirradiationsby 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 representedby 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 selectedby 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 thebody 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 causedby inhalation of these aerosols can depend

### on 

### a varie'tyof factors, such as breathing rates, particle sizes, and physical activity. Estimating the total dose to individuals
over a specific timeperiod requires specifying the distribution of particle depositions in the respiratory tract and the mathematical characteristics of the clearance parameters. The EPA currently uses assumptions establishedby 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 fractionsD,, 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 usesAMADs 

### of 1micron. 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 thepulmonary 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 determinedby averaging the 23 ICWadult 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 

### - ICWGI 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 themodel 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 characterizedby f,, which represents that fraction of the nuclide ingested which is absorbed intabody 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 formulatedby 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 orby contaminated ground surfaces.

RADRISK 

### computes doses forboth 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 ratesmay 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 ismore rapid at early times after exposure, doses are computed annually for the first several years and for progressively longer periods thereafter, dividing by the lengthCOMPARTMENT

N-P a (D3=0.30) b

Figure 5-3. 

### The ICRP Task Group lung Model for particulates.

The colmnslabeled

### D, W, and Y correspond, respectively, to rapid, intewediate, and slaw clearnnce of the inspiredmaterial (indays, weeks, or years). The symbols Tand 

### 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 activitymedian aerodpamic diameter 

### AHAD 1 pn and represent the fraction of the inspiredmaterial 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-activitymodel 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 (seeChapter 

### 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 individualswho ingest the material under different conditions. For example, f,values for uraniumwere found to range from 0.005 to 0.05 for industrialworkers, but a higher average value of 0.2

(0.12to 0.31) is indicatedby dietary data from persons not ocoupationally exposed (ICFS79). EPA 

### has used the 0.2 value for uranium ingestionby 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, themass increases during childhood and continues to increase until adulthood, at which time the net growth of the organ ceases; theremay be 

### a gradual decrease in mass (forsome 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 18.3g. For this illustration, the same values are used as employedby 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 1-131by the thyroid is similar to that of stable iodine in the diet and can be influencedby sex and dietary differences. There can be considerable variation among populations.

### Dunning and Schwarz (Du81)found amean 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 thepossible 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 thepaper 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 radionuclidewithin 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 formost 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 correspondingE 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 inmilk, 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 ParameterVariabililty 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 suchvariation gets averaged over a population and/or over a lifetime. Nevertheless, it shouldbe 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.

Parametervariation 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 fromwhom 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 theparameters 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 (seeChapter 6); 

### if doses are also higher in childhood, the enhanced effect on risk will be compounded.

5.3.1.6.9 

### PastApproaches 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 uncertaintyby 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 thesemodels were developed to yield point estimates. Despite these and other difficulties, propagation of parameter uncertainty through the dosimetric equation canprovide a measure of themodel uncertainty. Application of thesemethods 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 thepotential 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 calculationof the energy deposition in the target tissues is a minor contributor to the overall uncertainty.
Classification of Uncertainty in Radionuclide Dose
Establishingnumerical values of uncertainty formodel 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 thepopulation being evaluated. Most risk
assessorsuse 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 ispresented. Because of the problems cited above with respect to the development of models andmodel 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 someperspective 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 dosimetricmodels.
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 toxicmetals 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 isbased on judgment, a preferable term would be "credibility interval" (NIH85).]
Group I 

### - Essential Elements and TheirAnalogs

Essential elements are controlledby 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 isperhaps 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 tobe large. However, the gases do not distributeuniformly 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 toAr-41, Kr-85, Xe-133, and Rn-222.

Group 111 

### - Well-Studied ToxicMetals

A number of elements have been extensively studied in animals with limited information available forman. Examples here include toxic elements encountered in industrial activities, e.g., mercury, cadmium, lead, and uranium, for which studieswere carried out to help establish safeworking 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 toblood 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 forpolonium, 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 toman 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 theAgency'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 thephysics. 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 radioisotopesand aspects of dosimetry pose unique problems. While the effect of these problemsmay be to increase the uncertainty in dose estimates, the extent of such an increase has yet tobe evaluated.

Long-Lived Bone Seekers
Radioisotopes with effective half-lives that are short compared to the average life span are expected tobe 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 distributionmay be nonuniform with respect tobiological 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 thebone 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 liningmay 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 tophotons 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 tobe 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. Inboth 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 thephotons 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 (thedose rate for a unit

### concentration) in air is calculated for a source of photons with energy E,. At all points in an infiniteuniform 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 rateper unit air concentration (radm3/Cis);

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 (radkg/J)

= 

### 5.93Ei2 (grad/MeVCi s); and 

### P = 

### density of air (g/m3).
The above equation presumes that for each nuclide transformation, onephoton 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 fielddepends on the height of the receptor above the surface, and the dose rate factor calculat-ionis more complicated. The radiation flux per unit solid angle is strongly dependent on the angle of incidence. It increases from the value forphotons incident from immediately below the receptor to a maximum close to the horizon. Attenuation and buildup due to scatteringmust be considered to calculate the dose rate factor. Secondary scatteringprovides 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 tobe used quite extensively even for exposure conditions forwhich they are not strictly applicable


# ... because more realistic estimates are considerably more difficult and expensive [tomake]

## ."

### 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 dasAbsetzen Kleiner in der Luft Suspendierten Teilchen in der Menschlichen Lunge bei der Atmung, PflugersArch. 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.

ICRP59InternationalCommission 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.

ICRP75InternationalCommission on Radiological Protection, Report on the Task Group on Reference Man, ICRP PublicationNo. 23, Pergamon Press, Oxford, 1975.

ICRP77InternationalCommission 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 Radionuclidesby 

### Workers, LCRP PublicationNo. 30, Pergamon Press, Oxford, 1979.

ICRP80InternationalCommission 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.

ICRP84InternationalCommission on Radiological Protection, "A 

### Compilation of the Major Concepts and Quantities in Use by ICRP", ICRF PublicationNo. 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 theAnnual Meeting of the Health Physics Society, Seattle, Washington, July 20-25, 1980.

Ke80a Kerr., G.D., 

### "AReview 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., 

### "ANew Look at the Dosimetry of 14c

### Released to theAtmosphere 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 (BEIR111), Washington, D.C., 1980.

NCRP71National Council on Radiation Protectionand Measurements, Basic Radiation ProtectionCriteria,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 theNational Institutes of HealthAd Hoc Working Group to Develop Radioepidemiological Tables, NIH PublicationNo. 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., PostnatalEnlargement 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 toVariability 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 Equivalentper Microcurie-Day for Source and Target Organs of anAdult forVarious 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 ofAlpha-Emitting Radionuclides in Bone with ParticularEmphasis 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 geneticeffects, and other detrimental health effects caused by exposure tolow levels of ionizing radiation. Ionizing radiation refers toradiation that stripselectrons from atoms ina medium through which it passes. The highly reactive electrons and ions createdby thisprocess ina livingcell can produce, through a series of chemical reactions, permanent changes (mutations)inthe cell's geneticmaterial, the DNA. These may 

### result 

### incell death orinan abnormally functioning cell. A mutation ina germ cell (spermorovum) may be transmitted toan offspring and be expressed asa genetic defect inthat offspringor inan individual of 

### a subsequent generation; such a defect is commonly referredtoasa geneticeffect. There is also strong evidence that the induction of a mutationby ionizing radiation ina non-germ (somatic) cell can serve asa stepinthe development of a
cancer. Finally, mutational orother events, includingpossible cellkilling, produced by ionizing radiation inrapidly growing and differentiatingtissues of an embryo or fetus can give risetobirth defects; these arereferredtoasteratological effects. At acute doses above about 25 rads, radiation induces other deleterious effectsinman; however, for the low doses and dose rates of interest i n t h i sdocument, only those three kinds of effects referred toabove arethought tobe significant.

Most important from the standpoint of the total societal risk from exposures tolow-level ionizing radiation are the risks of
cancer and geneticmutations. Consistent with our current understanding of their origins 

### interms of 

### DNA damage, these are believedtobe stochastic effects; i.e., theprobability (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 geneticeffects, moreover, is there any convincing evidence for a "threshold," i.e., some dose levelbelow which the risk is zero. Hence, so farasisknown, any dose of ionizing radiation, no matter how small, might give risetoa cancer ortoa genetic effect infuture generations. Conversely, there is

no way tobe certainthat a given dose of radiation, no matter how large, has caused an observed cancer inan individual orwillcause one inthe future.
Beginning nearly with the discovery of x-rays in 1895but especially sinceWorld War 11, an enormous amount of research has been conducted intothebiological effectsof ionizing radiation. This research continues atthe level of themolecule, the cell, the tissue, thewhole laboratory animal, andman. There aretwo fundamental aspects tomost of thiswork:1. 

### Estimating the radiation dose to a target (cell, tissue, etc.)



## . This aspect (dosimetry), which may involve consideration ofphysiological,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 thepurpose of assessing the risk toman from exposures to ionizing radiation, themost important information comes from human epidemiological studies inwhich thenumber of health effects observed in an irradiated population is compared to that in an unirradiated controlpopulation. Thehuman epidemiological data regarding radiation-induced cancer are extensive. As a result, the risk can be estimated to within an order ofmagnitude with a high degree of confidence. Perhaps for only one other carcinogen

## - tobacco smoke

## - is it possible to estimate risksmore reliably.
Nevertheless, there are gaps in the human data on radiation risks. No clear-cut evidence of excess genetic effectshas 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 standpointof environmental exposures. Since the epidemiological data are incomplete inmany respects, risk
assessorsmust rely onmathematical models to estimate the risk from exposures to low-level ionizing radiation. The choice of models, of necessity, involves subjective judgments but shouldbe based on all relevant sources of data collectedby 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 phrasedby the President of theAcademy, "We believe that the report will be helpful to the EPA and other agencies as they reassess radiation protection standards. It provides the scientificbases upon which standards may be decided after nonscientific socialvalues have been taken into account."

In this discussion, thevarious assumptionsmade in calculating radiation risksbased on the 1980NAS report are outlined, and these risk estimates are compared with those prepared by other scientific groups, such as the 1972NAS BEIR Committee (NAS72),the United Nations Scientific Committee on the Effects ofAtomic 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 contaminationby radioactive materials also leads to the ingestion or inhalation of thematerial 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 6.2.9to 6.2.11.
Organ doses can also result fromhigh-LET radiation, such as that associated with alpha particles. The cancer risks when highLET radiation is distributedmore or less uniformly within a body organ is the third situation considered (Section6.3). Because densely ionizing alphaparticles 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 (Section6.4).
In Section 6.5, the causes of uncertainty in the cancer risk estimates and themagnitude of this uncertainty are reviewed so that both thepublic and EPA decision makers have a proper understanding of the degree of confidence toplace 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 themodels 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 JapaneseA-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).

Themost important source of epidemiological data on radiogenic
cancer is thepopulation of JapaneseA-bomb survivors. TheA-bomb survivors have been studied for more than 38 years, and most of them (theLife Span Study Sample) have been followed since 1950 in a carefullyplanned andmonitored epidemiological survey (Ka82,Wa83). They are the largest group that has been studied, and they provide themost detailed information on the response pattern for organsby

### age and sex over a wide range of doses of low-LET radiation. Unfortunately, the 1980 BEIR Committee's analysis of theA-bomb survivor data collectedup to 1974was prepared before bias in the dose estimates for the survivors (the tentative 1965dose 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 "BEIRV 

### Committee" of theNational Academy of Sciences ispreparing a report on radiation risks in light of DS86 and other new information. A detailed reevaluation of EPA's current risk estimates is indicatedwhen 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 ofuncertainty associatedwith these assumptions isprobably greater than the uncertainty of the estimated risk per unit dose among theA-bomb survivors or other
sources of risk estimates for radiogenic cancer inhumans.

