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Health Effects of Exposure to Radon: BEIR VI Appendix B Comparative Dosimetry INTRODUCTION In order to extrapolate the lung-cancer risk derived from the epidemiological analysis of the underground miner data to the general population, a number of steps are needed. The approach to dealing with smoking, gender, and age at exposure were discussed in chapter 3. In addition, it is necessary to account for the differences in exposure conditions between mines and homes, in breathing rates for different activities in mining and typical home behavior, and in respiratory physiology between men, women, children, and infants. These factors determine the relationships between exposure and dose and the possibility of differing relationships between exposure and dose for miners and for the general population needs to be considered in extrapolating a risk model from miners to the general population. The comparative dosimetry approach used by the committee follows that described by the Panel on Dosimetric Assumptions Affecting the Application of Radon Risk Estimates (NRC 1991) and the BEIR IV Committee (NRC 1988). In chapters 2 and 3, the committee reviewed the biological and epidemiological data suggesting that lung-cancer risk varies directly with exposure to radon and its decay products It is now assumed that lung-cancer risks varies directly with the dose of alpha energy delivered to the appropriate cellular targets, and that the dose can be estimated from the exposure using a dosimetric model. A dosimetric model is employed to estimate the dose received by particular classes within the general population such as adult males (not miners), adult women, children, and infants as well as to adult male miners. The dosimetric
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Health Effects of Exposure to Radon: BEIR VI model takes into account the exposure conditions in homes and in mines as well as the relevant physiological characteristics of the population groups. The ratio of the dose of alpha energy per unit exposure for a particular population group (men, women, children, infant) as given by the radon concentration to the dose per unit radon concentration to the miners is given by K: K = [Dosehome/Exposurehome]/[Dosemine/Exposuremine] (1) The K-factor includes diverse environmental and physiological factors and the use of this double ratio greatly simplifies the risk assessment for indoor radon. This chapter addresses dosimetry of radon progeny in the lung and presents the calculated K-factor values by reviewing the information available on exposure conditions in homes and mines, presenting the dosimetric model used and then presenting the resulting distributions of K-factor values that were calculated. EXPOSURE Introduction The formation and decay sequence for 222Rn was shown in Figure 1-1. Because 222Rn has an almost 4-day half-life, it has time to penetrate through the soil and building materials into the indoor environment where it decays into its progeny. There is some recent evidence that in spite of its short half-life, 55 seconds, 220Rn can also penetrate into structures in significant amounts. However, the data are limited and the extent of the thoron problem is quite uncertain as discussed in a subsequent section of this chapter. The short-lived decay products, 218Po (Radium-A), 214Pb (Radium-B), 214Bi (Radium-C), and 214Po (Radium-C'), represent a rapid sequence of decays that result in two α-decays, two β-decays and several γ-emissions following the decay. To illustrate the behavior of the activity of the radioactive products of the radon decay, the activity of each of the short-lived isotopes is plotted as a function of time for initially pure 222Rn in Figure B-1. Because 222Rn has a longer half-life than either of the four short-lived products, the progeny reach the same activity (number of decays per unit time) as the radon. The mixture then decays with the 3.8 day half-life of the radon. Each 222Rn decay results in four progeny decays so that the total activity is then the sum of these individual decay-product concentrations. The activity is the product of the decay constant (In 2/half-life) times the number of radioactive atoms. Thus, the short-lived 218Po can have an activity equal to the 222Rn because a large decay constant times a small number of atoms becomes equal to a small decay constant times a large number of atoms. If the products formed by the decay of the radon were to remain in the air, then there would also be equal activity concentrations of 218Po, 214Pb, and other progeny. The resulting mixture is said to be in secular equilibrium. As used in the monitoring of uranium mines, an equilibrium mixture of these decay products
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Health Effects of Exposure to Radon: BEIR VI FIGURE B-1 Normalized in-growth of decay-product activities of an atmosphere initially containing only 222Rn. at 3700 Bqm-3 (100 pCiL-1) is called a working level (WL). Thus, a 370 Bqm-3 (10 pCiL-1) equilibrium mixture represents 0.1 working level. The cumulative exposure to such activity can be expressed as the amount of activity in WL multiplied by time of exposure. In occupational exposure assessments for miners, this cumulative exposure has historically been given in Working Level Months, WLM, where it is assumed there are 170 hours in a working month. The WLM is calculated as where (WL)i is the average concentration of radon decay products during exposure interval expressed in WL and ti is the number of hours of the exposure. The cumulative exposure when spending all of the time in a house at a given decay product concentration is more than four times that for occupational exposure (8766 compared to 2000 hours worked on an annual basis).
