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Health Risks of Radon and Other Internally Deposited Alpha-Emitters: BEIR IV (1988)
Commission on Life Sciences (CLS)

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2 Radon INTRODUCTION Of the several isotopes of radon, radon-222 has the most im- portant impact on human health (see the box entitled "Isotopes of Radons. An inert gas at temperatures above -61.8°C, radon-222 is a naturally occurring decay product of radium-226, the fifth daugh- ter of uranium-238. Both uranium-238 and radium-226 are present in most soils and rocks in widely varied concentrations.2t As radon forms from the decay of radium-226, it can leave the soil or rock and enter the surrounding air or water. Radon gas thus becomes ubiquitous, and its concentration is increased by the presence of a rich source and by low ventilation in the vicinity of a source. As illustrated in Figure 2.1, radon decays with a half-life of 3.82 days into a series of solid, short-lived radioisotopes collectively referred to as radon daughters or progeny (Figure 2-1) (see Annex 2B). Two of these daughters, polonium-218 and polonium-214, emit alpha parti- cles, which, when emission occurs in the lung, can damage the cells lining the airways. The resulting biological changes can ultimately lead to lung cancer. Underground mining was the first occupation associated with an increased risk of lung cancer. Uranium ores contain particularly high concentrations of radium, and radon-daughter exposure has been associated with Jung cancer in uranium miners. Miners of other types of ore can also be placed at risk by the combination of a sufficiently strong source of radon and inadequate ventilation. 24

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RADON 25 ISOTOPES OF RADON The committee's discussion of radon is limited to radon- 222, the most common isotope. Other radioisotopes of radon- radon-219 (actinon) and radon-220 (thoron) occur naturally and have alpha-emitting decay products. Actinon has an ex- tremely short half-life (3.9 s). Accordingly, concentrations of actinon and its daughters are extremely low, and decay of acti- non contributes little to human exposure. Because of its short hal£life (56 s), the concentration of thoron is also usually low. Dosimetric considerations suggest that the dose to the tracheo- bronchial epithelium from thoron progeny is, for an equal con- centration of inhaled alpha energy, less by a factor of 3 than that due to the progeny of radon-222.26 The potential for Jung cancer due to inhalation of thoron cannot be addressed directly, because the available epidem~ological data are based almost exclusively on exposures to radon-222 and its daughters. Radon progeny are also present in the air of dwellings. Their source Is the underlying soil, but building materials, water used routinely in the building, and utility natural gas also contribute. The concentration of radon progeny in dwellings is highly variable and depends mainly on the pressure in the house and on the ventilation. Because of their wide distribution, radon daughters are a major source of exposure to radioactivity for the general public, as well as for special occupational groups. The estimated dose to the~bronchial epithelium from radon daughters far exceeds that to any other organ from natural background radiation.20 The recent recognition that some homes have high concentrations of radon has focused concern on the potential lung-cancer risk associated with environmental radon. Measured concentrations of radon in homes in the United States appear to follow a log-normal distribution.24 In addressing the risks associated with radon exposure, the com- mittee responsible for this report considered the extensive infor- mation accumulated during nearly a century of research on radon. Epidemiological studies have described the risks associated with radon-daughter exposure of underground miners; animal studies have provided complementary data; and experimental and theoretical re- search has provided insights into radon-daughter carcinogenesis. The

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26 MASS NUMBER I t620y , 1 RADON ~ 3.82d 218 214 210 206 HEALTH RISKS OF RAD ON AND O THER ALPHA-EMT TTERS . _ . ... '..., 3.05m · ·.~ ASTATINE ~ RADON (RAD!UM A) 0.019% . . .,, .. c~4 ~ ~ LEAD ~ ~ 1 ~ BISM ;T;I · ' 26.8m · ~ ' 19.7m . . . (RADIUM B) ~ _ (RADIUM C) /3 ,POLOI IIUM.' . ~ 0.0001648 _ (RADIUM C' _ - · ~ THALLIUM | ~ _ 3.1m THE SHORT-LIVED RADON DAUGHTERS LEAD ,l3 _ r 22y ~ .. cx 0.000002% a 0.00013% crow ~1: I POLONIUM | Sd r 138d z~ ~ FIGURE 2.1 The radon decay chain. An arrow pointing downward indicates a decay by alpha-particle emission; an arrow pointing to the right indicates a decay by beta-particle emission. The historical symbols for the nuclides are in parentheses below the modern symbols. Most of the decays take place along the unbranched chain marked by the double arrows. The negligible percentage of the decays going along the single arrows is shown at critical points. The end of the chain, lead-206, is stable, not radioactive. Half-lives are shown for each isotope with ~ = seconds, m = minutes, d = days, and y = years. research, described in detail in the appendixes of this report, is briefly summarized below. DOSIMETRY By convention, the concentration of radon daughters is mea- sured In working levels (WL), and cumulative exposures over time are measured in working-level months (WI.M) (see box entitled ~Spe- cial Quantities and Units for Radon Exposures"~. As described in Annex 2B, the relationship between exposure, measured as WI,M,