6.2.1 Assumptions Needed toMake 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 themost part, only following doses of ionizingradiation that are relatively high when compared to those likely to occur as 

### a result of the combination ofbackground 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 radiosensitivityof a particular tissue. For example, themost probable radiogenic cancer for women isbreast cancer. As described below, the incidence of radiogenic breast cancer does not seem to diminish when the dose isprotracted
over a long period of time. For example, the number of excess cancers per unit dose among Japanesewomen, who received acute doses, is about the sameper 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 carcinogenicper unit dose for breast tissue as the acute exposures from A-bomb gamma radiation.Moreover, the femaleA-bomb survivors show an excess of breast cancer at doses below 20 radswhich is linearly proportional to that observed at several hundred rads (To84). (Evidenceof 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 ofwhole-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 ofmagnitude as is sometimes claimed. Based on data from irradiated tinea capitispatients, induction of thyroid
cancer also seems tobe linearwith doses down to 10 rads or lower

(NCRP85). However, formost other cancers, a statistically significant excesshas 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 DoseResponse Functions

The 1980NAS report (NAS80)examined only three dose response functions in detail: (1)linear, inwhich the number of effects (risk)is directly proportional to dose at all doses; (2)linearquadratic, inwhich risk isvery 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 compatiblewith 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 confoundedby an assumed neutron dose component. The chance that a quadratic response function underlies the excess cancer observed in theNagasaki 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 thepurposes of this rulemaking, a quadratic response cannot be used to estimate
cancer risks due to ionizing radiation.The 

### 1980NAS HEIR Committee considered only the Japanese mortality data in its analysis ofpossible dose response functions

(NAS80). Based on the T65 dose estimates, this Committee concluded that the excessmortality from solid cancers and leukemia among the A-bomb survivors is compatiblewith 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 scientistswho prepared that report, opinion was not unanimous, and the subsequent reassessment of theA-bomb dose seriouslyweakens the Committee's conclusion. The Committee's analysis of dose response functionswas based on the assumption thatmost of the observed
excess leukemia and solid cancers among survivors in Hiroshima resulted from neutrons (seeTablesV-13, A-7, EquationsV-10, V-11 inNAS80). 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 thatmost of the response attributed to neutrons was causedby 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 therewill generally be a higher response per unit dose, there will also be an unpredictable change in the shape of the dose response exhibitedby the data. Reanalysis of the Japanese experience after completion of the dose reassessment may thenprovide more definitive informationon the dose response of theA-bomb survivors; nevertheless, it is unlikely to produce a consensus on the dose response at environmental levels, i.e.,about 100mrad/yr. This isbecause 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 (seebelow), no data on human cancers that are incompatiblewith 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, shouldbe based on scientificallycredible risk models that are unlikely to understate the risk. The linear model fulfills this criterion. Given the current bias in the doses assigned toA-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 Committeepreferred the linear-quadraticdose responsemodel. In thismodel, 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 

### 2.5times 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 discussedbelow. 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 cellularDNA requirethe formation of interacting pairs of "sublesions." The interacting pairs canbe produced by a single traversingparticle, or track, orby two separate tracks, giving rise, respectively,to a linear and quadratic term in the dose response relationship. According to the theory, a sublesionmay 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 causedby separate tracksbecomes less important, and the magnitude of the D' term diminishes. Hence, the theory predicts that at sufficiently low doses or dose rates, the response shouldbe
a linear function of dose. Moreover, the constant of proportionality is the same in both cases; i.e.,

### a.

Results ofmany 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 forhigh-LET radiation (exceptfor 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 effectivenessof repair, and therefore the shape of the dose response curve, can in principle vary substantially with cell type1980 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 twoprojection 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 themore credible linear and linearquadratic response functions, the difference is less, a factor of about 2.2. For a given dose responsemodel, results obtained with the twoprojection models for solid cancers differby about a factor of 3.

Differences in the estimated cancer risk introducedby the choice of the risk projection model are also appreciable. As pointed out above, the 1980 NAS analysis indicatesthat relative lifetime risk estimates exceed absolute risk estimates by about a factor of 3 (seeTable 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 1980NAS 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 I11values were calculated for this table using riskmodel data fromNAS72 and NAS80. The BEIR I11values 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.

TheNAS 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 projectionmodels (if any) have been specifiedby the UNSCEAR (UNSCEAR77)or the ICRP

(ICRP77). UNSCEAR apparently presumes the same age distributions as in the epidemiological studies it cited, mainly theA-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 riskmodels with less than lifetime expressionTable 

### 6-1.Comparative estimate of cancer fatalities induced by
a single 10-rad, low-LET radiation exposure to the general population (averagevalue 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, TableV-25.increase in solid cancers that has occurred among irradiated populations (Xa82).

Toproject 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 expressionperiod. 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 100,000persons. This analysis considers not only death due to radiogenic cancer,but also theprobabilities of other competing causes of death which are, of course, much larger and vary considerablywith age (Bu81,Co78). Basically, it calculates for ages 0 to 110 the risk of death due to all causes by applying the 1970mortality data from theNational Center for Health Statistics (NCHS75)to a cohort of 100,000 persons. Additional details of the life table analysis areprovided 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 theNAS BEIR 111 Committee by EPA and used by the Committee to prepare its risk estimates. Therefore, thepopulation base and calculations shouldbe essentially the same in both the NAS and EPA analyses.

Both absolute and relative risk models have been considered to project the observed risks ofmost 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 calculatedby 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 (theBEIR 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 expressionperiod for leukemia and bone cancer is less than the follow-up period, the same risk values would be calculated for thesecancers using eitherprojection method. For a dose tothewhole body, thisprocedure yields about 400 fatalitiesper 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 alsopresented estimates of excess soft tissue cancer incidence risk coefficients for specific sites, as a function of age atexposure, initsTable V-14. By summing the site-specific risks, it then arrivedatan estimate forthe whole-body riskof cancer incidence (otherthan leukemia and bone cancer) as given inTable V30. Finally, by using the weighted incidence/mortality ratios given inTable V-15 of the same report (NAS80), the results inTable V-30 can be expressed interms of mortality toyield (forlifetime exposure) a risk estimate of about 242 and 776 cancer fatalitiesper 10' person-rad, depending on whether an absolute or a relative risk projectionmodel, respectively, isused toestimate lifetimerisk. These values 

## -areabout 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 allpresume a uniform dose toalltissues atrisk inthe body. In practice, such uniform whole-body exposures seldom occur, particularly for ingested orinhaledradioactivity. The next section describes how thisrisk estimate is apportioned forwholebody exposure when consideringthe risks followingthe exposure of specificorgans.
6.2.9 Organ Risks

For most sources of environmental contamination, inhalation and ingestionof radioactivity aremore common than external exposure. Inmany cases, depending on the chemical andphysical characteristicsof the radioactive material, inhalation and ingestionresult in

### a nonuniform distribution of radioactive materials within thebody so that some organ systems receive much higher doses than others. For example, since iodine isotopes concentrate preferentially inthe thyroidgland, the dose tothis organ can be orders of magnitude largerthan the average dose tothe body.

To determine theprobability that fatal cancer occurs ata particular site, EPA has performed lifetable analyses for each
cancer typeusing the information on cancer incidence andmortality inNAS80. NAS80 published incidence risk coefficients (NAS80 Table V-14) andmortality toincidence ratios (NAS80 Table V-15). The data inTables 6-3 and 6-4 are from these tables with the exception of themortality toincidence ratios forthyroid and lung cancer. Since not allforms of thyroid cancer can be induced by radiation and since, forthose that are, a more reasonable mortality to incidence ratiowould be 0.1 (NRC85), EPA has used that value inits calculations. Lung cancer incidence andmortality have both shownan increasing trend between 1970 and 1980. Since incidence leads mortality, an uncorrectedmortality 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 studiesby 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 theprocedure used inNAS80, absolute risks at an absorbed dose rate of 1mrad/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 theNAS80 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 theNCHS 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 riskTable 6-3. Site-specific incidence risk coefficients per rad-y)



### .

Age atExposure 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, TableV-14Table 6-4. Site-specificmortality to incidence risk ratios.

Site Male Female

Thyroid Breast Lung Esophagus Stomach Intestine Liver Pancreas Urinary Lymphoma Other

Source: NASBO, TableV-15, except thyroid and lung (seetext)



## .

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 inNAS80. These differences reflect a correction in the exposure interval data for each age group and the use of final rather thanpreliminary 1970mortality 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 calculatedby 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 obtainedby dividing the value for alphaparticles (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 fromNAS80 Table V-17. EPA hasfollowed the BEIR 111 Committeeb practice of using the absolute riskmodel projections for Leukemia and bone cancer with the relative riskprojection for allother cancers. Sincethe expressionperiod for leukemia andbone cancer is 27 years, there is
no differencebetween the number of cancers projected for a 30-year and a lifetime follow-up period.

Table 6-8 shows the average mortality risksper unit absorbed dose forthe combined leukemia/bone and constrained relative risk models. The risk, ingeneral, decreases with increasing age at exposure. For a constant, uniform absorbed dose ratetoallorgans and tissues, about 60 percent of the risk is conferred by the exposures inthe first 20 years of life.

The mortality toincidence ratios inTable 6-4 were used to convert themortality risk estimates inTable 6-8 toincidence risk estimates. For leukemia and bone cancer, the incidence risks are considered tobe equal as inNAS80. The resultant incidence risks
are shown inTable 6-9.

Table 6-10 summarizes the low LET cancermortality riskper unit dose for a uniform, lifetime dose toeach siteinTable 6-8. Since site-specific dose factors for estimatingrisk of esophageal
cancer and lymphoma arenot available, these risks areadded to those already inthe category "other." The proportional risk column shows the fraction of the total risk associatedwith each 

### sitewhen allorgans and tissues areuniformly irradiated.

6.2.10 Thyroid Cancer from Iodine-131 and Iodine-129

Iodine-131 has been reportedtobe only one-tenth as effective
asx-rays orgamma rays ininducing thyroid cancer (NAS72, NCRP77, NCRP85). BEIR 111 reported estimates of factors of 10-80 times reduction for 1-131 compared tox-rays and noted the estimates were derivedprimarily from animal experiments (NAS80). However, one study inrats reportedthat 1-131 was just aseffective asx-rays in inducing thyroid cancer, leadingan NRC review group toselect onethirdasthe minimum ratioof 1-131 tox-ray effects that is compatible with both oldand new data (NRC85).

It would be prudent touse this factoruntil further information from animal studies or some human data aredeveloped. 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 orbeta radiations from other radionuclides.

6.2.11 Cancer Risks for a Constant Intake Rate

The fatal cancer risks shown inthe tables of this chapter presume a lifetime exposure ata constant dose rate. Even fora

### - Table 6-5. BEIR 111 L-L model for excess fatal cancers other than leukemia andbone cancer.

Age 

### atExposure Group 0-9 10-19 20-34 35-49 50+ A11

Risk Coefficients per rad-y) forAbsolute 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) forRelativeRisk Model

Male Female General

Cohort Deaths 

### at

## lo-' rad/y forRelative Risk Model

Male Female General

Risk per Unit Dose per rad) forRelativeRisk Model

Male Female General

Source: NAS80, TableV-20Table 

### 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 OtherTable 6-7. BEIR 111 L-L model for excess incidenceof (andmortality 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 .0760 .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 29.4Table 6-7. BEIR I11 L-L model forexcess incidence of (andmortality from) leukemia and bone cancer (absoluteriskmodel) (Continued)



## .

Age atExposure Site 0-9 10-19 20-34 20-34 35-49 A11

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, TableV-17 (seetext)

### Table 6-8. 

### Site-specific mortalit risk per unit dose 

### (%.CE-6peg 

### rad) for combined leukemia- 

# gone and constrained relative risk 

### model.

Age at Exposure
Site 0-9 10-19 20-34 35-50 50+ 

### A11

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-6per rad) for combined leukemia-bone and constrained relative risk 

### model.

Age atExposure

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-specificmortality risk per unit dose from low-LET radiation, averages for both sexes and 

### all ages.

Cancer EPA89"." EPA84".bUNSCEAR77 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 estimatesuse different age distributionsandperiods of expression.

EPA Radionuclides Background InformationDocument; EPA

520/1-84-022-1 (EPA84).

Normalized for risk of fatal cancer (excludes genetic effects)



## . Five additional target organs that have thehighest doses
are assigned 0.08 each for 

### a total of 0.4. Estimated proportions for this document.



## -dosimetricmode: with age invariant parameters, dose rates vary with time for a constant intake rate. This variation reflects the tirnedependent activity levels associatedwith 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 commitmentmodelto calculate dose rates following a fixedperiod (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 formany long-lived radionuclides. Therefore, EPA estimates risks for constant radionuclide intakesby first determiningdose 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 keVper pm in tissue, 5-MeV alphaparticles result in energy deposition ofmore than 100 keVper 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 relativebiological effectiveness (RBE)of high-LET alpha radiations is often 10 ormore 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 forhigh- and low-LET irradiations, respectively,then the RBE will decrease with increasingdose.

6.3.1 Quality Factors and RBE forAlpha 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 assessmentby 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 appropriatenessof this numerical factor for estimating fatalradiogenic 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 shouldbe noted that prior to ICRP Report 26 (ICRP79),the quality factor assigned to alpha particle irradiationwas 10. That is, the biological effect from a given dose of alphaparticles 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 forhigh-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 radiationdoes 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 calculatingthe risks from alpha particles (see Section 6.3.3).