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Health Effects of Exposure to Radon: BEIR VI More appropriately, the activity of the radon decay products is described by a quantity called the Potential Alpha Energy Concentration (PAEC). The total airborne potential alpha energy concentration, PAEC, is calculated as follows: Cp(Jm-3) = (5.79C1 + 28.6C2 + 21.0C3) × 10-10 (3) where Cp is the PAEC in Jm-3, and C1, C2, C3 are the activity concentrations of 218Po, 214Pb, and 214Bi in Bqm-3, respectively. This quantity incorporates the deposition of energy into the air. Exposure can be then expressed as the PAEC multiplied by the length of exposure in hours. This exposure is reported in the scientific units of Joule-hours per cubic meter (Jhm-3). Although much of the prior epidemiological studies on the radon risk as observed in underground miner populations have used exposure measured in WLM, these values can be easily converted into proper SI units by multiplying the original estimates of exposure in WLM by 3.5 × 10-3 Jhm-3 WLM-1. In either homes or mines, decay products are lost from the air by attachment to environmental surfaces. In homes these surfaces include walls, floors, furniture, and the people in the room. The decay products also attach to airborne particles. The processes that control the airborne concentrations of the decay products are shown in Figure B-2. The attachment to the airborne particles maintains the radioactivity in the air since submicron particles typically remain suspended in the air over long times. The ratio of decay products to radon, termed the equilibrium factor, F, ranges from 0.2 to 0.8 with typical values of 0.35 to 0.4 (Hopke and others 1995). FIGURE B-2 Behavior of radon decay products in indoor air.
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Health Effects of Exposure to Radon: BEIR VI Another concept relating to the decay products is that of the ''unattached" fraction. Although it is now known that the decay-product atoms are ultrafine particles (0.5 to 2 nm in diameter), an operationally defined quantity called the "unattached" fraction has been used to describe this activity. However, there is no general agreement on a procedure for estimating its value. The "unattached" fraction should not be considered as single atoms of progeny; we now know that the unattached fraction actually comprises small clusters of molecules. These unattached decay products have much higher mobilities in the air than the attached activity and can more effectively deposit in the respiratory tract. Thus, the "unattached" fraction has been given emphasis in estimating the health effects of radon decay products. Typically most of the "unattached" activity is 218Po and the value of unattached fraction, fp, is usually in the range of 0.01 to 0.10 in indoor air, although it may be higher if the concentrations of particles in the air are very low. One does not actually measure unattached particle numbers but rather unattached decay-product activity in Bqm-3. Thus, the PAEC can be obtained directly from the unattached activity. AIRBORNE PARTICLE PROPERTIES The particles in indoor air are quite different in size from those that are typically encountered in the outdoor atmosphere. There are also many types of particle sources in a home such as open flames of a gas stove or candles, cooking, aerosol-spray products, and tobacco-smoking. In general, particles larger than 1 µm cannot penetrate into a house. A typical distribution of room air particles is shown in Figure B-3. The points are the average of 10 separate measurements. Particle sizes in typical room air are around 50 to 100 nanometers (nm). The size distributions for particles generated by gas stoves and by cigarette-smoking are shown in Figures B-4 and B-5. These distributions show that there are much higher particle concentrations when particle sources are operating. The high temperatures in the gas stove flame produces small particles with a peak in the distribution at about 20 nanometers while the lower temperatures in cigarette burning generates particles in the 100 to 300 nanometer (0.1 to 0.3 µm) range. The particle concentrations decrease over time because of surface deposition and ventilation. To obtain estimates of the activity-weighted size distributions, the number distributions are multiplied by the probability of attachment for the unattached radon decay product (Porstendörfer and others 1979). Exposure in Normally Occupied Homes Over the past decade, improvements in measurement technology have allowed the direct measurement of the activity-weighted size distribution of the radon decay products in normally occupied homes. Methods have been developed by which the entire radioactive aerosol size distribution can be