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RADON 27 SPECIAL QUANTITIES AND UNITS FOR RADON EXPOSURES The working level (WL) is defined as any combination of the short-lived radon daughters in 1 liter of air that results in the ultimate release of 1.3 x 105 MeV of potential alpha energy. As detailed in Annex 2B, this is approx~rnately the amount of alpha energy emitted by the short-hal£life daughters in equilibrium with 100 psi of radon. Exposure of a ratter to this concentration for a working month of 170 h (or twice this concentration for half as long, etc.) is defined as a working-level month (W~M). Note that the cumulative exposure in W[M ~ the sum of the products of radon-progeny concentrations and the tunes of exposure. For historical reasons, time is quantified into blocks of 170 h when the concentration is expressed in WL. This can lead to confusion in domestic environments, because a 3(}day month is 720 h. Exposure to ~ WL for 720 h results in a cumulative exposure of 4.235 W[M. Home occupancy for 12 in/day at 1 WL would result in a cumulative exposure of about 2.12 W[M per month of occupancy. and dose to target cells and tissues in the respiratory tract is ex- tremely complex and depends on both biological and nonbiological factors.20 Because of differences in the circumstances of exposure, it cannot be assumed a priori that exposure to 1 W[M in a home and to ~ W[M in a mine will result in the same dose of alpha radiation to cells in the target tissues of the respiratory tract. Thus, an under- standing of the dosimetry of radon daughters in the respiratory tract is essential for extrapolating risk estimates derived from epidem~o- logical studies of miners to the general population in indoor domestic environments. Factors influencing the dosimetry of radon daughters include physical characteristics of the inhaled air, breathing patterns, and the biological characteristics of~the Jung (Table 2.1~. Radon daughters are initially formed as condensation nuclei. Although most of these attach to aerosols immediately after forma- tion, a variable proportion remain unattached and are referred to as the unattached fraction. This fraction is an important determinant of the dose received by target cells; as the unattached fraction in- creases, the dose also increases because of the efficient deposition of

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28 HEALTH RISKS OF RADON AND OTHER ALPNA-E3MITTERS TABLE 2-1 Factors Influencing the Dose to Target Cells in the Respiratory Tract from Radon Exposure Characteristics of Inhaled Air Fraction of daughters unattached to particles Aerosol characteristics Equilibrium of radon with its daughters Breathing Pattern Tidal volume Respiratory frequency Nose or mouth breathing Characteristics of Lung Bronchial morphometry Mucociliary clearance rate Mucus thickness Location of target cells the unattached daughters in the airways. The particle size distribu- tion in the inhaled air also influences the dose to the airways, because particles of different sizes deposit preferentially in different genera- tions of the Jung airways. The specific mixture of radon daughters also affects the dose to target cells, but to a smaller extent. The amount of radon daughters inhaled varies directly with the minute ventilation, i.e., the total volume of air inhaled In each minute. The deposition of radon daughters within the lungs, however, does not depend in a simple fashion on the minute ventilation, but varies with the flow rates in each airway generation. The flow rates vary with both tidal volume and breathing frequency. The proportions of oral and nasal breathing also influence the relationship between ex- posure and dose. A substantial proportion of the unattached radon daughters deposits in the nose with nasal breathing, whereas it is likely that a smaller fraction deposits in the mouth with oral breath- ~ng. Characteristics of the lung also influence the relationship be- tween exposure and dose (Table ~-1~. The sizes and branching pat- terns of the airways affect depositions and can differ between children and adults and between males and females. The rate of mucociliary clearance and the thickness of the mucus layer in the airways also enter into dose calculations, as does the location of the target cells in the bronchial epithelium. As outlined in Part 3 of Appendix VIT, smoking and presumably other pollutants modulate these factors. The effect of the physical and biological factors outlined in Table 2-1 on the dosimetry of radon daughters can be estimated by computer modeling (see Annex 2B). The committee used the results of such

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RADON 29 models to provide guidance on estimating the risk of lung cancer due to radon in indoor environments. HUMAN AND ANIMAL STUDIES The association of radon-daughter exposure with human lung cancer has been the subject of extensive epidemiological studies of underground miners. The {ung-cancer hazard faced by underground miners was first recognized by Harting and Hesse in 18797 on the basis of their autopsy observations of European miners. Excess Jung- cancer occurrence has been found in uranium miners in the United States, Czechoslovakia, France, and Canada and in other under- ground miners exposed to radon daughters, including Newfoundland fluorspar miners, Swedish metal miners, British iron and tin miners, Trench iron miners, Chinese tin miners, and American metal miners (see Appendix IV). Epidemiological studies of these mining groups have shown increasing lung-cancer risk as cumulative exposure to radon daughters increases and have provided some insights into the combined effects of cigarette smoking and radon-daughter exposure (see Appendix VIl). Exposures of animals to radon and its daughters have con- firmed that exposure to radon daughters causes lung cancer (see Ap- pendix HI). Animal experiments have also provided data on exposure- response relationships and on the modifying effects of exposure rate and the physical characteristics of the inhaled radon (see Appendix NIT). Animal models have proved less useful for studying the inter- action of radon daughters with cigarette smoking because of the difficulties of replicating smoking patterns in humans with animals. The committee also considered relevant information from the extensive literature on the biology and epidemiology of lung cancer. This malignancy, although relatively uncommon at the start of the twentieth century, has become the leading cause of cancer death in the United States.35 Most lung cancers are caused by cigarette smok- ing; only 5-10%0 of the total cases occur in lifelong nonsmokers.35 38 In cigarette smokers, the risk of developing lung cancer increases with the number of cigarettes smoked daily and with the number of years of smoking.3 35 The risk of Jung cancer for a smoker is some 10 times higher than that for a nonsmoker, and up to 20 times higher for heavy smokers. Because cigarette smoking predominates as the cause of Jung cancer, the committee needed to address separately the risks of radon-daughter exposure for smokers and for nonsmokers.