. In 1980, the ICRP published the task group report "Biological Effects of InhaledRadionuclides," which compared the results of animal experiments on radiocarcinogenesisfollowing the inhalation of alpha-particle and beta-particle emitters (ICRP80). The task group concluded that: "...theexperimental animal data tend to support the decisionby the ICRP to change the recommendedquality 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 responseper unit dose of alpha radiation ismaximal 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: "ForhighLET radiation, such as from internally deposited alpha-emitting radionuclides,the linear hypothesis is less likely to lead to overestimatesof the risk andmay, 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 (butnot 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 consistencyof these data with a quadratic (orthreshold) response is not remarkable and, perhaps, not relevant to evaluating risks at lowdoses. In contrast to the dialpainter data, the incidence of bone cancer following shortlived radium-224 irradiation, observed in spondyliticsby 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 ofbone cancer. Closely related tothe 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, fromboth human and animal studies, that repeated exposures to radium-224 alphaparticles are 5 

### timesmore effective in inducing bone sarcomas than a single exposure that delivers the same dose (Ma83,NAS80). The 1980NAS BEIR Committee took this into account in its estimates of bone cancer fatalities,which EPA isusing.

6.3.3 Assump~ions Made by EPA for Evaluating 

### ~he Dose from Alpha-Parcicle Emit~ers

EPA has evaluated the risk to specificbody 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 factormay 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 (relativeto a low dose rate, low-LET risk model which includes a DREF of 2.5) 

### overpredicts the observed incidence of leukemia inpersons receiving thorotrast (thorium oxides) (Mo79)and in the U.S. radium dial painters (Sp83). Nevertheless, in view of thepaucity 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 forhigh-LET radiations arebased on a linear dose response function. For bone cancer and leukemia, EPA
uses the absolute risk projection model described in theprevious section. For other cancers, theAgency uses relative risk projections.

TheAgency's estimates of the risk of fatal cancer due to a uniform organ dose invarious 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 linearmodel for a uniformly distributedwhole-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 forbone cancer. As outlined above, the risk estimate for this cancer in the BEIR I11 report is based directly on data forhigh-LET (alpha) radiation. Some readersmay note that the risk estimate in Table 6-11, 19 bone cancer fatalities per lo6person-rad, is less than the 27 fatalities listed in Table A-27 ofNAS80 for alphaparticles. This is because the analysis in Appendix A of NAS80 (butnot ChapterV of that report) assumes that in addition to a 2-yearminimum inductionperiod, 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 smallerwhen 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 situationwhere 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 radioactivenoble 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 aerosolparticles in air. When inhaled, the radonprogeny are deposited on the surfaces of the larger bronchi of the lung. Since two of these radionuclidesdecay by alpha-particle emission, the bronchial epithelium is irradiatedby 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 generalpublic. Recently the National Academy of Sciences, BEIR IV Committee, and the InternationalCommission on Radiological Protection reviewed the question of radon risks and reported their conclusions (NAS88,ICRP87).
Although considerable progress has beenmade inmodeling the deposition of radon daughters in the lung, it is not yet possible to



### - characterize adequately the bronchial dose deliveredby alpha radiation from inhaled Rn-222 progeny (NAS88). This is inpart due to the uncertainty concerning the kinds of cells inwhich bronchialTable 6-1.1. Site-specific mortality riskper unit dose from internally deposited alpha emitters averaged forboth

sexes and aliages.

Site Mortality risk per 

## lo6personProportional 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.atedand 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. Itmust 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, therewould 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 UndergroundMiners

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 105MeV 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 100pCi/L of Rn-222. The WL unit was developed because the concentrationof specific radonprogeny 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 radonprogeny 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 formembers of the generalpopulation 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 adjustmentswere 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, radiationdose from radon daughters. This approach to quantifying exposure, correcting for differences in these factors,was recommendedby Evans (Ev69)and is consistent with the original derivation of the working level (Ho57).



## - The BEIR IV Cornittee, however, concluded that the tracheobronchial "doseper WLM in homes, as compared to that inmines, differsby less than a factor of 

### 2," and advised that the dose and risk per WLM exposure in residences and inmines should beconsidered 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 "exposureequivalentr' 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 0.75is still employed. Discussion of the support for this estimate can be found in EPA86.

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 itsmodel to reflect changes, as contained in consecutive reports (EPA79,EPA82, EPA83a, EPA83b, EPA84, EPA85,and EPA86). The Agency initiallyprojected 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 inductionperiod 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 3percent 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 1percent 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 survivorshave 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 studiedpopulations. The A-bomb survivor data indicatewomen are as sensitive asmen 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. TheAECB estimate for lifetime exposure to Canadian males is 830 fatalitiesper million 

### - person-WLM (Th82). In Table 6-12, this estimate has been adjusted for the U.S. 1970male and femalepopulation.

.
- 

### TheNational Institute for Occupational Safety and Health reviewed published data onminer studiesused as a basis for estimated risk coefficients and pointed out some of the strengths

- 

### and limitationsof selected studies (NIOSH87).

. 

### The occupational exposure groups that constitute the epidemiological database for the risk estimates are as follows:



### I. U.S.UraniumMiners (NIOSH87)
(a) Strengths: A 

### large, clearly defined, well-traced cohort with some smoking histories and exposure records on the
samepersons. Standard sampling techniques were used to make measurements.
(b) 

### Limitations: There were fewmeasurements in smallmines, work histories were self-reported, exposures were high, andpotential error due to excursions in exposure levels is high.
(c) Follow-up: 19 years in 1977.



### 2. Czechoslovakian UraniumMiners (NIOSH8
(a) Strengths: Extensive exposure data with a large number of low level exposures and limited exposure to other undergroundmining. Many possible confounding factors have been investigated and eliminated.(b) Limitations: Exposure estimates prior to 

### 1960 based on radon gasmeasurements. Person years at risk not determined in standardmanner. 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. Priormining experiencewas carefully traced. Exposures prior to 1967may be disputed.
(b) Limitations: Median age of the cohort was 39 years in 1977. Thoron and gamma exposuresmay have been high but not accounted for. Smokinghistory 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 confounderswas 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 forprior 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: 14years in 1980.

### Table 6-12. 

### Risk estimate for exposures to radonprogeny

Organization Model Fatalitiesper Exposure period Expression lo6person-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 40years 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 inparentheses express the risk inmore customary units, in which a continuous annual exposure to 1WL 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 Projection6.4.4 Recent Radon Risk Estimates

6.4.4.1BEIR 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 (theU.S.,Ontario, and Eldorado uranium miners and theMalmberget ironminers) 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 generalizationof standard SMRmethods. 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 cancermortality rate

y(a) = 1.2, if a is less than 55years 1.0, if a is between 55 and 

### 64years 0.4,if a is greater than 64 years

W, = WLM incurredbetween 

### 5 and 15 years prior to age a

W, = WLM incurredmore than 15 years prior to age a
The Committee model is, therefore, an age-specific, relative-risk projection model with a 5-year latentperiod 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 twomajor
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 fourminer 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 errormay differ among the studies; (2)errors ofmisclassificationof cause of death; (3)errors in smoking status of individualminers, 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 (minerdata all for males); (2)effect of age (miner data contain no information on exposures before about age 20); (3)effect of smoking (minerdata contain poor information on smoking status); (4)temporal expression of risk (notenough miners have died to establish accurately the pattern of lifetime risk from radon exposure), and 

### (5)extrapolation frommining to indoor envirenments (what are significant differences in the air inmines compared to air indoors?). After reviewing the various sources of uncertainty, the BEIR IV Committee concluded

[p42],"



### ... The imprecisionthat 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 themodeling is strongly influencedby the U.S. uraniumminer 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: cumulativeWLM (CWLM)are calculated from measured radon levels for only 10.3percent of the miners, varying amounts of estimation are required for about 36.1percent of theminers, and guesswork is used for about 

### 53.6percent 

### of theminers (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 thepotential 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 IVmodel is run with the 1980 lifetable and vital statistics at an exposure level of 0.001WLM per year, the reference risk canbe calculated.

BEIR IV Relative Risk Model 

### - Lifetime Exposure and LifetimeRisk


### 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 coefficientsby 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/106PWLMY Ontario 1958-1981 0.5%-1.3% 3-7 cases/106

### PWLMY Average 1% 10 cases/106PWLMY

### The relative risk model then developed for a constant exposure rate is:

= 

### themortality 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 firstwas to correct for cocarcinogenic influences inmines. To account for unidentified, unproven carcinogens that might be present inmine environmentsbut not elsewhere, only 80percent of the risk was attributed to radon. The second adjustmentwas for dosimetric corrections. The dose to bronchial epithelium used by the Task Group for persons indoorswas estimated to be only 80percent as large as that for persons in 

## - mines; therefore, the risk to thepublic from radon was considered to be 80percent of the risk ofminers.Adjusting the average relative risk coefficient of 1percent per WLM by these two factors gives 

### a risk coefficient of 0.64percent per WLM:

The third adjustment made by the Task Group is related to age. Since reports of JapaneseA-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 20by 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 0.64percent 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 0.001WLM 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 theAECB to establish an upper bound of the range. The lower bound of the rangewas establishedby averaging the UNSCEAR and ICRP estimates. TheAgency chose not to include theNCRP estimate in its determination of the lower bound because this estimatewas believed to be outside the lower bound. With this procedure, the EPA arrived at relative risk coefficients of 1.2percent to 2.8percent per WLM exposure equivalent (300to 700 fatalities per million person-WLM exposure equivalent) as estimates of thepossible 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 ofmuch of current scientific opinion.

The lower limit of the range of 1985 EPA relative risk coefficients, 1.2percent per effective WLM, was similar to that derived by theAd 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 1.1percent per WLM (St85). On theother hand, three studies 

### - two on miners (Ra84, Ho86) and one on residential exposure (Ed83, 84) 

### - indicated a relative risk coefficient greaterthan 3 percent per WLM, perhaps aslarge as 3.6 percent.

The EPA therefore increasedtheupper limitof itsestimated range of relative risk coefficients. To estimatethe risk due to Rn-222 progeny, theEPA used the range of relativerisk coefficients of 1to4 percent per WLM. (SeeEPA86 fora more detailed discussion.) Based on 1980 vital statistics, thisyielded, for members of the general public, a range of lifetimerisks from 380 to 1,520 fatal cases per

## lo6WLM (expressedinexposure equivalents). In standard exposure units, uncorrected forbreathing rate and age, this corresponds to230 to920 casesper

## lo6WLM. Coincidentally, the geometricmean estimate obtained inthisway with 1980vital statistics, 4.6xlO-'/~LMin standardunits of exposure, is numerically the same as that obtained using a 3percent relative risk coefficient and 1970 vital statistics (seeTable 6-9).

. However, 

### inlight of thetwo recently published consensusbased reports, BEIR IVand ICRP 50, and a recent report on the Czech miner groups (Se88), theAgency has reviewed itsbasis forradon

. risk estimation. Comparable relativerisk coefficients forminers (age-constant relative risk) yield a coefficient of around 1percent

. in ICRP 50, 1.34 percent inBEIR IV, and 1.5percent inthe Czechs. This suggests that the range, 1percent to4 percent, usedby EPA may be too wide. Nevertheless, note that only 5 of the 20 or so studies forwhich there are some data are included 

### inthese estimates.

The BEIR IVCommittee noted andmodeled a drop inrelative risk with increasing time of exposure and a decreasingrelative risk with increasing age after exposure (NAS88). The Czech miners show a similar response pattern (Se88). Though the Committee didnote a dose rateeffect inthe U.S. uranium miner cohort, i.e., a decrease inriskper unit exposure athigh dose rates, it was not included in themodel (NAS88). The possibility of a similar dose-rate effect
was found recently ina study on Port Radium uranium miners (Ho87).

The ICRP 50 Task Group worked from a different database and developed a simplermodel with fewer age- and time-dependent parameters. The Task Group provided a 3 times higher risk for exposurebetween birth and 20 years of age than after 20 years of age (ICRP87). The finding inthe recent Czech report that risk priortoage 30 is2 to2.5 times greater than after age 30 lends
some support tothe ICRP conclusions (Se88).

Both BEIR IVand ICRP 50 models treat radon and smoking risks 

## - asmultiplicative. This conclusion isbased primarily on data from the U.S. uranium miner cohort. Although apparentlybased on weaker evidence, the report on Malmberget miners and the recent report onCzech 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) interactionshave been reported inman, and animal studies canbe found to justify any position (Ch81,Ch85, Cr78). In prior calculations,EPA has always treated the interactionbetween radon daughters and cigarette smoke as multiplicative. EPA will continue to treat the radon daughter-smoke interactionasmultiplicative at this time.
Important unresolved issuespertaining to the risks from inhaled radon progeny remain. At the advice of the Radiation Advisory Committee of EPA's ScienceAdvisory Board, EPA will continue to use relative risk models but shall include both BEIR IV and ICRP 50model calculations to illustrate the difference in results from the two models. The ICRP 50model will be slightly modified. The risk reduction factor of 0.8 to compensate for differences in dosimetry will be removed to place the ICRP 50model and 

### BEIR IVmodel on a comparativebasis. Calculations in the ICRP 50model will be made using risk coefficientsof 2.4percent 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 ofbackground radon exposure (seesection below).