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Health Effects of Exposure to Radon: BEIR VI FIGURE B-3 Particle size distribution for room air. The curve is a fit to averages (points not shown). Each average derived from 10 separate measurements. Figure taken from Li and Hopke (1993). deduced from the measured activity collected on or penetrating through a series of screens. The first activity-size measurements in indoor and ambient air were made by Sinclair and others (1977) using a specially designed high-volume-flow diffusion battery. They observed bimodal distributions with activity mode diameters of 7.5 and 150 nm in indoor atmospheres and 30 and 500 nm outdoors in New York City. Similar results were reported by George and Breslin (1980). In Göttingen, West Germany, Becker and others (1984) only observed the larger mode using a modified impactor method with a minimum detectable size of 10 nm. More extensive measurements in New York City by the group at the Environmental Measurements Laboratory (EML) have been reported by Knutson and others (1984) using several different types of diffusion batteries as well as cascade impactors. They again observed modes around 10 nm and 130 nm in the PAEC-weighted size distribution measured with a low-volume screen diffusion battery. Four samples taken with a medium-volume (25 L min-1) screen diffusion battery showed a major mode at 80 to 110 nm and a minor mode containing 8 to 9% of
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Health Effects of Exposure to Radon: BEIR VI FIGURE B-4 Particle size distribution from a gas stove burner. Each point represents the average of 10 separate measurements. Figure taken from Li and Hopke (1993). the PAEC with a diameter < 5 nm. Finally, the same group made measurements at Socorro, NM (George and others 1984). They report that the major mode was only slightly different from that found in New York. However, the minor mode was always < 5 nm, distinctly smaller than in the New York distributions. One of the problems with the extension of screen diffusion batteries to smaller particle sizes is the substantial collection efficiency of the high mesh-number screens typically used in diffusion batteries designed to cover the range of particle size from 5 to 500 nm. At normally used flow rates, a single 635-mesh screen has greater than 90% efficiency for collecting 1 nm particles, the size of "unattached" 218Po having a diffusion coefficient of the order of 0.05 cm2s-1. A theory of screen penetration (Cheng and Yeh 1980; Cheng and others 1980; Yeh and others 1982) had been developed in the late 1970s and this theory was validated to 4 nm by Scheibel and Porstendörfer (1984). The limitations of high mesh-number screens were recognized and it then became possible to examine new types of diffusion battery designs that could be extended to smaller particle diameters. Reineking and others (1985), Reineking and Porstendörfer (1986) and Reineking and others (1988) use the high-volume flow diffusion batteries de-
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Health Effects of Exposure to Radon: BEIR VI FIGURE B-5 Particle size distribution for sidestream cigarette smoke. Each point represents the average of 10 separate measurements. Figure taken from Li and Hopke (1993). scribed in Reineking and Porstendörfer (1986) to obtain activity-size distributions. They obtained their size distributions by fitting lognormal distributions using an algorithm that searches the solution space for an acceptable fit (SIMPLEX). Size distributions of indoor air in rooms without and with the addition of particles from running an electric motor are presented in Figures B-6 and B-7. Holub and Knutson (1987) report the development of low-flow diffusion batteries with low mesh number screens and extension of the EML batteries to smaller sizes. Tu and Knutson (1988a,b) have used the 25 L min-1 screen diffusion batteries to measure the 218Po-weighted size distributions in the presence of several specific aerosol sources. The results of these measurements are presented in Figures B-8 and B-9. The presence of a mode around 10 nm is again observed in curve 1 in Figure B-8. Only in curve 1 (no active aerosol sources) in Figure B-9 is a mode at 1 nm observed. In all of the other cases, the activity is attached to the aerosol present in the house. The attachment was confirmed by independently measuring the aerosol-size distributions using an electrical aerosol analyzer (Liu and Pui 1975) and the attachment coefficients recommended by Porstendörfer and others (1979). There was a high level of agreement between the measured and calculated activity-weighted size distributions.
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Health Effects of Exposure to Radon: BEIR VI FIGURE B-6 Activity-weighted size distribution of the indoor aerosol in a closed room without active aerosol sources. Figure taken from Reineking and Portstendörfer (1986). FIGURE B-7 Activity-weighted size distribution of the indoor aerosol in a closed room with an operating electric motor. Figure taken from Reineking and Porstendörfer (1986).