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30 HEALTH RISKS OF RADON AND OTHER ALPHA-EMITTERS The committee's analyses of the interaction between radiation and smoking are cliscussed in Part 2 of Appendix VIl. The committee was faced with the challenge of using the epidemi- ological and experunental evidence in combination with its under- standing of the dosimetry of radon daughters to address the topical issue of exposure In domestic environments. This followed from the charge to the committee to develop risk coefficients applicable to exposure to radon in homes. The rationale for this charge emerges from recognition of the potential for radon in domestic environments as a public-health hazard. The urgency of completing this task has increased as more data have become available on indoor radon in the United States. Radon concentrations in American homes have not been systematically surveyed, but available data indicate that some homes have levels approaching or greater than the control lev- els in underground mines.25 Radon daughters are also a hazard for underground miners. Thus, the committee was also charged with ad- dressing the risks for occupational exposure to underground miners. Other expert groups and individual investigators have derived risk estimates for radon-daughter exposure (see Appendix VIl). The approaches have been diverse and based variously on the epidemi- ological data, on extensive dosimetric modeling, and on informal expert judgment. The resulting risk estimates have a wide range. The committee did not select any of the published lung-cancer risk estimates associated with exposure to radon and its progeny as ap- propriate for meeting its charge. THE COMMITTEE'S APPROACH TO ESTIMATION OF LUNG- CANCER RISK Evaluation of the lung-cancer risk associated with radon daugh- ters was the most challenging task faced by the committee. Nu- merous studies of underground miners exposed to radon daughters have shown an increased risk of lung cancer, in comparison with un- exposed populations. Animal studies have confirmed this risk, and the development of multicomponent dosimetric mo~lels has provided an understanding of factors influencing the carcinogenic potential of exposure. However, the human data are for occupational exposure in an underground environment and do not address directly the risks at the generally lower levels of exposure that are typically of concern in the home.

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RADON 31 Two approaches that have been used previously to characterize the risks associated with radon-daughter exposure were considered by the committee: dosimetric models of the respiratory tract and statistical models applied to one or more of the epidemiological data sets. The dosunetric approach provides an estimate of risk that is based on modeling the dose to target cells in the respiratory tract. Di- verse dosimetric models have been developed; all require assumptions (some not subject to direct verification) concerning the deposition of radon daughters in the Jung and the nature and location of the tar- get cells for cancer induction. Additional assumptions concerning the carcinogenic potential of alpha radiation in the respiratory tract are also required. The comrn~ttee preferred an epidem~ological approach that provides risk coefficients based directly on the substantial body of available human data. While the committee did not use dosimet- ric models for calculating the lung-cancer risk coefficients, it found such models useful in applying its risk model, derived from studies of underground miners, to the general population. Rather than basing its risk estimates solely on review of pub fished reports, the committee obtained and analyzed the original exposure and follow-up data from four of the most important epi- demiological studies of underground miners (see the next section). Each of the studies has limitations, but a combined analysis of these major data sets permitted a comprehensive assessment of the risk associated with radon-daughter exposure and other factors influenc- ing this risk. In analyzing the epidem~ological data, the committee used a descriptive approach, rather than methods based on models of carcinogenesis. The comm~ttee's analytical approach was appro- priate for meeting its charge with a minimum of assumptions as to the underlying mechanisms of cancer initiation and promotion. Al- though a few epidemiological studies of Jung-cancer risk associated with indoor domestic exposure to radon have been reported, these studies have been preliminary and are inadequate for the purpose of risk estimation. In the future, however, epidemiological studies of indoor exposure may serve as a basis for lung-cancer risk estimates. THE COMMITTEE'S ANALYSIS OF THE RISK OF LUNG CANCER ASSOCIATED WITH EXPOSURE TO RADON PROGENY The committee's risk estimates for radon-daughter exposures are based largely on its own reanalysis of the four principal data sets on

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32 HEALTH RISKS OF RADON AND OTHER ALPHA-EMITTERS the epidemiological follow-up of underground miners. The commit- tee obtained data based on two Canadian uranium-miner cohorts, Eldorado Beaveriodget° and Ontario;~7~~9 on Swedish iron miners, MaImberget;29 and on Colorado Plateau uranium miners. 9 i2 t3 37 (see box entitled "Characteristics of the Four Underground-Miner Groups Analyzed by the Committees. The committee attempted to obtain comparable data from the studies of Czechoslovakian ura- nium miners, but was unsuccessful. For the first three of the cohorts, we obtained data on individual miners; for the Colorado cohort, we were able to obtain only detailed summaries of the type described below. Some of the miners in the Colorado cohort were occupa- tionally exposed to radon progeny prior to their employment in uranium mines. Although exposures have been estimated for this earlier period,~3 their accuracy is uncertain and they were not con- sidered in the committee's analysis. Cigarette-smoking information on all exposed subjects was available only for the Colorado cohort, so the comm~ttee's primary dose-response analysis does not include this factor. However, the committee's analyses of the combined effects of smoking and radon exposure on the Colorado cohort, described in Appendix VII, support the results presented here (and in Annex 2A). Recent developments in statistical methods have provided better approaches for analyzing data from occupational-cohort studies than were available when these data were first analyzed. With these mul- tivariate methods for analysm of the data, it is possible to examine systematically many aspects of risk estimation that have been the most uncertain in the past, particularly the temporal patterns of excess risk. By analyzing the combined data from the four cohort populations, it was the comm~ttee's intent to gain a clearer under- standing of appropriate models for describing the risk associated with exposure and to obtain a more meaningful comparison of the risks in these primary cohorts. The committee first carried out separate but parallel analyses of the four cohorts to gain a clearer understanding of the determi- nants of risk within each. The committee then carried out a formal analysis of the combined cohort data to obtain better estimates of the effects that seemed important and consistent in the separate analyses. This approach led to the development of a relative-risk time-since-exposure model, based on the combined data, which is more complex than ordinarily used for estimating radiation risks. However, it Is the simplest mathematical expression that adequately