Table 6-13. Lifetime risk from radon daughter exposure of lung
cancer death (per106

### 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 theA-bomb survivordata 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 stablevalues 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 inductionby 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 theparameters 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 estimatesbecame 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 theA-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 themix of gamma ray and neutron radiationswas quite different in the two citieswhere 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 significantnumbers 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.-JapanJoint 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 shieldedby terrain or were in factories,have so farbeen excluded from

### * * * DRAFT 

### * * *
the analysisbecause 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 differencesbetween T65 and DS86 are: (1)the neutron dose in DS86 is decreased to 10 percent of its formervalue 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 Hiroshimabut decreases in Nagasaki relative to T65; (3)transmission of gamma rays through wooden structures isdecreasedby 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 (linearor linear quadratic) will be increasedby 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 themagnitude of risk coefficients.

6.5.3 Sampling Variation

Besides the systematicbias 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 limitedby statistical fluctuations due to sample size. The uncertainty bounds 

### (21 SD) attached to the gamma-ray risk coefficient in the BEIR 111 linear model are about +25 percent, for either leukemia (TableV-8) or for all other cancers 

# able V-11). For the latter groups of cancers,however, the neutron RBE was constrainedto the value obtained from analysis of the leukemiadata. If this constraint is removed, the uncertainty in the estimate increases to +I50 percent (TableV-9). This increase reflects the large uncertainty associatedwith the neutron contribution in the analysis and the strong correlationbetween neutron and gamma-ray doses.Following the dosimetry reassessment,neutron doses will decrease markedly but will remain correlated with gamma-ray doses.
Finally, it shouldbe 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 (150to 350 rads) and dose rates (>5 rads min-') may overestimate the effects of either low doses (0-20rads or less) or of any dose delivered at dose rates of 5 rad y-lor lessby 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, fitsmuch 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 2.5less than the coefficient derived from the linearmodel. Thus, while the animal data indicate a reduced effectivenessof radiation at low doses and.low dose rates, the available human data would suggest that the linear dose responsemodel 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 responsemodel, a "transportation model" is needed to apply the risks from an observed irradiatedgroup to another population having different demographic characteristics. A typical example is the application of the Japanese data forA-bomb survivors to Western people. Seymour Jablon (Directorof theMedical Follow-upAgency of the National Research Council, NAS) has called this the "transportationproblem," a helpful designation because it is often confused with the risk projection problem described below. However, there ismore than a geographic aspect to the "transportationproblem,"Risk estimates for one sex must sometimesbe based on data for the other. In transporting risk estimates from one group to another, onemay have to consider habits influencing health statcs, such as differencesbetween 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 observationperiodwas transferrable but that relative risk was not. Therefore, the Committee calculatedwhat 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 theA-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 arepurely 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 Japanesebaseline rates differ drastically with respect to mortality from specific cancers, this approach implies some large differences in thepredicted number of specific cancers resulting from a given dose of radiation in the two countries. Themost important differences relate to cancers of the breast, lung, and stomach. Baseline rates of breast and lung cancers are higher in the United Statesby 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 shouldbe 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 20percent; on the other hand, applying the relative risk model to stomach cancer alone would lower the whole-body risk by about 8percent. 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 ofmore than +20 percent in the total risk estimate. Thus, in the case of uniform whole-body doses, the amount of uncertainty introducedby 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, suchmodels 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 lifetimebut 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, theBEIR 

### 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 theA-bomb survivor population strongly suggests that the relative risk projection model better describes the variation in risk of solid tumors over time

(NIH85). However, theremay 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-upbut 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 asmuch 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 coefficientsused here are those presented in the BEIR 111 report. 

## - As yet, there is little information on the ultimate effects of exposure during childhood. As theA-bomb survivor population ages, more information will become available on the cancer mortality ofpersons irradiated when they were young. Table 6-2 indicates that themore 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 theprojected effects on this group was a major contributor to the increase in risk foundwith 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 10percent) 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

Themost 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)theprojection 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 lo6person-rad, is taken from the NAS BEIR 111 Committee report (NAS1980),incorporating themost conservative model assumptionsutilized by the Committee 

### - i.e.,

### a linear dose response and age-specific relative risks projected over a lifetime for solid tumors (L-RRmodel). For reasons discussed above, it would now appear that estimates of average lifetime risk based on the L-RR model assumptionsmust be revised upwards 

### - to roughly 1,200 fatal cancers/106 person-rad. Although further analysis of the A-bomb survivordata may increase this estimate, the conservatism inherent in themodel's assumption supports the view that the 1,200/1O6value is an upper bound, pending release of NAS BEIR V report now in preparation.

Animal data would suggest that the linear dose responsemay 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 estimatemay 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/106personrad, 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 shouldbe emphasized that this estimate cannot be regarded as conservative in the sense ofproviding any significantmargin of safety with respect to public health protection. The decision by EPA to employ the central estimate of 400 fatalities/106person-rad and a range of 120-1,200 fatalities/106person-rad was reviewed and approvedby a special panel set up by the Agency's outside RadiationAdvisory Committee and by the Committee itself, as an interim measure for this proposed rulemaking.

The uncertainty in risks for specific cancer sitesmay be substantially larger than the uncertainty in the whole-body risk. One
reason is that the epidemiologicaldata pertaining to some sitesmay 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 differmarkedly between populations.

The uncertainties in risk associated with internally deposited alpha emitters are often greater than for low-LET radiation. Human epidemiologicaldata on the risks from alpha emitters are largely confined to: (1) lung cancer inducedby radon decay products (see below); (2)bone cancer inducedby 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 alsoproblematic.

Formany alpha-emitting radionuclides, themost 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 regardingthemetabolism of the radionuclide, uniformity of radionuclide deposition in an organ, and the short range of alphaparticles in tissue (see Chapter 5).



### In the case of alpha irradiation of the lungby 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/106WLM and an uncertainty range of 160-720 fatal lung cancers/106WLM.

6.6 

### OTHER RADIATION-INDUCED HEALTH EFFECTS

The earliest report of radiation-inducedhealth effectswas in 1896 (Mo67),and it dealt with 

## - acute effects in skin generally causedby very largex-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, sopossible acute effects will not need to be considered in assessing the risk to the general population from routine radionuclide emissions.

Radiation-induced carcinogenesiswas the first delayed health effect described: the first casewas reported in 1902 (Vo02),and 94 cases of skin cancer and 5 of leukemia were reported

### by 1911 (Up75). Radiationinduced genetic changeswere noted soon afterward. In 1927, H.J.Muller describedx-ray-induced mutations in animals (in the insect,Drosophila), and in 1928, L.J.Stadler reported

### a similar finding inplants (Ki62). At about the same time, radiation effects on the developing human embryo
were observed. Case reports in 1929 showed a high rate ofmicrocephaly

(smallhead size) and central nervous system disturbance and one case of skeletaldefects in children irradiated 

### -- in utero (UNSCEAR69). These effects, at unrecordedbut 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).

Forpurposes 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 theirprogeny and in future generations.

Of the possible consequences of radiation exposure, the genetic risk ismore subtle than the somatic risk, since it affects not the persons exposed, but relates only to subsequentprogeny. 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 (nearly1,000years) are needed for near comp1.ete expression of genetic effects. Genetic risk is incurredby fertile people when radiationdamages the nucleus of the cellswhich 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 damagemay 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 causedby mutations in somatic cells because, whereas somatic risk is expressed in the person exposed, genetic risk is expressed only in progeny and, in general, overmany subsequent generations. Moreover, the types of damage incurred often differ in kind from cancer and cancerdeath. Historically, research on genetic effects and development of risk estimates have proceeded independently of the research on carcinogenesis. Neither the dose responsemodels 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 "disordersand 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 recessivemutations in autosomal chromosomes, mutations in sex-linked (x-linked)chromosomes, chromosome aberrations

(physicalrearrangement or removal of part of the genetic message on the chromosomeor abnormal numbers of chromosomes), and irregularly inherited disorders (geneticconditionswith complex causes, constitutional and degenerative diseases, etc.).

Estimates of the genetic risk per generation are conventionally based on a 30-yr reproductivegeneration. That is, the median parental age for production of children is defined as age 30 (one-halfthe 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 ofgenetic effects includesboth first generation estimates and total geneticburden 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 extrapolatesto what is expected forhumans. 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 spontaneousmutation rate per gene site is dividedby the average radiation-inducedmutation rateper gene observed inmouse studies, to obtain the relative radiation mutation risk in humans. The inverse of this relative radiationmutation risk is the expected "doubling dose" for radiation-inducedmutations inman. The doubling dose is the exposure in rads which will double the current genetic malformation level inman and usually is used to estimate equilibrium effects or all future generation effects.
A doubling dose estimate assumes that the total population ofboth sexes is equally irradiated, as occurs frombackground 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 toman. Some of the committee reports assessing the situation have attempted to indicate the range of uncertainty; others have simply used a central estimate (seeTable 6-18). The same uncertainties exist for the latter (central estimates) as for the former.

Most of the difference is causedby the newer information used in each report. Note that all of these estimates are based on theextrapolation 

### of animal data to humans. Groups differ in their interpretation of how genetic experiments in animalsmight be expressed in humans. While there are no comparablehuman 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 shouldbe 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, andmost 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, theRBE 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 aremade 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 incorporationof 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 inmice, also ranged from about 4 to 50

(UNSCEAR82,Gr83a, Ga82). However, an RBE of 4 forplutonium-239 compared to chronic gamma radiationwas reported for specific locus mutations observed in neonatemice (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

(chronicx-ray or gamma radiation exposure).

Type of genetic Current incidence Effectsper lo6liveborn 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 estimatesby 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 inmice

(NAS80,pp. 109-110)



## .
Source: NAS80.

### Table 6-18. Summary of genetic zisk estimates per

## lo6liveborn of low-dose rate, low-LET radiation in a 30-yr generation.

Source Serious hereditary effects

First generation Equilibrium (allgenerations)

BEAR, 1956 (NAS72) 

## ---

BEIR I, 1972 (NAS72) 49" (12-200)
UNSCEAR, 1972 (UNSCEAR72) 9" (6-15)
UNSCEAR, 1977 (UNSCEAR77) 63


### ICRP, 1980 (Of80) 8 9

BEIR 111, 1980 (NAS80) 19" (5-75)

UNSCEAR, 1982 (UNSCEAR82) 22

Numbers inparentheses are the range of estimates.

" 

### Geometric mean of the lower and upper bounds of the estimates The geometricmean 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 (allgenerations)

Range 28

## - 298

Geometric Mean 91

- - 

### 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, itwould 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, inplants, in insects, and inmammals (UNSCEAR'72,77,82),and inperipheral blood lymphocytesof persons exposed to radiation (UNSCEAR82,Ev79, Po78). However, they cannot be used forpredicting genetic risk inprogeny of exposed persons. Some believe such changes tobe a direct expression of damage analogous to that induced by radiation in germ cells. At least, aberrations inperipheral lymphocytes show that radiation-induced chromosome damage can occur invivo inhumans. Since human data are so sparse, they canbe 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.

Databelow (Table6-20), collected by Van Buul (Va80),on induction of reciprocal translocations in spermatogonia invarious 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 therewere no human data on this genetic injury in the majority of cases, assuming that animal results and human results would be similarwould 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

## 70.02 to2.90

## + 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 inthe doubling-dose method of estimation is that there isa proportionality between radiation-induced and spontaneous mutation rates. Some of theuncertainty

### was removed in the 1982 UNSCEAR report with the observation that intwo-test systems (fruit fliesandbacteria), there is a proportionality between spontaneous and induced mutation rates ata number of individual gene sites. There is still some question astowhether
ornot the sitesthat have been examined arerepresentativeof ailsitesand a11 gene loci, with developing evidence that themouse 7-locus system ismore sensitive toradiation than othermembers of themouse genome (Ne88). The doubling-dose estimate, however, seems better supported than the directestimate.

There is some uncertainty astowhich 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 tothemost recent, has listedan increased number of conditions and diseases that have a genetic component and hence may be increased by exposure toionizingradiations.

6.6.4.1 Observations on Human Populations

A study of thebirth cohort consistingof children of the Japanese A-bomb survivors was initiated inmid-1946. In a detailed monograph, Nee1 and Schull (Ne56) outlinedthebackground of this first study and made a detailed analysis of the findings toJanuary 1954 when the study terminated. The study was designed to determine: (1)ifduringthe firstyear of life, any differences couldbe observed inchildren born toexposedparents when compared tochildren born to suitable control parents, and (2) if differences existed, how they shouldbe interpreted (Ne56).'Thisstudy 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 70,082persons have reported data on si.xendpoints. 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 (sexchromosome aneuploidy) and frequency ofbiochemicalvariants (avariant enzyme or protein electrophoresis pattern) have been measured on large subsets of this cohort.