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Health Effects of Exposure to Radon: BEIR VI FIGURE B-8 218Po-weighted size distributions measured in house I. 1. Cooking (5 min.); 2. Frying food; 3. Cooking soap; 4. Cigarette smoldering. Figure taken from Tu and Knutson (1988a). FIGURE B-9 218Po-weighted size distributions measured in house II after a kerosene heater was operated for: 1.0 min; 2.80 min; 3.20 min. Figure taken from Tu and Knutson (1988a).
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Health Effects of Exposure to Radon: BEIR VI Several other groups including the National Radiation Protection Board (NRPB) of the United Kingdom and the Australian Radiation Laboratory (ARL) have also developed these graded-screen diffusion batteries for activity-weighted size distribution measurements. Intercomparison experiments between the measurement methods have been performed (Hopke and others 1992). Generally good agreement was obtained. Several automated systems to make use of this methodology have been developed. Strong (1988) used 6 sampling heads containing 0, 1, 3, 7 18, and 45 stainless steel, 400-mesh wire screens. He has measured the size distributions in several rooms in two houses at two times of the year. The size distributions observed in the kitchen are presented in Figure B-10. These results are summarized in Table B-1. It should be noted that in the "kitchen" curve in Figure B-10, a trimodal distribution is observed; a true "unattached" fraction at 1 nm, a nuclei mode at 10 nm, and an accumulation mode at 100 to 130 nm. In the table, the "unattached" fractions presented are the integrated values from the size distributions. A problem of interpretation then arises regarding the "unattached'' fraction since Strong integrates the distribution up to > 10 nm to obtain the fraction that he attributes as being "unattached." For the distribution in a kitchen during cooking, the activity median diameter for the "unattached" fraction is given as 11 nm. This designation is a clear departure from the original purpose for defining an "unattached" fraction. The advent of these more sophisticated and sensitive measurement systems has necessitated either a more precise definition of the meaning of the "unattached" fraction or cessation of its use (Hopke 1992). Subsequent to these original measurements, Strong (1989) modified his system by changing the screens to one 200-mesh screen and 1, 4, 14, and 45 400- FIGURE B-10 PAEC size distributions measured in a rural house kitchen by Strong (1988).
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Health Effects of Exposure to Radon: BEIR VI TABLE B-9 Relative distribution of time in various activities for the time spent indoors at home (taken from ICRP 1994 and LUDEP default settings from Jarvis and others 1993) Population Subgroup Sleeping Sitting Light Activity Heavy Activity Infant, 1 yr 58.33% 12.50% 25.00% — Child, 10 yr 41.67% 12.28% 30.56% — Female adult 55.00% 15.00% 30.00% — Male adult 55.00% 15.00% 30.00% — Miner adult 0.00% 31.30% 68.80% — TABLE B-10 Morphometric, physiologic, and lung volumes for children and adults (taken from ICRP 1994) Infant, 1 yr Child, 10 yr Adult Male Adult Female Height (cm) 75 138 176 163 Weight (kg) 10 33 73 60 TLCa (mL) Mean 548 2869 6982 4968 Standard Deviation 61.5 — 700 600 FRCb (mL) Mean 244 1484 3301 2681 Standard Deviation 26 311 600 500 VCc (mL) Mean 377 2326 5018 3551 Standard Deviation 47 — 560 420 VDd (mL) Mean 20 78 146 124 Standard Deviation — 14.9 22.5 21.0 Anatomical Dead Space (mL) VBBe 6.81 25.10 50.00 40.08 Vbbf 4.70 26.45 47.06 40.19 VETg 8.73 26.49 48.75 44.18 Smallest Bronchiole (cm) 0.1068 0.1429 0.1651 0.1587 Largest Bronchus (cm) 0.7502 1.3114 1.6500 1.5342 a Total lung capacity. b Functional residual capacity: volume of air that remains in the lungs after exhalation. c Vital capacity: maximum volume of air breathed in during inspiration. d Total anatomical dead space: the volume in which no gas exchange occurs. e Bronchial dead space volume. f Bronchiolar dead space volume. g Extrathoracic dead space volume.