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34 HEALTH RISKS OF RADON AND OTHER ALPHA-EMITTERS describes the level and temporal pattern of risk in the four cohorts. This section provides a summary of the committee's approach and results. Detailed discussion of our statistical models and methods and of their application to these cohort data is given in Annex 2A. Undoubtedly, many factors influence the occurrence of lung can- cer in miners exposed to radon daughters. In carrying out its own original analyses of the data on the cohorts, the committee focused on the following potential risk factors: cumulative exposure, dura- tion of exposure, age at which risk is being evaluated, age at first exposure, time since cessation of exposure, and time since each part of the exposure. In the one cohort for which it was possible to do so (the Colorado Plateau uranium miners), the ejects of smoking were also evaluated. The original investigators of these data relied primarily on calculation of standardized mortality ratios (SMRs) by exposure category, a method that provides a useful but limited anal- ysis of the data. In addition to a thorough investigation of effects of the above risk factors, a substantial part of the analysis was in terms of comparisons purely within the cohort data, as opposed to comparison with data on external populations, as in analyses based on SMRs. STATISTICAL MODELS AND METHODS The committee's general approach was to examine how the age- specific relative risk depends on the variables of interest. This was done by making a cross-classification of numbers of Jung-cancer cleaths and person-years at risk, by categories of these variables, and then fitting models to the rates given by the ratio of deaths to person-years in such a tabular cross-classification. The committee fitted regression models with a Poisson probability mode] for the number of deaths in each cell of the table, where the expected value was taken as the product of the person-years at risk for the cell and a cancer rate given by a parametric model. For the case of purely internal cohort comparisons, not relying on external rates, this is a grouped-data analog of the widely used Cox relative-risk regression method.2 For the case of comparison with external rates, it is a gen- eralization of standard SMR methods that provides more detailed examination of the relative risk. (A very useful reference for these methods is the survey paper by Bresiow.~) The parametric models for this analysis were expressed in terms of the excess relative risk, that is, the ratio of the excess risk to the

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148 HEALTH RISKS OF RADON AND OTHER ALPHA-EMITTERS 1.2 e 1.0 15~ 0.8 Oh o ,¢ 0.6 Cal z O 0.4 Or m z 0.2 O . _ · J-E Model O J-B Model 0 H-P Model T J-B 10 15 20 BREATHING RATE l/min FIGURE 2B-4 Mean bronchial dose as a function of minute volume.26 30 generations by as much as 50%, so that the total deposition in- creases less rapidly than the increased alpha energy inhaled. Re- cently, James27 has reported that the total deposition is proportional to the square root of the minute volume over the range of particle sizes in which deposition is due to particle diffusion. The variation in the mean bronchial dose as a function of minute volume for the three models is shown in Figure 2~4 for particles with an AMAD of 0.15 ,um. Note that the curves are concave downward because of the square root variation of dose with minute volume. In summary, the results of efforts to mode! the dose to the Jung show significant differences in the estimates of the dose per unit of exposure in the tracheobronchial region of the lung: factors of 2 to 3 are typical. Unfortunately, there are few experimental data with which these estimates can be compared, so it Is not possible to make a fully informed choice among the models on the basis of measure- ments. Recently, Cohen4 has reported on measurements leading to correction factors of about 2 in an updated H-P five-Iobe mode.

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RADON DOSIMETRY 149 Furthermore, some recent experiments indicate that the dose at bi- furcations is not uniform and may be considerably larger than those calculated with current models.3i These observations have not, how- ever, been confirmed by Cohen's studies.4 Although the deposition pattern predicted by any particular mode! cannot be validated in viva, the average deposition patterns (and hence, the average dose) for all the models are sufficiently s~rnilar so that any of them can be used with about equal confidence. Even though these three lung models give different values for any particular case, the dose estimates are converging as more infor- mation becomes available. A survey of dose calculations for miners made in the period 1951-1981 by the National Council on Radiation Protection and Measurements (NCRP)34 included 28 calculations of the dose per W[M to the bronchial tissue. The results ranged from 0.7 to 140 mGy/W[M (0.07 to 14 rad/W[M). In addition to differences in the loci of the bronchi for which the dose was calcu- lated, differences in input values of the parameters describing the radon daughters in the air and of the parameters characterizing the dosimetry models were responsible for much of the spread. Jacobi22 estimated that input of the same aerosol characteristics into these models would reduce the spread to about 3 to 10 mGy/W[M (0.3 to 1 rad/W[M). Agreement is even better when the average dose to the trachealobronchial tree is calculated. For the most advanced calculations models, the variation in mean doses can be as little as 20~o.35 However, the degree of agreement is better for some aerosol distributions then others, being best for air in mines and poorer for the cleaner air in homes where the unattached fraction may be higher. EXTRAPOLATION OF DOSES FROM MINES TO HOMES The committee's estimates of lung-cancer risks due to the inhala- tion of radon and its progeny are based on data derived from coning populations. These estimates can be used to calculate the risk per W[M in nonoccupational situations, but the attendant assumptions and uncertainties associated with this change in exposure conditions must be considered. There are potentially important differences be- tween the environmental conditions in which exposure is sustained in a mine and a home and between the physical characteristics of miners and members of the general population. Just as there is no single lung mode} that has been shown to be best in all cases of radon