There were smallbut 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 theparental 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 estlmatewas 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 overlapsmuch of the range based on extrapolation from
mouse data.

The most recent report evaluated the following possible genetic effects: (1).untowardpregnancy 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 findingsof 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 compatiblewith 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 JapaneseAbomb survivors are published.

6.6.4.2 Ranges of Estimates Providedby Various Models
Following recommendationsof 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 obtainedwith acute low-LET exposures are equally well describedby a linear-quadratic function. Moreover, in some of these cases, it has been found that a reduction in dose rate (or fractionationof dose)produced a reductjon in the quadratic termseen at high doses with little or no effect on the linear component. Such observations can be quali-tativelyexplained, as previously discussed in reference to somatic effects (Section6.2.21,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 estimatesmade by different committees based on the linear non-threshold model vary, the agreement is reasonably good. Some of the committeesmade estimates in terms of a range. These ranges are expressed as a singlevalue by taking the geometricmean 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 "doublingdose" 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 affectedby ionizing radiation are not known nor have all loci associated with "serious" genetic conditions been identified. Because the risk estimatedby the directmethod 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 estimatesprovide a better estimate of uncertainty than the UNSCEAR 1982 and ICRP estimatesbecause the BEIR I11 Committee assigned a range of uncertainty formultifactorial diseases (15percent to 

### < 50percent) that reflects the uncertainty in the numbersbetter than the other estimates 

### (5percent 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)forpurposes of calculating genetic risk. The factor of 

### 3 increase in risk forhigh-dose rate, low-LET radiation, noted earlier, is also used. The weighted RBE forhigh-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 forhigh- 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 inTable 6-21 are limited, like allother human genetic risk estimates, by the lack of confirming evidence of genetic effects inhumans. These estimates depend on a presumed resemblance of radiation effects inanimals tothose inhumans. The largest human source of data, the Japanese A-bomb survivors, appears atbest toprovide an estimate of the doubling dose for calculating the genetic risk inman which isnot statistical.1y significant (Ne88)



### .
In developing the average mutation rate forthe two sexes used inthe calculation of the relativemutation risk, the BEIR I11 Committee postulatedthat the induced mutation rate in females was about 40 percent of that inmales (NAS80). Studies by Dobson, et al., show that the basis forthe assumption was invalid andthat human oocytes shouldhave a risk equivalent to that of human spermatogonia. This would increasethe risk estimate obtained from doubling-dose methods by a factor of 1.43


### - Table 6-21. Estimated frequency of genetic disorders ina birth 

## - cohort due toexposure of theparents to 1radper

. generation.

Serious heritablg disorders

- (Casesper 10 liveborn)

. Radiation First generation Allgenerations

Low Dose Rate, Low-LET 20
High Dose Rate, Low-LET 60

(Do83, Do84, D088)



## . Recently Dobson etal. (Do88) have shown that
mouse oocytes arevery sensitivetoradiation, doses of 4 to 12 rads killing 50 percent of the immature mouse oocytes. Immature oocytes inwomen are not so easily killed. Dobson etal. (Do88) have also shown the existence of a special, hypersensitive, non-DNA lethality target (apparentlytheplasma membrane) inimmature mouse oocytes. Irradiationwith low energy neutrons, whose recoilprotons have track lengths lessthan a celldiameter, induces genetic effects in immature mouse oocytes and yields effec~s similar tothose observed in other cells (Do88). Immature human oocytes do not have the same hypersensitive target asmouse oocytes and so shouldbe as 

## - susceptible as spermatogonia togeneticeffects of radiation.
Unfortunately, BEIR 111 and, sinceit isbased on BGIR 111, BEIR IVhave embedded sex-sensit.iv3.t~differences intheir riskestimates. 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 

### 44percent 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 ismade 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 contributionsto various disordersprobably 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 includemild mutations which are but slightly detrimental in their heterozygous state. However, theymay be sustained by advances inmedical science, thuspersisting and accumulating for generations. To what extent this occurswill 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 inducedby irradiationand 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 (N1000years), the background mutation ratewill change and come into equilibrium with the new level of radiationbackground. There will be an accumulation of new radiation-induced mutations until the backgroundmutation rate has reached equilibrium with this continued insult.
While predicting 1,000years 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 radiationwill induce 260 cases of serious heritable disorders per lo6live births in all generations. This is for a backgroundmutation rate leading to 29,120 cases of seriousheritable disorders per 106Live 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 (seeNAS80, 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 inbackground level of genetic effects after 30 generations or more.

Increase in Increase in serious heritable background disordersper 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 4higher 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 dealsmostly with case reports. (St21,Mu29, Go29). However, the irradiation exposureswere 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 forhigh-LET radiation compared to an RBE of 20 x/+ a factor of 5

Underestimate of the doubling dose required inman 

### 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 referenceto low estimate in Table 6-15

(El In reference to high estimate in Table 6-15.

Id) If the most recent analysis of the JapaneseA-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 (seeRu53 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 sensitivityof 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 exposurewas 12.5 R for structuraldefects and 1 R for functional defects. He also suggested human exposurebetween ovulation and about 

### 7weeks gestational age could lead to structural defects, and exposures from about 6 weeks gestational age until birth could lead to functionaldefects. In a later review (Ru71), Rugh suggested structural defects in the skeleton might be induced
as late as the 10thweek of gestation and functional defects as early as the 4thweek. It shouldbe 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 theremay be no threshold for radiationinduced congenital effects in the early human fetus. In the case of humanmicrocephaly (small head size) and mental retardation,at least, some data support this theory (Ot83,84).

However, formost 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 increasewas still greater if there was concurrent exposure to radiosensitizing chemicals such as iodoacetimideor 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 inwhich mice were exposed to 5, 20, or 100 R on the eighth day ofpregnancy. 

### He concluded that the doseresponse function for induction of skeletal effectswas Linear, or nearly linear,with no observable threshold. This appears consistent with 

### a report by Russell (Ru57), which suggested a threshold for some effectswhereas 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 shouldbe 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 dosesmay induce additional abnormalities in components that aremost sensitiveat other stages of development, andmay further modify expression of the changes induced inparts of the embryo at maximum sensitivity during the time of irradiation. In the first case, damagemay 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.

(Theembryonic period, when organs develop, is the period from conception through 

### 7weeks gestational age. The fetalperiod, a time of 

### -- in utero growth, is the period from 8 weeks gestational age to birth.) The different organ and tissueprimordia develop independently and at different rates. However, they are in contact through chemical induction or evocation (Ar54). These chemical messages between cells are important inbringing about orderly development and the correct timing and fitting together of parts of organs or organisms. While radiation can disrupt thispattern, interpretationof the responsemay 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/fetushas a different timetable. In fact, each half (left/right)of an embryo/fetusmay 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 distinctionbetween minor variations of normal and frankmalformation, therefore, is an arbitrary one, and each investigatormust establishhis 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 relevanceof such studies to man. There are species differences in development attributablepartly to the differj-ngcomplexity 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 secondyear 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 glialmultiplication, occurs shortlybefore mid-gestation inman, but at about the time of birth in the rat (Do73). In thls respect, then, the rat ismuch less neurologicallymature 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 toman (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 (ratherthan 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, extrapolationof even qualitative estimates of effectswill 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 shouldbe given as the simplest function that fits the data

(often linear or linearwith a threshold). No attempt shouldbe made to develop complex dose responsemodels 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 programstarted in I.950 i.o st.udy chi]-drenexposed 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)increasedmicrocephaly;

(3)increased mortality, especially i-nfantmortality; (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 ofmicrocephaly and mental. retardation have been investigated to any great extent. In the most recent report, Otake and Schull (Ot83, 84) showed that mental retardationwas particularly associatedwith exposure between 8 and 15weeks 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 4.16t0.4cases 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 shewas 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 gestationwas 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 asmild (IQ50-70), moderate (IQ 35-49), severe (IQ20-341, and profound (IQ

### < 20)

(WH075). However, some investigators use onlymild mental retardation (IQ50-70) 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 retardationmay 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 likemanner, sincemental 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-sis 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 calculatedby Otake and Schull for the Japanese survivors (0t84). Otake and Schull (0t84)gave an estiinate foi: onecase entitled, "The Relationship of Mental Retardation toAbsorbed Fetal Exposure inthe Tensitive' Period When A11 'Contro1s"re Combined." This estimate oE frequency of mental retardation, 0.416 per 

### LOO rads, could be directly applicable toa U.S. population. In this case, the risk estimate would be about four cases of severe mental retardation per 1,000 livebirths per rad of exposure during the 8th and 15thweek of gestation.

The ICRP published an excellent review of biology andpossible mechanisms of occurrence of radiation-induced brain damage, 

### -- inutero (ICRP86). ICRP estimates: (1)for exposures from the 8ththrough the 15thweek after conception, the riskof severemental retardation is 4 x per rad, with a confidence interval of 2.5 x to5.5 x

## lom3,and (2) for exposures from the 16ththrough the 25th week afterconception, the riskof severemental retardation is 1x per rad. However, a thresholdbelow 50 rad cannot be excluded (ICRP86).

If the estimate isapplicable, the low-LET background radiation

. (about 15mrads) delivered duringthe 

### 8- to 15-week gestational age- sensitiveperiod could induce a risk of 6 x 10" cases of severe mental retardation per livebirth. This can be compared toan estimate of a spontaneous occurrence of 0.6 x 10" to3.1

### x cases

- of idiopathic severemental retardation per live birth (EPA84).



### - 6.6.8.2 Teratologic Effects: Microcephaly inHumans

Plumrner (P152) reportedmicrocephaly associated withmental retardation inJapanese A-bomb survivors exposed 

### -- inutero. Wood (Wo65, 66) reported both were increased. The diagnosis of reduced head circumference was based on "normal distribution" statistical theory (Wo66); i.e., ina population, theprobability of having a given head circumference isexpected tobe normally distributed around themean head circumference forthatpopulation.

For example, in a population of live-born children, 2.275 percent willhave a head circumference 2 standarddeviations ormore smaller than themean, 0.621 percent willhave a head circumference 2.5 standard deviations ormore smaller than themean, and 0.135 percent willhave a head circumference 3 standard deviations ormore smaller than themean (statisticalestimates based on a normal distribution).

For most of the studies of the Japanese A-bomb survivors exposed inutero, if thehead circumferencewastwo ormore standard 

### -- deviations smaller than themean forthe appropriate controls inthe unexposed population, the casewas classifiedashaving reducedhead circumference even ifthe datahad not been adjusted for differences in stature (Ta67, Mi72, Wo65). While a definitive relationship between reduced head circumference andmental retardation has not been established, there isevidence that they arerelated.Studies of the Japanese survivors show a re1ationshj.pbetween reducedhead 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 50percent (RERF78)to 70percent (Wo66),while the fraction of those selected for reduced head circumferencewho had mental retardation has been reported as 11 percent (Wo66)to 

### 22percent

(Mi72). Thus, while the relationship appears to exist, it has not been quantified.

Themajority of the cases of reduced head size are observed in those exposed in the first trimester of gestation, particularly the 6th or 7th to 15thweeks of gestation (Mi59,Wo66, Mi72, Wo65, Ta67). Most recently, it has been shown that reduction inhead circumferencewas a linear function of dose (1~84).However, the authors noted that the analysiswas based on T65 dosimetry, and the data shouldbe reanalyzed after completion of the dosimetry reassessment currently in progress. These findings of reduction inhead circumference,with 

### a window of effect in the same timeperiod of gestation asmental 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 farbeen consistent with a linear, no-threshold dose response during the criticalperiod.

6.6.8.3 Other Teratologic Effects

JapaneseA-bomb survivors exposed -- in utero also showed a number of structural abnormalities and, particularly in those who were microcephalic, retarded growth (Wo65). No estimatehas been made of the radiation-related incidenceor dose-response relationships for these abnormalities. However, UNSCEAR (UNSCEAR77)made a very tentative estimatebased on animal studies that the increased incidence of structuralabnormalities in animalsmay be 0.005cases per R per live born, but stated that projection to humans was unwarranted. In 1986,UNSCEAR assumed the risk of an absolute increase ofmalformed fetuses of the order of 5E-3 per rad seen in animalsmight 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 shouldbe 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 formost 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 withthe 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 1986UNSCEAR 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 inhumans. 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 severemental retardation of 4 x lo-' per rad over the period of 8 to 15weeks 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 inman (Table 6-24)



## .
6.7 Summary of EPA's Radiation Risk Factors 

### - A Perspective

Table 6-25 summarizes EPAtsestimate 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 relationshipbetween 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 inutero radiation exposure

Type of Risk Risk per Rad Risk per Event ina toConceptus 100mrad per Year Background

Cancer 5.3 

## lo4 4.8 x

## lo-'

Mental Retardation 4 (exposureat 8 

### - 15weeks)

Mental Retardation 1 (exposureat 16

### - 25 weeks)

Malformation 5 x 5.8 x 10.'