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Health Effects of Exposure to Radon: BEIR VI ments) were calculated from activity concentrations associated with the activity-weighted size distributions for the home environment and the dose-conversion coefficient. This physical dose-rate per unit radon concentration is then divided by the estimated average dose to a miner based on the exposure results presented above. The miner dose-rate per unit radon concentration is 7.0 nGy h-1/(Bqm-3). The dose-rates per unit radon are calculated for men, women, children, and infants based on the activity-weighted size distributions measured. Each value of the dose per unit radon concentration in homes is divided by the dose per unit radon concentration in mines to yield a series of K values. The distributions for the K values for houses with a smoker are presented in Figures B-21 to B-24. The distributions for the K values for homes without a smoker are shown in Figures B-25 to B-28. The median values for the K factor for the various classes of individual in the 2 exposure scenarios (smoker and non-smoker) are given in Table B-11. It can be seen that all of these values are near to 1 so that the dose per unit exposure per unit 222Rn concentration are essentially the same in mines and homes. The measurements in homes (Hopke and others 1995) showed that the median dose per unit radon was relatively insensitive to the aerosol conditions in the homes. FIGURE B-21 Cumulative frequency distribution for the K factor for adult males based on the set of measurements in smoker homes.
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Health Effects of Exposure to Radon: BEIR VI FIGURE B-22 Cumulative frequency distribution for the K factor for adult females based on the set of measurements in smoker homes. FIGURE B-23 Cumulative frequency distribution for the K factor for children (10 yrs of age) based on the set of measurements in smoker homes.
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Health Effects of Exposure to Radon: BEIR VI FIGURE B-24 Cumulative frequency distribution for the K factor for 1 year old infants based on the set of measurements in smoker homes. In the case of the homes with a smoker, there is increased radon-progeny activity per unit radon gas concentration as shown by the increased equilibrium factor, but because the activity tends to be carried by larger particles, it is less effective in delivering dose to the respiratory tract. In the case of the non-smoker homes, the number of the larger particles is smaller and the amount of airborne progeny activity per unit radon gas concentration is also less, but because it tends to be carried by smaller particles, the average dose rate per unit activity concentration is higher. The previous NRC analyses of lung dosimetry in mines and homes (NRC 1991) found K factor values around 0.75 for males and females and somewhat higher for children and infants. In that earlier NRC report, a sensitivity analysis was performed to ascertain the effects of changes in input parameters to the dosimetric model on the resulting K values. The dosimetric model used here is very similar to that used in that report and a new sensitivity analysis was not undertaken. In that analysis, it was found that changes in the breathing rate had the largest effects on the dosimetry. Although breathing rate has some effect on the amount and the location of activity deposition in the lungs, the largest effect of changing the breathing rate is the resulting change in the total volume of air inhaled and thus the total amount of PAEC inhaled. The median K factor in the new calculation has increased from previously calculated values to close to 1. This change
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Health Effects of Exposure to Radon: BEIR VI FIGURE B-25 Cumulative frequency distribution for the K factor for adult males based on the set of measurements in non-smoker homes. FIGURE B-26 Cumulative frequency distribution for the K factor for adult females based on the set of measurements in non-smoker homes.
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Health Effects of Exposure to Radon: BEIR VI FIGURE B-27 Cumulative frequency distribution for the K factor for children (10 yrs of age) based on the set of measurements in non-smoker homes. FIGURE B-28 Cumulative frequency distribution for the K factor for 1 year old infants based on the set of measurements in non-smoker homes.