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150 HEALTH RISKS OF RADON AND OTHER ALPHA-EMITTERS exposure, the committee finds that there is no single characterization of the environment in mines or homes that can be called typical. For the same working-level exposure, there can be a wide variation of particle size, equilibrium factors, and percent unattached fraction. In addition, a miner working underground! and a person living in a home have different levels of activity and, thus, different breathing patterns. Rather than postulating some average or typical values for the various factors that affect the dosimetry of radon progeny in mines and homes and calculating appropriate dose per WI,M conversion factors, the committee proposes a methodology that can be applied when any radon environment is sufficiently characterized. By using the ratio of the dose per W[M in nones to that in homes obtained by means of the lung models described above, it is possible to extrapolate the risk values obtained for underground miners to people living in homes, at a reasonable level of approximation. Let K be a dimensionless factor that, when multiplied by the risk to miners per W[M, will give the risk to an individual in a home per W[M. (Ri~k~m (Rinks K = (WLM)m (WLM)h (2B-2) While miners exposed to 1 WL receive 1 W[M during a 17~h working month, a person living in a home at 1 WL would normally be exposed to more than 1 W[M in a month. If an occupancy factor (O.F.) is defined as the fraction of a 72~h month that is spent in a house, then 1 WI' will result in 720/170 O.F. W[M per month of occupancy. Since risk is proportional to dose, K varies as the ratio of the dose per W[M in houses and mines. a, Riskh/WLMh Riskm/WLMm ~ ~ Dosem/wLMm Doseh/WLMh (2~3) As described above, the dose to the lung depends upon both aerosol and physiological factors. Therefore, to a first approximation, the dimensionless proportionality constant K can be expressed in terms of the partial dose conversion factors obtained from a specified lung model. K Oc Ph fh Fh (MV)h . . ./WLMh Pm fm Em (MV)m · · ./WLMm (2~4)

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RADON DOSIMETRY 151 where P is dose conversion factor for particle size, f is dose conver- sion factor for the unattached fraction, F is dose conversion factor for the equilibrium factor, MV is dose conversion factor for the minute volume; all these parameters have units of red per W[M. Physiolog- ical factors such as mucus thickness, mucus transport rate, bronchial morphometry, depth of the target cells, and the particle deposition fraction can be assumed to be approximately equal in miners and persons living in homes. These factors undoubtedly vary from per- son to person; but except for the possibility of Correlated damage to the lung, they should be about equal in miners and others, and there- fore would largely cancel out in Equation 2~4. Differences between smokers and nonsmokers, however, may be important (see Part 3 of Appendix VIl). The committee considered the possibility of systematic di~er- ences between miners and others, for example, prevalence of ~nhala- tion via the mouth, but found that data were not available for evalu- ation. However, sex and age are factors to be considered. Although most calculations of aerosol deposition are based on the anatomy of the male, the female airway geometry is similar, and scaling factors can be used for estimating the dimensions of individual airways.34 In the case of the growing child, the alveolar area is not fully developed at birth. However, the ciliated airways are complete. Calculated deposition patternsi536 indicate higher relative deposition in the first few generations in the lungs of children for aerosols of 0.2-,um diameter or less. For adults of working age and the same smoking status, it is likely that mucociliary clearance rates are similar in miners and in the general population. However, there is little information concerning clearance in other groups in the population and none for children. It is known that clearance in the elderly is delayed.~° Most lung models show an increase of bronchial dose per unit exposure until about age 6, after which it falls off and becomes nearly constant after age 10. The NEA Organisation for Economic Co-operation and Development has recommended that age dependency of dose equivalent per unit exposure be neglected.35 It is instructive to examine the variations of the proportionality factor K, given by Equation 2B-4, under the various dosimetry mod- els, as illustrated by the results shown in Figures 2B-2, 2B-3, and 2~4. In doing so, it will be shown that for a range of conditions typical in mines and homes, under a given model, K is reasonably

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152 HEALTH RISKS OF RADON AND OTHER ALPHA-EMITTERS stable, even though the variation between models in the calculated dose per W[M varies by a factor of 2 or more. The first three terms in Equation 2B-4, the conversion factors for particle size, unattached fraction, and equilibrium factor, are rep- resentative of the physical environment in the mine or home and are, to some extent, interconnected. For example, with increasing aerosol concentration the dose per W[M conversion factor for equilibrium (F) generally increases somewhat, while that for the unattached frac- tion (0 decreases. However, these factors are independent to a first order and K can be approximated by the product of the ratios given in Equation 2B-4. George and colleagues.8 9 made particle size measurements in homes in New York and New Jersey and in mines in Canada and Colorado. They found that particle size was log normally distributed with an activity mean diameter of 0.12 Am in homes and 0.17 Am in the mines. From Figure 2B-3 it is seen that all three lung models give about 1 rad/W~M for 0.12-pm particles compared to approximately 0.7 rad/W[M for 0.17-,um particles. For these particle distributions Ph/Pm is about 1.4. The amount of activity that is unattached depends on the am- bient environment. Because of low ventilation, the mass concen- tration of airborne dust was probably high in early mines and the unattached fraction low. More recently, the ventilation in mines has been greatly increased. However, the number concentration of air- borne dust particles may have become higher due to the widespread use of underground diesel equipment. In a careful study of several uranium mines, George et al.9 found that the unattached fraction ranged from 0.004 to 0.16, with a mean value of 0.04. Recent investigations have provided some preliminary data on the fractions of unattached RaA and the size distribution of the car- rier aerosol for the attached fraction in the home environment.8 The number and size of particles in the air of a home vary considerably throughout the day, depending, among other factors, on the pres- ence of cigarette smoke and the cooking of food. This, in turn, has a large effect on how many ions become attached to particulates. Particulate concentrations are generally lower at night but rarely fall below 104 particles/mI. In the absence of specific aerosol sources in the home, Reineking et al.38 calculated, on the basis of measured concentrations of particulates, that the unattached fraction ranged between 0.06 and 0.15 (mean value, 0.1~. With additional aerosol sources, the unattached fraction fell below 0.05. This value for the