(exposureat2 

### - 8 weeks) Pre-implantation 1x 10 -2 3.8 x Loss (exposureat 

### 0 

### - 2 weeks)

To provide a perspective on the risk of fatal radiogenic cancers and thehereditary damage due toradiation, EPA has calculatedthe risk from background radiation tothe U.S. population using the risk factors summarized inTable 6-24. The risk frombackground radiationprovides a useful perspective for the risks causedby emissions of radionuclides. Unlike cigarette smoking, auto accidents, and othermeasures 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 largelyunavoidable; therefore, it isa good benchmark for judging the estimated risks from radionuclide emissions. Moreover, tothe degree that the estimated risk of radionuclides isbiased, the same bias ispresent inthe risk estimates forbackground radiation.

The absorbed dose rate from low-LET background radiation has three major components: cosmic radiation, which averages about 28 mrad/yr inthe United States; terrestrial sources, such asradium in soil, which contribute an average of 28 mrad/yr (NCRP87); and the low-LET dose resultingfrom internal emitters. The lastdiffers among organs, tosome extent, but for softtissues it is about 24 mrad/yr (NCRP87). Other minor radiation sources such as fallout from nuclear weapons tests, cosmogenic radionuclides, naturally occurring radioactive materials inbuildings, airlinetravel, and
consumer products, contribute about another 7 mrad fora total lowLET whole-body dose of about 87 mrad/yr. The lung andbone receive somewhat larger doses, not included inthe 87 mrad/yr estimate, due tohigh-LET radiations (seebelow). Although extremes do occur, the distribution of thisbackground annual dose tothe U.S. population is relatively narrow. A population-weighted analysis indicatesthatKO percent of the U,S, popuiati.on wou1.d receive annual. doses that
arebetween 

### 75rnrad/gr and 115rnrad/yr (EPA8ll.

As outlined in Section 6.2, the BEIR 111 linear, relativerisk models yield, for 1-ifetirneexposure tolow-LET radiation, an average lifetime riskof 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 tocalculatetheaverage lifetimerisk due tolow-LET background radiation as follows. The average duration of exposure inthis group is 70.7 yr, and at 90 mrad/yr, the average lifetime dose is 6.4 rads. The risk of fatal cancer perperson inthis group is:

or about 0.24 percent of alldeaths. The vital statisticsused in EPArsradiation risk analyses indicatethat theprobability of dying from cancer inthe United States from allcauses is about 0.16, i.e., 16percent. Thus, the 0.24 percent result for the BEIR I11 linear dose response model. indicates that about 1.5 percent of allU.S. cancer isdue tolow-LET background radiation. The BEIR 111 linear-quadratic model indicatesthat about 0.1percent of alldeaths are due tolow-LET background radiation or about 0.6 percent of allcancer deaths.

Table 6-11 indicates a risk of 5.6xlO-'rad-' for alpha emitters inlung tissue. UNSCEAR estimated that in "normal" areasthe annual absorbed dose inthe lungs from alpha emitters otherthan radon decay products would be about 0.51mrad (UNSCEAR77). The individual lifetime cancer risk from thisexposure is:

which is about 1/100 of the risk due tolow-LET background radiation calculatedby means of the BEIR I11 linearmodel.

The 1982 UNSCEAK report indicatesthat the average annual absorbed dose tothe endosteal surfaces of bone due tonaturally occurring, high-LET alpha radiation is about 6mrad/yr, based on a quality factor of 20 and an absorbed dose equivalent of 120mrem/yr (UNSCEAR82). Table 6-11 indicatesthat the individual lifetime risk of fatalbone cancer due tothisportion of the naturally occurring radiation background is:

The exposure due tonaturally occurringbackground radon-222 progeny inthe indoor environment isnot well known. The 1982 UNSCEAR report lists for the United States an indoor concentration of about 0.004 worlcing levels (15Bq/m3) (UNSCEAR82)



## . This estimate isnot based on a national survey and isknown tobe exceededby as much asa factor of 10 ormore in some houses. However, aspointedTable 6-25. Summary ofEPAQ radiation risk factors

Risk Significant Risk.Factor Exposure Period Nominal Range

Low LET (10-"ad-'i

~eratolo~ical:~ Severemental 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 30year 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 exposuremay 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 0.2-0.3WLM/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/106WLM, a lifetime risk of 0.64percent is estimated. For comparison, roughly 5percent 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 4percent of all cancer deaths. This is 2.5 times the 1.61percent 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 inmale smokers, this risk estimate cannotbe 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 105,000per lo6live births, about 10.5percent of live births (NAS80,UNSCEAR82). The low-LET background radiation dose of about 87mrad/year in soft tissue results in a genetically significant dose of 2.6rads during the 30-year reproductive generation. Since this dose would have occurred in a largenumber 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 logmean of this range to obtain an average value for estimating genetic risk. Based on this average value, the background radiation causes about 680genetic effectsper 10' live births (seeSection 6.6). This result indicates that about 0.6 percent of the current spontaneous incidence of serious congenital and genetic abnormalitiesmay be due to the low-LET background radiation.REFERENCES
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### J. R. Parker and V. R. West, Washington, D.C., 1973.

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### 78, Washington, D.C., 1984.

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National Council on Radiation Protection and Measurements, IonizingRadiation Exposure of the Population of the United States,NCRP Report No. 93,Bethesda, MD, 1987.ORNL 8 4

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National Institute for Occupational Safety and Health. Radon Progeny in Underground Mines, DHHS (NIOSH) PublicationNO. 88-101, USPHS, CDC, NIOSH, Cincinnati, Ohio, 1987
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Oak Ridge National Laboratory, Age Dependent Estimation of Radiation Dose, [in press], 1984.
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### RERF

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### St. C1. Renard, Lung Cancer in Swedish Iron Miners Exposed to Low Doses of Radon Daughters,N. Engl. J. Med., 310, 1485-1494, 1984.

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### Atomic Bomb Casualty Commission, Hiroshima, 1966.7. SUMMARY OF UNCERTAINTIES IN DOSESAND 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 calculationof 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 calculationalmethods 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, thisparameter is an important
source of uncertainty.

o For particulate emissions, the rate of deposition of particulates at a given location canvary by over a factor of 10 (seeSection 

### 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 surfacesgenerally contributesmore to the radionuclide concentrations in foods than does uptake from soil,which tends to reduce the importanceof 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 inparticle 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 contributorsto 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 approachesmay be used to describe and quantify uncertainty (Mo78,He79, Cr88, Sc80). One method to assess uncertainty is to assignprobability distributionsto each of the parameters and then to propagate the uncertainties through the dose and risk calculations. These approaches are currently being evaluatedby 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 characterizationof uncertainty is provided, whereby the uncertainty in the key calculational parameters for each category is described. No attempt ismade to establishprobability distributions for the parameters or to propagate rigorously the uncertainties. Instead, approximate ranges of doses and risks are provided. The ranges arebased on inspection of the important contributors to risk and the discussion of uncertaintiesprovided 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 followingdiscussion, a distinction ismade 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 ofmagnitude depending on the soil conditions and the type of plant. These characteristics canvary 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 averagevalue 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 INACTIVEURANIUM 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 exposurehave demonstrated that the offsite exposures are dominated by radon and itsprogeny (DOE86). Accordingly, the uncertainty analysis is limited to radon exposures.

Each of the 24 sites differs to varying degrees in the size and characteristicsof 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 1pCi/m2-secper pCi/g of Ra-226 in the tailings. The concentrationsof Ra-226 in the piles have been extensively studied, and their uncertainty is not an important contributorto 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 entirepile, the uncertainty is reduced due to averaging.

For the fivepiles 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 fluxmeasurements have not been routinelymade 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 associatedwith predicting radon emanation is eliminated because the parameters of interest are average or integratedvalues. For the stabilized piles, the radon emanation ratemay 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 factorsyields the average annual radionuclide concentrationsat the locations of interest. The dispersion factors used in the risk assessments were obtained using site-specificmeteorological 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 conditionsdo 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 theuncertainties in calculating an atmospheric dispersion factor ata specificlocation and point in time may be large (Mi88), when appliedtoaverage annual conditions,
as isthe case for themaximum individual doses, orwhen integrated over time and space, as isthe case for thepopulation doses, the dispersion factors aremore reliable. As discussed inChapter 4, validation studies for Gaussian dispersionmodels indicate an uncertainty of approximately a factor of 2 for average annual concentrations for locationswithin 10km of the release and a factor of 4-10 for locationsbetween 30-140 km of a release.

7.2.3 Location of theMaximum Individual

One of the largest sources of uncertainty incalculatingthe maximum individual doses and risks isuncertainty inthe location of themaximum individual and the duration of exposure. The location of themaximum individual was determined through siteunits. The Agency's ground-rule when estimatingrisk isthat individuals remain atthese 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 fortheir entire lives are given appropriate consideration.

7.2.4 Risk Factor

The risk factor for exposure toradon daughters used inthis risk assessment is 360 lung cancer fatalitiesper 1.OE+6WLM. The uncertainties associated with the risk factor for exposure toradon 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 arrivedatin Chapter 6. These values represent estimates forthe average individual for the case of constant lifetime exposure. The question of sensitive subgroups isnot addressed. Major sources of uncertainty inradon risk factors forthe general public include: (1)uncertainties inthe epidemiological data on miners (especiallywith respect toexposure estimation); (2) uncertainties inthe sex, age and time dependencies of the risk; (3) uncertainties inhow differences inexposure conditions between mines and homes, particularlywith respect tounattached fraction, affect risk; and (4) uncertainties regarding thepossible synergisticeffects of radon and cigarette smoke incausing lung
cancer.
7.3 DOE RADON

Five sites are included inthe DOE radon category: a mill tailings siteinMonticello, Utah, theMiddlesex Sampling Plant site inMiddlesex, N.J., theWeldon Spring siteinWeldon Spring, MO, the Niagara Falls Storage Site inNiagara Falls, NY, and theFeedMaterials Production Center in Fernald, Ohio, As is the case for the uraniummill 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 theMonticello 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 sitevisit 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 chemicalprocessing 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 additionaluncertainty 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 concentrationsby 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 theportions 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 sixmines, the radon release rate was calculatedbased onmonitored levels of radon daughters. The radon concentrationwas estimated assuming a daughter equilibrium fraction of 0.07 observed at othermines. As applied to any one mine, this approach introduces a large uncertainty since the degreeof 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 formines with rugged terrain. For thesemines, the Industrial Source Complex Dispersion (ICS)model was used. Though this model is widely used for complex terrain and release configurations, the resultsmay be uncertain by a factor of 2 or more.

7.6 PHOSPHOGYPSUM STACKS

Like uranium mill tailings, the exposures fromphosphogypsum 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, suchmodels must be used discriminantly and, when possible, shouldbe supplementedwith fieldmeasurements (Be76, Mi75).
Dispersion modeling of radon andparticulates 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 ispreferable toperform field measurements at themaximum individual location to confirmpredicted values. For population exposures, the uncertainties are smaller because the pile behaves effectively as a point source.

For particulate radionuclides,theuncertainties 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 weremade 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 individualvariability. Both these apply primarily to themaximum 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 estimateduncertainty of about a factor of 

### 3 for population exposures. Risk estimates reported for themaximum 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 surfacemines, it was necessary to establish categories ofmines which were used to identify reference mines for use in the risk assessment. One referencemine 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 othermines.
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 calculatemaximum individual and population doses. As applied to population doses, the uncertainties are small because individual site variabilities are averaged out. With regard to themaximum 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 inpathway, 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 largebecause 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 assumptionswere used tomodel resuspensionwhich likely resulted in an overestimate of exposures from thispathway.

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 radionuclidesand pathways contributingto 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 contributorsto uncertainty (DOE75;Me82, Wr86).

7.9.2 Uranium Conversion Facilities

Separate assessmentswere performed for the two uranium conversion facilities. Site-specific source terms, meteorology, demography, and land use were used. Accordingly, the uncertainties associatedwith using model or reference facilities were avoided. The limitingpathways of exposure are the inhalation of Ra-226, Th230, and U-238. The important sources of uncertainty are as described for uraniummills.-1.9.3 Fuel FabricationFacilities

The five operating fuel fabrication facilitieseach have wellcharacterized source terms and site characteristics. The dose and risk assessmentswere 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 inhalationof 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 commercialnuclear power plants in the United States. The radionuclide emissions and site characteristics of each facility are well defined and maintained current in accordancewith regulatory requirements. In addition, each facility is required to demonstrate semiannually that its offsite impacts are maintained within the guidelines set forth inAppendix I to 10 CFR 50. In brief, these guidelines are 5mrem/yr to the whole body or any organ and 25mrem/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 fewmrem/yr. The populati.onimpacts are reported to be about 

### 1 to 10person rems/yr per plant.

For thepurpose of this assessment, average annual source terms for BWRs and PWRs were used. A reference sitewith 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 themaximally exposed individuals,based on values reported by each utility, were less thanAppendix 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 shouldbe 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 andDCF for Sr-90 also contribute to uncertainty. Numerous studies have been published describing the uncertainties in ingestionpathway 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 inNUREG/CR-2612 (NUREG81)to have a geometric mean of 0.33and 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 FACILITIESAND 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 radionuclidesand 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, alongwith 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 ofmodel facilitieswas defined in Section 3.2.2, which is used to extrapolateto the category as a whole. This is the first source of significant uncertainty, especially in terms of the representativenessof themodel facilities for any one hospital. The representativeness issue is significant when attempting to characterizethe doses and risks to the offsite maximum individual at any one facility, and, as is discussedbelow, 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, themodel facilities
are fairly conservative representatives.

The first and perhaps the most important source of uncertainty is the radionuclide emission rates. Source term data were obtainedfrom surveys of 100 hospitals (performedby 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 releasepoints and analyses designed to ensure that the discharges do not exceed the offsitemaximum 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 ofuncertainty is the assumed location of the maximum individual. The analysis is based on the assumption that the nearest individual is 100meters away. This is a reasonable assumption for most facilities;however, for some facilities this assumption may be highly conservative.

Forpopulations, 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 thepopulation doses because these are long-lived radionuclides.

Given the uncertainty in the location of the nearby individual and land use patterns, the uncertainties associatedwith atmospheric dispersion, the exposurepathways (including food), and the dose conversion factors are not likely to contribute significantly to the uncertainty in risk. The external and internaldosimetry 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 individualsand the populations in the vicinity of hospitals is dominatedby uncertainty in the source term. In addition, uncertainties in demographic characteristics and the risk factor are important contributorsto uncertainty.

These observations also apply to most other NRC-licensedfacilities 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 radiopharmaceuticalmanufacturers, 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 risksbeing greater than the indicatedvalues 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 sitesmay 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 factorsdo not contribute significantly to uncertainty. In addition, the food ingestionpathway is not significant. As a result, uncertainties due to these factors are small.

7.10.3 NRC Licensed Facilities-Source Material Licensees

Sections 3.7and 3.8of Volume 11 discuss emissions fromNRC 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 7.9on 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 facilitiesor a conservatively selectednumber of reference facilities. Accordingly, representativenessis not an important
source of uncertainty. However, the environmental behavior and internaldosimetry of uranium and thorium is uncertain because the particle size and chemical composition of the emissions are not welldefined. Nevertheless, conservative assumptionswere used which provides assurance that the doses and riskswere not underestimated.

7.10.4. Non-DOE Federal Facilities

As discussed in Section 3.10of Volume 11, this category consistsprimarily of shipyardsand 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 shipyardsand naval bases.

The airborne emissions from the shipyards are upper bound estimates based onmany years ofmonitoring. 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 themaximum individuals. For the population doses, the facilitieswith the average emissions for each category were used as reference facilities. This approach places an upper bound on the individual risks andmay 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 radionuclidesare variable among sites and as a function of time. The doses to individuals from the food ingestionpathways are also highly uncertain due to uncertainty in the environmental transportparameters 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 radionuclideemissions from industrial boilers are similar to those for utility boilers. For the purpose of the risk assessment, the industrial boiler with the largest particulate releasewas selected as the referenceplant. The radionuclide concentration in the particulate emissions from industrial boilers
was estimated indirectly based on the utility boiler experience. Since the largest industrialboiler was used as the reference facility, and since assumptions applicable to utility boiler emissions are conservative as applied to industrial boilers, thesource term is a generally conservativevaiue formost fncil~ities.

The actual site data at the referenceplants were used. As a result, someplants have higher or lowerpopulation 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 ofVolume 11, there are a large number of diverse DOE facilitieswith 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, ReactiveMetals, Inc. processes uranium ingots and releases particulate uranium, probably uranium oxide, in its multiple stack effluents at the site. The emissions aremonitored; however, the particle size and solubility is not monitored. As a result, the environmental behavior of the uranium and its inhalationdose conversion factors are uncertain. Another source of uncertainty is due to themethod used to model dispersion. The uranium is discharged at three stacks, but is modeled as a single source.

Due to the lack of onsitemeteorological 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 significantlyto overall uncertainty as long as the location of thismaximum individual is well defined. In this case, site specific informationwas 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 separateproduction 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 conservativeassumption 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 locationsmany 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 underestimateof 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 

### 50percent of the calculated dose was due to particulates (mostof the dose was due toAr-41), any overestimate in dose due to these assumptions is less than a factor of 2.

Since the doses are dominatedby external exposure toAr-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 overestimatedose 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 operationalhigh-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 accordwith a defined set of conservative assumptions.

The inhalationpathway was found to be the critical pathway. The critical radionuclidesare H-3, Kr-85 and 1-129 for theMRS and Yucca, and Pu-239 andAm-241 forWIPP. The most important source of uncertainty in these calculations is the assumption made with respect to the likelihoodof accidents occurring and releasing the defined types and quantities of radionuclides.REFERENCES

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Soil Concentration Profiles", Health Physics, 48(4):401,April 1985.
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Bennett, 

### B.G., "TransuranicElement Pathwaysto Man", in Transuranium Nuclides in the Environment, IAEA, Vienna, STI/PUB/410,p.367-382, 1976.
Crick,M.J., et al., Uncertainty Analysis of the Foodchain and Atmospheric Dispersion Modules of MARC, NRPB-R184,. May 1988

Department of Energy, FEIS, Remedial Actions at the Former Climax Uranium Company, Uranium Mill Site, Grand Junction,Mesa County, Colorado, UMTRA-DOE/AL, DOE/EIS0126-F, December 1986.

Department of Energy, Environmental Assessment, Remedial Action at the Tuba City Uranium Mill Tailings Site, Tuba City,Arizona, DOE/EA0317, November 1986.
Dunning, D.E.,Jr., and Schwarz, G, "Viability of human thyroid characteristics and estimated dose from ingested I-131t',Health Physics 40(5) : 625-642.
Fisher, D.L., et al, "The Solubility and Particle Size Distribution of Uranium Oxide Aerosols Associatedwith Uranium Metal Fabrication Facilities", Abstracts from the 31st Annual Meeting of the Health Physics Society. Pittsburgh, Pa, June 29 

### - July 3, 1986,pg S85.

Garten, C.T. 1978. "A review of parameter values used to assess the transport of plutonium, uranium and thorium in terrestrial food chains", Environ. Res, 17(3): 437-452.

Garten, 

### C.T.,Jr., "Statistical uncertainties in predicting plutonium dose to lung and bone from contaminated soils", Health Physics, 39(1): 99-103, 1980.XaDG Hart, K.P.,and Levins, 

### D.M., "Steady State 

### Rn Diffusion Through Tailings and Multiple Layers of Covering Material", Health Physics, 50(3):369. March 1986.
Ibrahim, S.A.,and Wicker, F.W., "ComparativeUptake of U and Th by Native Plants at a 

### U. Production Site", Health Physics, 54(4):413,April 1988.

International Commission on Radiological Protection, Lung Cancer Risks from Inhalation Exposure to Radon Daughters, ICRP 50, 1987.
Johnson, J.E., et al., "Transfer Coefficients of Selected Radionuclides to Animal Products. I. Comparison ofMilk and Meat from Dairy Cows and Goats", Health Physics, 54(2):161, February 1988.
Landa, E.R, "Radium-226 Contents and Rn Emanation Coefficients of Particle-Size Fractions ofAlkaline, acid and Mixed 

### U. Mill Tailings", Health Physics, 52(3):303, March 1987.
Linsalata, P., Eisenbud,M., and Penna Franca,P., "Soil-to-Edible Vegetable and Soil-to-Farm Animal Tissue ConcentrationRatios for Ra, Th, U and Rare Earth Elements", Abstracts of the 33rdAnnual Meeting of the Health Physics Society, Boston, Mass., July 4 

### - 8, 1988, pg S51.
McDowell-Bayer, L.M., and Travis, C.C., "A Review of parameters describing terrestrial food-chain transport of Lead-210 and Radium-226", Nuclear Safety 21(4):486-495, 1980.
Mewhinney, J.A.,Griffith, W.C., and Muggenburg, B.A., "The influence of aerosol size on retention and translocation of 241 

### Am following inhalation of 24AM02 by beagles", Health Physics, 

### 42(5) :611-627, 1982.
Milhorn, R.C., et al., "Measured Plutonium Resuspension and Resulting Dose from Agricultural Aerations of an Old Field at Savannah River Plant in S.E.USA", IAEA, Vienna, STI/PUB/410,

### p. 409, 1975.
Miller, C.W., et al., "Examinationof the Uncertainty in Air Concentration Prediction Using the Hanford Field Data", Health Physics, 55(2):443, 1988.Morgan, M.G.,et al., "Technical Uncertainty in Policy Analysis," Chapter 6 in Final Report to Division of Policy Research and Analysis, 

### NSF Grant PRA-7913070, 1982.
Morris, R.C., and Fraley, L Jr.,"Environmental Effects of Rn-222 Fluence Rate from Unreclaimed Uranium Tailings,"Health Physics, 52(Suppl 1):550,Abstract of Papers presented at the 32nd Annual Meeting of the HPS, Salt Lake City, Utah, July 5 

### - 9, 1987.

Morgan, M.G., et al., "Aprobabilistic methodology for estimating air pollution health effects from coal fired power plants," Energy System and Policy 2(3):287,1978.
Murphy, B.D., and Ohr, S.Y., "The Limitations of Atmospheric DispersionData and Their Contributionto Uncertainties in Dose Assessment." Health Physics 48(3):315, March 1985.

NCRP8 5 National Council on Radiation Protection and

~~~~ ~ ~-~. Measurements, Evaluation of Occupational and Environmental Exposures to Radon and Radon Daughters,NCRP Report No. 78,May 31, 1984.
Ng82 Ng, Y.C. "A review of transfer factors for assessing the dose from radionuclides in agricultural products," Nuclear Safety 23(1): 57 

### - 71, 1982.
Ng77 Ng, Y.C., et al., "Transfer coefficients for the prediction of the dose toman via the forage-cow-milk pathway from radionuclides released to the biosphere," 

### U.S. DOE Report UCRL-51939, 1977.

NUREG79 U.S. Nuclear Regulatory Commission, A statistical analysis of selected parameters for predicting food chain transport and internaldose of radionuclides. Hoffman, F.O., and Baes, C.F., I11 (eds),NUREG/CR-1004, 1979.

NUREG81 U.S. Nuclear Regulatory Commission,Variability in dose estimate associatedwith food chain transport and inqestion of selected radionuclides,by F.O. Hoffman, R.H. Gardner, and K.F.Eckerman,Schwarz, G., and iioffnan, 

### F.O.,"Imprecision of dose predictions for radionuclides released to the environment:an application of a Monte-Carlo simulation technique." Environment International, 4: 280-297, 1980.
Schwarz, 

### G., and Dunning, D.E., Jr.,"Imprecision in estimates of dose from ingested 137 Cs due to variability in human biological characteristics." Health Physics, 43(5) : 631-646, 1982.
Sheppard,M.I.,"RadionuclidePartitioning Coefficients in Soil and Plants and Their Correlation." Health Physics, 49(1):106,July 1985.
Sykes,R.I., and Lewellen, W.S., "Results from Two Model Evaluation Studies", abstract from the 31stAnnual Meeting of the HPS, Pittsburgh, Pa.,June 29

## - July 3, 1986, Health Physics, 5(Supp 1):S32,June 1986.
Tadmor, 

### J. "Sensitivity Analysis of the Influence of Source Term and Environmental Parameters on the Radiological Risk of Coal-Fired Plants," Health Physics, 51(1):61,July 1986.

Tracy, B.L.,Prantl, F.A.,and Quinn, J.M., "Transferof 226Ra, 210Pb and uranium from soil to garden produce: assessment of risk," Health Physics, 44(5): 469-477, 1983.
Whicker, F.W., and Kirchner, T.B., "Uncertainty Analysis of Predicted 1-131 and Cs-137 Concentrations inMilk from Fallout," abstracts from the 33rdAnnual Health Physics Society Meeting, Boston, Mass., July 4 

### - 8, 1988,Vol. 54 (Supp1):S20.

Wrenn, M.E.,"Metabolism and Toxicity of Uranium," ibid. S82.

### APPENDIX A

ASSESSMENTMETHODOLOGYA.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 PATHWAYMODELING

A.2.1 Individual Assessment

The nearby individuals were assessed on the followingbasis:



### (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 concentrationsby 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



## .
TableA-1. Presumed sources of food for urban and rural sites.

Food Urban/Lowproductivity Rural

Vegetables Meat MilkF1 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 concentrationsdue 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 considerationwas 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 100years 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 summarizesAIRDOS-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 reactingbody 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 Stateswith 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 isbased 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 factorsdo 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) Surfacebuildup 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) Exposureperiod-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 fraction-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)
Consumption time delay-meat (days) TSUBSTableA-2. AIRDOS-EPA parameters used for generic siteassessments (Continued)

Symbolic variable Description Value

FSUBG

FSUBL

TSUBB

P

TAUBEF

MSUBB VSUBM Produce fraction (gardenof interest) Leafy veg fraction (gardenof interest)

Soilbuildup 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 fractionleafyvegetablesTable? 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 0.0at. A-i,~:t,~iiii v~iiric!:; iisiiii fox. e?i<inteiiLdependen?.factors. ((?ani:irrt~ed)

Nb Y I..OE-7 

### Nd Y 3.OK4 Ni. 

### W 5. OF-? NP 

### W I..OE-Z 0 I 9.5E-1.

P0 

### W 1 .OR-1 Pr Y ?.OR-4 PI1 

### Y I. OR-?* RH 

### W 7. OE-I Kh 

## 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.

Sh W 1 .O1!-1 SC Y 1 .O4-4 Sf? 

### W R .OR-] 

### sn~ 

## w ?.RE-4 Stl 

### W ?.OR-2



### sr 

## 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.IE-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-3Table A-4. Cattledensities and vegetable crop distributions for use withAIKDOS-EPA

Dairy cattle State density iilkm2

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 Maryland 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/km2TableA-4. Cattledensitiesand vegetable crop distributions for use withALRDOS-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/rciyr) (mz/rCiyr)Risk factors for selected radionuclides (seeTable A-3 for default inhalation class and ingestion f, values). (Continued) Table A-5.

Nuclide Inhal. Ingest. Immer


## . Surface (rCi-') (r~i-') (m3/@Ciyr) (m2/rciyr)Table

### A-5. Risk factors for selected radionuclides (seeTable A-3 for default inhalation class and ingestion 

### f, values)



## . (Continued)

Nuclide Inhal. Ingest. Immer


## . Surface (pCim') (~ci-') (m3/pciyr) (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-05REFERENCES



### 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 forEstimating 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.

NRC75 Memo from 

### K. Eckerman,N. Dayem, R. Emch, Radiological Assessment Branch, Division of Technical Review, Nuclear Regulatory Commission, Code Input Data forMan-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.

Sj84 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 StatesDepartment ofAgriculture, 1972, Food Consumption of Households in the United States (Seasonsand Year 1965-1966), Household Food Consumption Survey 1965-1966, Report No. 12,

### Agricultural Reseach Service, USDA, Washington, DC (March1972).

### APPENDIX B

MECHANICS OF THE LIFE TABLE IMPLEMENTATIONOF THE RISK ESTIMATESB.1 INTRODUCTION

This appendix describes the mechanics of the life table implementationof the risk estimates derived in Chapter 

### 6.

B.2 LIFE TABLEANALYSIS TO ESTIMATE THE RISK OF EXCESS CANCER

Radiation effects canbe classifiedas 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 forwhich 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 (solidtumors and leukemia) and the expression, in later generations, of genetic effects. In order to estimate these effects, instantaneousdose rates for each organ at specified times are sent to a subroutineadaptation 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 thismethodology 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 actualmortality 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 (latencyperiod) before a
cancer is clinically observed. Following the latencyperiod, the probability of occurrence of a cancer during a given year is assumedto 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 victimmay experienceyears 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 concentrationof 1pCi/cm3in air or 1 pCi/cm2on ground surfaces. Since themodels used in RADRISK are linear, these resultsmay 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 110years. The span from 0 to 110years 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 incrementalprobability PR of dying during that year due to radiation-induced cancer of the given organ. In general, for them-th year, the calculations are:



### M (m) = total number of deaths in cohort during yearm,

= [PN(m)

## + PR(m) x N(m)Q(m)

= incremental number of deaths duringyearm due to radiation-induced cancer of a given organN(m+l) = number of survivors atthebeginning of yearm il

= N(rn) 

### - M(m) (N(l)=100,000)



## .
PR isassumed tobe small relative toPN, an assumption which isreasonable only for low-level exposures (BuEl), such asthose consideredhere. The total number of incremental deaths forthe cohort 

### isthen obtainedby summing Q(m) over allorgans for 110 years.


### Inadditiontoproviding an estimate of the incremental number of deaths, the lifetablemethodology can be used toestimate the total number of years of lifelosttothose dying of radiation-induced cancer, the average number of years of lifelostper incremental mortality, and the decrease inthepopulation's lifeexpectancy. The totalnumber of years of lifelosttothose dying of radiation-induced cancer 

### is computed asthe difference between the totalnumber of years of lifelivedby the cohort assuming no incremental radiation risk, and the totalnumber of years of life livedby the same cohort assuming the incremental risk from radiation. The decrease inthepopulation's lifeexpectancy can be calculated asthe total years of lifelost dividedby tha original cohort size (N(1)=100,000),

Either absolute or relative risk factors can be used. Absolute risk factors, given interms of deaths per unit dose, arebased on the assumption that there is some absolute number of deaths in

### a population exposed ata given age per unit of dose. Relative risk factors, thepercentage increase inthe ambient cancer death rate per unit dose, arebased on the assumption that tha annual rate of radiation-induced excess cancer deaths, due toa. specific type of cancer, isproportional tothe ambient rate of occurrence of fatal
cancers of that type. EitBer the absolute or the relative risk factor is assumed toapply uniformly during a plateau period, beginning atthe end of the latentperiod.

The estimates of incremental deaths inthe cohort from chronic exposure are identical tothose obtained if a corresponding stationarypopulation (i.e., a population inwhich equal numbers of persons areborn and die ineach year) issubjected toan acute radiation dose of the same magnitude. Since the totalperson-years livedby the cohort inthis study isapproximately 7.07 million, the estimates of incremental mortality inthe cohort from chronic irradiation alsoapply toa one-year dose of the same magnitude toapopillati.on of this size, age distribution,and age-specific mortality rates. More precise life table estimates for a specific popu:l.ationcan be obtained by altering the structure of the cohort to reflect the age distribution of a particularpopulation at risk.



### In addition, since the stationary population is formedby superposition of all age groups in the cohort, each age group corresponds to a segment of the stationarypopulation 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 ratewould be numerically equal to that calculated for the stationary population exposed to an annual dose of the samemagnitude. Thus, the risk estimatesmay be reported as a lifetime risk (thecohort 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 fromyear to yearmay be obtainedby 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 IncreasedRisks 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 PublicationNo. (HRA)75-1150, Public Health & Service, Health Resources Administration, National Center for Health Statistics, Rockville, Maryland, 1975.

ICRP77 International Commission on Radiological Protection, "Recommendationsof 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 EKVIR3NMENTALRISK ASSESSI<EN'I'S

C.l INTRODUCTION

The doses and risks attributable to airborne emissions from the various facilities and categories of facilities addressed inVolume I1 have been estimated using themodels and assumptions described in this volume. The calculational methods usemonitored data characterizing airborne emissions and then apply mathematical models to estimate the radionuclide concentrationsand 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 ofmrem/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 tomeasure 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 thepopulation. Accordingly, the actual orpotential 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 shouldbe 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, themodels used to calculate doses and risks are generalizations and simplifications of theprocesses which result in exposure and risk. In addition, our ability to model theprocesses is also limited to a degree by the availability of data characterizing each site and our understanding of theprocesses.
In Volume 11, doses and risks for each category arepresented
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 valuespresented, however, areof 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-6Lifetime 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 emissionmay 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 occurswhen, due to limited information and uncertainty in the calculational parameters, conservative assumptionsare 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 inVolume I of the 

### BID. This appendix summarizes the quantitativeuncertainty analysis techniques currently under review by the Office.

C.2 QUANTITATIVEUNCERTAINTY ANALYSIS

The use of quantitative uncertainty analysis to address environmental risks became widespread followingthe Reactor Safety Study (NRC751,and was recommended by theAgency in support of environmental risk assessments in 1984 (EPA84). The technique results in a range ofvalues 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 canbe 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 therebybe expressed as "best estimate" values, 90percentile values or 99percentile values, etc. Figure C-l presents an example of the output of such an analysis. The results are expressed as a cumulativeprobability distribution. Inspectionof 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 1mortality per 10,000 years, and that the best estimate (i.e.,the 50percentile value) is less than 0.1 fatalityper 10,000 years.
Though the concept is simple, the implementation and interpretationof uncertainty analyses performed in support of environmental risk assessment has evolved into an area of specialization founded in work performed at CarnegieMellon University (Mo78). The use of quantitativeuncertainty 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 theoutput 

### of a riskassessment using quantitativeuncertainty analyses 

### (fromRi83).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 approachmost appropriately applied to the analysespresented inVolume I1 of this BID. Each application uses a somewhat different calculational approach and set of input data. The appropriatenessof 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 uncertaintiespertain to the formulationof mathematical models used to predict risk and the degree to which they accurately represent reality. One way to address this source of uncertainty is toperform 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 alternativemodels in light of the available data and to assignprobabilities 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 considerationof 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 formulationofmetabolic models is a serious problem in estimating dose conversion factors formany radionuclides. Modeling uncertainty for dispersion and pathway calculationspose similar problems. As a result, theAgency'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 issuehas to do with whether all significant radionuclidesand pathways of exposurehave 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 ismorelikely 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 contributorto overall uncertainty.

Uncertainties in the values of the calculational input

y in input parameters, such as dose and risk ion of both parameter andmodeling

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 theseparameters, along with appropriate unit conversions, results in an estimate of lifetime cancer risk due to exposure. Each of the fiveparameters has some degree of uncertainty, which contributesto the uncertainty in the calculated risk.

The source term (Ci/yr)is itself an estimated value which varies as a functionof time. However, since this is a lifetime risk, it is necessary to estimate the uncertainty in the average annual release rate overmany years. This distinction is important because it virtually eliminates the need to explicitly consider uncertainties associatedwith the time-varying nature of the source term. If the concernwas with themaximum 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 measurementsmade over the area of the stack over prolonged periods of time, the source term couldbe accurately defined. However, the source term has been approximated using a limited number of samples and a conservative set of assumptionswhich provides assurance that the real source term has not been underestimated. In a quantitativeuncertainty 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 ismultiplied by the release rate in Ci/yr, alongwith 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 themethod 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 themagnitude of this source of uncertainty ismuch 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. Consideringall of these factors,an uncertainty distribution is developed for the atmosphericdispersion factor. Note that the distribution of the atmospheric dispersion factors for themaximum 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 iswell characterized, and constructing a reasonableprobability distribution for this parameter will be a relatively straight forward exercise.

The fourthparameter, 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 thisparameter are based on epidemiological data and only apply to large populations. It is assumed that themaximum individual has the average radiosensitivity, and a risk factorprobability 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 isbeing estimated. Is the risk for a real or hypothetical person, is it themaximum or the average risk, and is it the current
or possible future risk that is of concern? The individuals constructing the distributionsmust 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 assignedprobability distributions. Thesedistributions are used as input tomodels that propagat.e the uncertainties.



### C.5 TECHNIQUES FOR PROPAGATINGUNCERTAINTIES

The basic approachesused to propagate uncertainties are method of moments techniques, or Monte Carlo techniques. Method ofmoments is the standard method for propagating error described in fundamental texts on statistics. Thismethod propagates errors by calculating a linear combination of themoments 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 themethod ofmoments is the use of a Monte Carlo, orMonte Carlo type, analysis. This approach can consume considerable computer resources but has the potential to yieldmore satisfying results. The technique calculates risk in the same manner as described above, except it performs the calculationmany times, each time randomly selectingan input value from each of the probability distributions representing each parameter. The output is a risk distribution. The more times the calculation is performed, themore complete the results. The number of repetitions will determine the precision of the output. Themore 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 aremost 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 issuepertinent to the implementationof a quantitative uncertainty analysis is thecompleteness and reliability of the data characterizing the distributionsof each of the calculationalparameters. The number of radionuclides, pathways and parameters used in the risk assessments (seethe AIRDOS input sheets in the Appendix to Volume 11) is very large. However, through a screeningprocess, the number of radionuclides and pathways that require explicit analysis canbe 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 distributionshave been constructed in the past which will facilitate thisprocess 

### (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. StatisticalAssoc., 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 (July1985).

Morgan, M. G., et al., "Sulfur control in coal-fired power plants: A probabilistic approach topolicy analysis," APCA Journal, 28(10) 993, 1978.

Moraan, M.G., and Henrion, M., Uncertainty:A

d. 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 ofAccident Risks in United StatesCommercial 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., "Analysesof Uncertainties in the EPA Ore Body Release and River Mode Exposure Pathway Models Used as the Bases for Proposed Geologic Repository Release Limits", FinalReport to Battelle Project Manager Division (BPMD) for the Department of Energy, June 

### 10, 1983.