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Health Effects of Exposure to Radon: BEIR VI TABLE B-11 Values of the K Factor Subject Type Home with a Smoker Home without a Smoker Male 0.94 0.95 Female 1.0 1.00 Child 0.99 0.98 Infant 1.08 1.05 primarily reflects a reduction of the miner breathing rate from the 1.8 m3 h-1 to 1.25 m3 h-1 in this present study. A comparison of the inputs to the dosimetric modeling that was used to derive the K factor is presented in Table B-12. As discussed above, the major difference between the 2 results was the change in breathing rates based on Ruzer and others (1995) and the South African Miner Data quoted by ICRP (ICRP 1994). In this analysis, the committee was able to utilize the more extensive set of the measurement in homes by Hopke and others (1995) and the more complete reconstruction of the activity-weighted size distributions in mines provided by Knutson and George (1992) and George (1993). For the earlier study, the growth of particles when respirated into the high humidity of the lungs was estimated to be a factor of 2. In the intervening time, a number of measurements on room air and a variety of materials (Li and Hopke 1993, 1994; Dua and others 1995; Dua and Hopke 1996) have suggested that this growth factor is quite reasonable for typical room air and since much of aerosol in the ventilation air in the mine is ambient aerosol, particularly in the size range below 1 µm where the attached activity is typically observed, this factor of 2 is also quite reasonable for mines as well. THORON There are 2 other naturally occurring isotopes of radon, 220Rn and 219Rn, commonly referred to as thoron and acton. 220Rn is a decay product of 232Th and 219Rn is a decay product of 235U. 219Rn has a half-life of 3.6 s and thus, it is not transported far after its formation to produce measurable exposure even to miners. However, even with a half-life of 55.6 s, 220Rn can be found present in the atmosphere (Schery 1992). Thoron undergoes a-decay through short-lived 216Po (0.15 s) which also undergoes a-decays to 212Pb with a half-life of 10.64 h. The decay chain, including half-lives and the energies of the emitted radiation, is outlined in Table B-13. Homes There have only been limited measurements of 220Rn and its decay products in indoor air. In the measurement of 220Rn there can be substantial variation
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Health Effects of Exposure to Radon: BEIR VI TABLE B-12 Comparison of input parametersa to comparative dose modeling Parameter 1991 Dose Panel BEIR VI Respiratory Physiology ICRP66 Defaultsb ICRP66 Defaults Breathing Rates (Miners) 1.80 m3 h-1 1.25 m3 h-1 Extrathoracic Deposition of Ultrafine Activity ~50% for ''unattached" (Breslin and others 1969) ~85% for 1 nm (Swift and others 1992) Breathing Rates (Domestic) ICRP66 Defaultsb ICRP66 Defaults Exposure Conditions in Mines Limited size distributions (Cooper and others 1973) 30 distributions recalculated from 1970 data (Knutson and George 1992) Exposure Conditions in Homes Measurements with created exposure scenarios (Li and Hopke 1991) 565 measurements in normally occupied homes (Hopke and others 1995) Hygroscopic Growth Estimated factor of 2 Factor of 2 based on series of published measurements (Li and Hopke 1993, 1994; Dua and others 1995; Dua and Hopke 1996) a Parameters are listed in tables B-8 through B-10 earlier in this appendix. b Although ICRP officially adopted the parameter values after the 1991 NRC report was issued, the parameter values were the same.
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Health Effects of Exposure to Radon: BEIR VI TABLE B-13 Properties of the 220Rn decay chain Type of Radiation Half-Life Radiation Energies (Move) Radon-220 (Th) α 55.6s 6.2883 Polonium-216 (ThA) α 0.15s 6.7785 Lead-212 (ThB) β 10.64h 0.331, 0.569 γ 0.23863, 0.30009 Bismuth-212 (ThC) β (64%) 60.6m 0.67, 0.93, 1.55, 2.27 γ 0.7272 α (36%) 6.051, 6.090 Polonium (ThD) a 0.298 ms 8.7844 Thallium-208 (ThC') b 3.053 min 1.796, 1.28, 1.52 g 2.6146, 0.5831, 0.5107 Lead-208 — Stable across the volume of a room because of the short half-life. Consequently, it is difficult to define representative values of 220Rn. Thus, most of the available measurements have been made of the 220Rn progeny rather than 220Rn. The available data for the ratio of PAEC arising from 220Rn decay products to that from 222Rn are summarized in Table B-14. Rannou (1987) has examined the variation of the PAEC(220Rn) as the PAEC(222Rn) changes and found that PAEC(220Rn) = PAEC (222Rn)0.4 (4) so that the exposure from the 222Rn progeny increases more rapidly than the 220Rn decay products. Thus, for dwellings with high 222Rn concentrations, it appears TABLE B-14 Approximate ratio of PAEC for 220Rn progeny to 222Rn progeny at various locations (Schery 1990) Location PAEC (222Rn) PAEC (220Rn) Italy (Latium), 50 dwellings 1.3 Canada (Elliot Lake), 90 dwellings 0.3 Hungary, 22 dwellings 0.5 Norway, 22 dwellings 0.5 Germany (western), 150 dwellings 0.8 Germany (southwestern), 95 dwellings 0.5 France (Finistere), 219 dwellings 0.3 United States (20 states), 68 measurements 0.6 United States (Colorado), 12 indoor locations (Martz and others 1990) 0.3 Hong Kong, 10 indoor sites 0.8