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RADON DOSIMETRY 153 unattached fraction (0.1) ~ somewhat higher than those previously assumed,~9 23 and is higher than the mean value (0.04) measured in uranium mines.7 9 Although the unattached fraction of RaA may be perhaps twice as high in homes compared to mines, extensive data on homes and mines have yet to be published. Assuming that the mines used to obtain the risk numbers for radon had an unattached fraction (f m) of about 4%, the three models shown in Figure 2B-4 give a conversion factor to the bronchial region of about 0.5 0.6 rad/W~M. For a diffusion coefficient of 0.005 cm2/s and an unattached fraction of To in homes, the same curves give conversion factors between 0.5 and 0.7 rad/W[M. The ratio fh/lm would then be about 1.2 (1.~1.4), depending on which lung model is chosen. - The equilibrium factor F is dependent on the ventilation rate. High ventilation rates result In low values of F. Jacobi and Eisfeld24 have shown that variation of the equilibrium factor has very little effect on the dose per unit exposure to the bronchial cells. Therefore, even though Fh and Em may be different, the ratio of the conversion factors is about I. Physiological factors must be considered also. Underground min- ers, usually adult males, spend about 40 h weekly engaged in mod- erate to heavy activity ~ the course of their work. Nonoccupational exposures of adults, both male and female, and children generally occur in the home during light activity or while asleep. So, in addi- tion to differing aerosols in homes and mines, the different breathing patterns of miners and the population in homes must be considered. The increased minute volume required for the metabolic cost of exercise is achieved by an increase in both the tidal volume and the frequency of breathing. The increased frequency of breathing decreases the mean residence time of aerosols in the lung and, by so doing, reduces the time available for diffusion to deposit particles on the bronchial airways. Exercise also has a role in how people inhale. With increasing ventilation there is a shift from nasal to oral breathing. For example, with exercise requiring ventilation of 35 liters/min, there is a shift from a pattern of 80~o nasal breathing in the resting subjects to about 50~o nasal breathing.37 Moreover, many normal people breath oronasally,39 and those with any form of nasal obstruction have a mainly oral form of breathing. The proportion of oral and nasal breathing influences the bronchial dose since, as noted above, about half of the unattached fraction is assumed to be deposited in the

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154 HEALTH RISKS OF RADON AND OTHER ALPHA-EMITTERS nose. increased oral breathing, therefore, increases bronchial doses, whereas increased nasal breathing reduces this dose. It can be argued that the larger fraction of unattached RaA in the home compared to that in a mine increases the dose to the bronchus for a given concentration of radon progeny in air. However, the increased minute ventilation of the miner is associated with more mouth breathing, which would tend to increase his dose. The committee could not find data on the pattern of nose or mouth breathing in miners and therefore could not examine this factor analytically. It is an obvious problem for further field investigation. ICRP,~9 NEA,35 and NCRP33 assign a breathing rate of 20 liters/m~n to underground miners and base their dose estimates on this minute volume. Ventilation rates in working coal miners have been measure; values obtained ranged from about 30 to 40 liter/min when averaged throughout the working shift. This Is about 1.5 to 2 times the 2~liter/min value assigned to miners and 3 to 4 times the 12.5 liter/m~n value assignee] to the general population in the NEA35 model. Moreover, individuals undertaking the same task have differing minute ventilation. Heavier miners breathe more, as do older miners. The very high ventilation rates found in coal miners might not be representative of values for uranium miners. In the case of the home environment, there are considerable uncertainties in attempting to estimate overall deposition patterns throughout the 24-h cycle. Not only is the distribution of oral/nasal breathing a relevant factor, but the tidal volume and frequency of breathing change continuously. The minute volume of a sleeping person is low, and the convective diffusion component of bronchial deposition may be insignificant.30 For a miner who averages a minute volume of 30 liters/m~n, the conversion factors from the three models shown in Figure 2~4 would be on the order of 0.8 rad/WLM. The average person in a home breathing, 12.5 liters/min would receive a dose of 0.45 rad/WLM. Thus, the conversion factor ratio due to these different minute vol- umes (MVh/MVm) would be 0.56. If the miners do a significantly larger amount of mouth breathing than occurs in homes, this ratio would be smaller. For the mean bronchial dose, differences between the models are small. The ratio of the conversion factors due to the different minute volumes would remain nearly the same for any of the three models because the shapes of the curves in Figure 2~4 are very similar. However, for basal cells at 22-,um depths as the target

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RADON DOSIMETRY 155 cells, the H-P mode] gives somewhat higher doses per unit exposure than do the other two models, and the variation in the ratio of the factor, MVh/MVm, in Equation 2B~2 due to the choice of the mode} becomes larger. As outlined in Equation 2~4, the four conversion factors making up K have ratios in homes to mines of 1.4, 1.2, 1.0, and 0.56. Therefore, for this specific case, the product of these factors is 0.94 and the mean dose received by a person exposed in a home to 1 WL of radon for 170 h is very nearly the same as the dose to a miner exposed to 1 WHIM. On the other hand, if the minute volume for miners is taken as 20 liters/min, the ratio K is increased to about 1.3. In homes where the unattached fraction is low (Rio), this ratio is less, about 0.8. Other investigators have reached similar conclusions. After reviewing the parameters relevant to a comparison of the WHIM in a mine and in the home, the NCRP concluded that the alpha dose to the surface of the bronchial tree (red per W[M) may be somewhat higher for environmental exposure than for underground exposure, largely due to a higher fraction of unattached RaA in the home.34 In contrast, the NEA has estimated smaller doses per WI,M for adult members of the general population than for miners.35 The NEA assumed a smaller unattached fraction (3%) than that assumed by the NCRP (7%o) and a slightly lower breathing rate than that assumed by the NCRP. Both the NCRP and the NEA analyses assume that miners inhale 20 liters/min, that is, light activity. Upon their review of the possible range of the input parameters for the dosimetric models, the NCRP concluded that the dose per WHIM in homes, as compared to that in mines, differs by less than a factor of 2. ~ summary, the committee believes that with the present state of scientific knowledge, it can neither choose between the three major radon lung models nor specify the best values of the factors which characterize the dose in a home or in a mine. However, all of the models are in agreement within about a factor of 3 or so. While the method of ratios used here does not directly calculate a dose to the lung tissues, it does allow the extrapolation of known risks in a mine to a home where dosimetry factors have been established. Future work on this methodology should be concentrated in areas that will improve the quantification of the constant K. This will include improving our knowledge of the environment in the mines from which the risk estimates were obtained, characterizing

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156 HEALTH RISKS OF RADON AND OTHER AlPHA-EMITTERS the indoor environment of homes more completely, and determining breathing characteristics of uranium miners by on-site measurements. Finally, high priority should be given to creating and validating a Jung mode} which retains the best features of the modem now in use. REFERENCES 1. Altshuler, N., N. Nelson, and M. Kuschner. 1964. Estimation of lung tissue dose from the inhalation of radon daughters. Health Phys. 10:1137-1161. 2. Camner, P. 1981. Influence of nose and mouth breathing on particle deposition in the lung. Health Phys. 40:99-100. 3. Chamberlain, A. C., and E. D. Dyson. 1956. The dose to the trachea and bronchi from the decay products of radon and thoron. Br. J. Radio. 29:317-325. 4. Cohen, B. S. In press. Deposition of ultrafine particles in human tra- cheobronchial tree. In Radon and Its Decay Products: Its Occurrence Properties and Health Effects, P. K. Hopke, ed. Symposium Series 331. Washington, D.C.: American Chemical Society. 5. Davies, C. N. 1974. Size distribution of atmospheric particles. Aerol. Sci. 5:293-300. 6. E`ry, R. M. 1970. Radon and its hazards. Pp. 13-32 in Personal Dosimetry and Area Monitoring Suitable for Radon and Daughter Products. Paris: Nuclear Energy Agency, Organization for Economic Co-operation and Development. 7. George, A. C., and A. J. Breslin. 1969. Deposition of radon daughters in humans exposed to uranium mine atmospheres. Health Phys. 17:115-124. 8. George, A. C., and A. J. Breslin. 1975. The distributions of ambient radon and radon daughters in residential buildings in the New Jersey-New York area. In The Natural Radiation Environment II, T. F. Gesell and W. M. Lowden, eds. Washington, D.C.: Technical Information Center, U.S. Department of Energy. 9. George, A. C., L. Hinchcliffe, and R. Sladow~ki. 1975. Size distribution of radon daughter particles in uranium mine atmospheres. Am. Ind. Hyg. Assoc. J. 34:484-490. 10. Goodman, R. M., B. M. Yergin, J. K. Landa, M. H. Golinuaux, and M. A. Sacker. 1978. Relationship of smoking history and pulmonary function tests to tracheal mucous velocity in nonsmokers, young smokers, ax-smokers and patients with chronic bronchitis. Am. Rev. Respir. Dis. 117:205-214. 11. Hadden G. G., C. O. Jones, and D. C. Morgan. 1967. A study of volumes of refined air in relation to dust exposures of coal miners. Pp. 37-48 in Inhaled Particles and Vapours II, C. N. Davies, ed. Elm~ford, N.Y.: Pergamon. 12. Harley, N. H. 1984. Comparing radon daughter dose: Environmental versus underground exposure. Rad. Protect. Dosimetry 7:371-375. 13. Harley, N. H., and B. S. Cohen. In press. Updating radon daughter bronchial dosimetry. In Radon and Its Decay Products: Its Occurrence Properties and Health Effects, P. K. Hopke, ed. Symposium Series 331. Washington, D.C.: American Chemical Society. 14. Harley N. H., and B. S. Pasternack. 1982. Environmental radon daughter alpha factors in five-lobed human lung. Health Phys 42:789-799.

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RADON DOSIMETRY 157 15. Hislop, A., D. C. F. Muir, M. Jacobsen, G. Simon, and L. Reid. 1972. Postnatal growth and function of the pre-acinar airways. Thorax 27:265- 274. 16. Holaday, D. A., D. E. Rushing, R. D. Coleman, P. F. Woolrich, H. L. Kusnetz, and W. F. Bale. 1957. Control of Radon and Daughters in Uranium Mines and Calculations of Biological EEects. U.S. Pulic Health Service Publication No. 494. Washington, D.C.: U.S. Government Printing Office. 17. Horsfield, K., F. G. Relea, and G. Cumming. 1976. Diameter, length and branching ratios in the bronchial tree. Respir. Physiol. 26:351-356. 18. International Commission on Radiological Protection (ICRP). 1975. Re- port of the Task Group on Reference Man. ICRP Publication 23. Oxford: Pergamon. International Commission on Radiological Protection (ICRP). 1981. Limits for the Inhalation of Radon Daughters by Workers. ICRP Publication 32. Oxford: Pergamon. 20. Jackson, P. O., J. A. Cooper, J. C. Langford, and M. R. Peterson. 1982. Characteristics of attached radon-222 daughters under both laboratory and underground uranium mine environments. Pp. 1031-1042 in Radiation Hazards in Mining: Control, Measurement, and Medical Aspects, M. Gomez, ed. New York: Society for Mining Engineers of the American Institute of Mining, Metallurgical, and Petroleum Engineers, Inc. 21. Jacobi, W. 1964. The dose to the human respiratory tract bv inhalation of ~ ^~^ _ , ,, _ ~ .. _,, ~ short-lived Mourn- and ^~VRn-decay products. Health Phys. 10:1163-1174. 22. Jacobi, W. 1977. Interpretation of measurements in uranium mines: Dose evaluation and biomedical aspects. Pp. 33-48. in Personal Dosimetry and Area Monitoring Suitable for Radon and Daughter Products. Paris: Nuclear Energy Agency, Organisation for Economic Co-operation and Development. 23. Jacobi, W., and K. Eisfeld. 1980. Dose to Tissue and Effective Dose Equiv- alent by Inhalations of Radon-222 and Their Short-Lived Daughters. GSF Report S-626. Neurherberg, Federal Republic of Germany: Geschelscheft fur Strahlen and Umwelfforschung. 24. Jacobi, W., and K. Eisfeld. Internal Dosimetry of Radon-222, Radon-220 and their Short-Lived Daughters. Pp. 131-143 in Proc. Second Special Symposium on Natural Radiations Environment, K. G. Vohra et al., eds., January 1981, Bhabha Atomic Research Centre, Bombay. New York: John Wiley & Sons. 25. James, A. C. 1977. Bronchial deposition of free ions and submicron particle studies in excised lung. Pp. 203-218 in Inhaled Particles IV, W. H. Walton, ed. New York: Pergamon. 26. James, A. C. 1986. Dosimetric approaches to risk assessment for indoor exposure to radon daughters. Radiat. Prot. Dosimetry 7:353-366. 27. James, A. C. In press. A reconsideration of cells at risk and other key factors in radon daughter dosimetry. In Radon and Its Decay Products: Its Occurrence Properties and Health Effects, P. K. Hopke, ed. Symposium Series 331. Washington, D.C.: American Chemical Society. 28. James, A. C., J. R. Greenhalgh, and A. Birchall. 1980. A domestic model for tissues of the human respiratory tract at risk from inhaled radon and thoron daughters. Pp. 1045-1048 in Radiation Protection. A Systematic Approach to Safety, Proceedings of the 5th Congress of International

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158 HEALTH RISKS OF RADON AND OTHER ALPHA-EMITTERS Radiation Protection Association (IRPA), Vol. 2, Jerusalem. New York: Pergamon. 29. Kraner, H. W., G. L. Schroder, and R. D. Evans. 1964. Measurements of the effects of atmospheric variables on radon-222 flux and soil-gas concentrations. Pp. 191-215 in The Natural Radiation Environment, J. A. S. Adams and W. M. Lowder, eds. Chicago: University of Chicago Press. 30. Martin, D., and W. Jacobi. 1972. Diffusion deposition of small-sized particles in the bronchial tree. Health Phys. 23:23-29. 31. Martonen, T. B., W. Hofman, and J. E. Lowe. 1987. Cigarette smoke and lung cancer. Health Phys. 52 :213-217. 32. National Council on Radiation Protection and Measurements (NCRP). 1975. Natural Background Radiation in the United States. NCRP Report 45. Washington, D.C.: National Council on Radiation Protection and Measurements. 33. National Council on Radiation Protection and Measurements. (NCRP). 1984. Exposures from the Uranium Series with Emphasis on Radon and Its Daughters. NCRP Report 77. Washington, D.C.: National Council on Radiation Protection and Measurements. 34. National Council on Radiation Protection and Measurements (NCRP). 1984. Evaluation of Occupational and Environmental Exposures to Radon and Radon Daughters in the United States. NCRP Report 78. Washington, D.C.: National Council on Radiation Protection and Measurements. 35. Nuclear Energy Agency (NEA), Group of Experts of the Organisation for Economic Co-operation and Development. 1983. Dosimetry Aspects of Exposure to Radon and Thoron Daughter Products. Paris: Nuclear Energy Agency, Organisation for Economic Co-operation and Development. 36. Phalen, R. F., M. J. Oldham, M. T. Kleinaman, and T. T. Cracker. In 37. 38. 39. 40. 41. 42. press. Tracheobronchial deposition predictions for infants, children and adolescents. Anat. Respir. Proctor, D. F. 1981. Oronasal breathing and studies of effects of air pollutants on the lung. Am. Rev. Respir. Dis. 123:242. Reineking, A., K. H. Becker, and J. Porstendorfer. 1985. Measurements of unattached fractions of radon daughters in houses. Sci. Total Environ. 45 :261-270. Sheppard, D., J. A. Nadel, and H. A. Boushey. 1981. Oronasal breathing and studies of effects of air pollutants on the lung (correspondence). Am. Rev. Respir. Dis. 123:242-243. United Nations Scientific Committee on the Effects of Atomic Radiation (UNSLEAR). 1977. Sources and Effects of Ionizing Radiation. Report E.77.IX.1. New York: United Nations, 1977. Walsh, P. J. 1970. Radiation dose to the respiratory tract of uranium miners a review of the literature. Environ. Res. 3:14-36. Walsh, P. J., and P. E. Hamrick. 1977. Radioactive materials determi- nants of dose to the respiratory tract. Pp. 233-242 in Handbook of Physiology, D. H. K. Lee, ed. Bethesda, Md.: American Physiological Society. 43. Weibel, E. R. 1963. Morphometry of the Human Lung. New York: Academic Press. 44. Yeh, H. C., and G. M. Schum. 1980. Models of human lung airways and their application to inhaled particle deposition. Bull. Math. Biol. 42 :462-480.

Representative terms from entire chapter:

radon daughters