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Health Effects of Exposure to Radon: BEIR VI that the 220Rn progeny will not be an important additional source of exposure and dose. In the measurements of the activity-weighted size distributions described above, the system used to make the activity measurements included a-spectroscopy so that 220Rn progeny could have been observed. Such activities were not observed and thus, for these homes, only 222Rn decay products provided exposure and dose. Steinhäusler (1996) has recently reviewed the information available regarding thoron exposure and dose. He concludes that the extent of information on 220Rn and its decay products is analogous to the state of affairs for 222Rn before the large-scale national radon surveys of the past 15 years. Few health studies have been conducted on the possible health effects of inhaling thoron decay products. Among residents in high-background areas of Brazil, China, and India, statistically significant increases have been observed for chromosome aberrations. The study in China found no increase in lung-cancer for thoron exposure at a mean concentration of 168 Bqm-3 (Wei and others 1993). Mines Exposure measurements for 222Rn based on gross-counting methods can be affected by the presence of 220Rn decay products. For most of the mines which employed the workers used in the epidemiological studies, measurements are not available to determine the presence or absence of 220Rn progeny. For the mines in the Colorado Plateau and the Grants Mineral Belt of New Mexico, negligible concentrations of 220Rn decay products have been found. There are substantial concentrations of 220Rn decay products in the mines of the Elliot Lake, Ontario (DSMA Atcon 1985), where the PAEC arising from 222Rn and 220Rn were approximately equal. Corkill and Dory (1984) made a retrospective study of exposure in the fluorspar mines of Newfoundland. However, they do not mention 220Rn or its decay products. A similar lack of information exists for the other mining environments. Bigu (1985) has examined theoretical models for estimating the PAEC levels arising from the presence of both radon isotopes. The relative amounts of the radon isotopes and their progeny can be estimated based on the weight ratio of 238U to 232Th. For the Ontario uranium mines, this ratio is assumed to be 1. Results were then calculated based on a number of models that had been previously developed to predict the airborne activity concentrations. Bigu found that a substantial number of the measurements fall outside of the theoretical bounds and there are substantial differences among the results calculated by the various available models. Thus, more sophisticated models will need to be developed to permit adequate estimation of the exposure to 220Rn progeny. Thus, it is not possible to assess the uncertainty from a concentration of 220Rn progeny to lung dose in the miner-based risk estimates. Steinhäusler (1996) also summarized the information available on occupa-
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Health Effects of Exposure to Radon: BEIR VI tionally exposed individuals (iron and rare-earth miners, workers in Th-processing plants and the monazite industry totaling 1557 persons). Increases in respiratory diseases, pancreatic cancer, and chromosome aberrations were found to be statistically significant. Among niobium miners in a Th-rich area, an increase in lung-cancer rates (observed/expected lung-cancer cases = 11.3) was found (Solli and others 1985). Among approximately 53,000 European and Japanese patients treated with thorium oxide injections, an increase in lifetime excess cancer risks for liver and bone cancer as well as leukemia has been observed. However, there were no excess lung-cancers. To reduce the uncertainties concerning the effects of 220Rn decay products, we will need to better characterize dosimetry of thoron progeny and obtain more data on exposures. CONCLUSIONS In this chapter we have reviewed the basis for the dosimetric extrapolation of risk from miners to the general population, taking into account differences in the aerosol characteristics, breathing rates, and the respiratory physiology of the various segments of the population (women, children, infants, as well as men). The constants have changed from the values of about 0.70 to 0.75 reported by the NAS Dosimetry Panel (NRC 1991) to values very close to 1. The differences in input parameters between the calculations are summarized in Table B-13. The primary cause of the shift in K value is attributable to the reduction in the miner breathing rate based on measurements of actual miners in Tajikistan and South Africa. The risks for infants are slightly higher than for miners, but for the other groups, the risks per unit exposure are essentially identical in homes and mines. This value around 1 is reasonable given the compensatory factors of particle concentration raising the airborne concentrations, but in sizes that are only weakly deposited in the lung. Thus, the doses per unit radon concentration are essentially the same although in each case, specific unusual aerosol conditions could substantially increase the dose per unit exposure.
Representative terms from entire chapter: