3.2.1.1. Death

Possible associations between exposure to radon and lung cancer mortality among underground miners are discussed in Section 3.2.1.7 (Cancer).

Excess mortality from noncancer diseases reported in some of the mining cohorts include all noncancer respiratory diseases, pneumoconioses, emphysema, interstitial pneumonitis, other (unspecified) chronic obstructive respiratory diseases, and tuberculosis (Lundin et al. 1971; Muller et al. 1985; Roscoe 1997; Roscoe et al. 1989, 1995; Samet et al. 1991; Tirmarche et al. 1993; Waxweiler et al. 1981). However, confounding factors such as exposure to other respiratory toxicants (most notably arsenic and silica dust), ethnicity, smoking history, and work experience were likely major contributors to mortalities from noncancer respiratory diseases. A statistically significant excess of mortality due to chronic nephritis and renal sclerosis was also reported in the U.S. uranium miner cohort, although it is unclear whether this was related to exposure to radon, uranium ore, or other mining conditions or to nonmining factors (Waxweiler et al. 1981).

No significant association was observed between cumulative exposure to radon progeny and death from cardiovascular diseases in cohorts of German uranium miners (Kreuzer et al. 2010) or Newfoundland fluorspar miners (Villeneueve et al. 2007a).

Limited information is available regarding exposure to radon and death in animals. No significant effects on longevity were observed in male Sprague-Dawley rats exposed to atmospheres of radon and radon progeny for 6 hours/day, 5 days/week during 18 months to obtain a cumulative exposure of 25 WLM (Morlier et al. 1994) or in rats exposed to a cumulative exposure of 20 WLM (1 hour exposures twice weekly for 42 total exposures) or 40 WLM (1-hour exposures twice weekly for 82 total exposures) (Chameaud et al. 1984). It should be noted that these studies employed relatively low dose rates.

3.2.1.2. Systemic Effects

No studies were located regarding gastrointestinal, musculoskeletal, hepatic, dermal, or body weight effects after inhalation exposure to radon and its progeny at exposure levels considered relevant to human health.

Respiratory Effects. Possible associations between exposure to radon and lung cancer are discussed in Section 3.2.1.7. Adverse noncancer respiratory effects have been observed in humans under occupational conditions and in laboratory animals exposed to radon and its progeny. Some studies of miner cohorts identified excess cases of nonmalignant respiratory diseases such as asthma, bronchitis, pneumoconioses, emphysema, interstitial pneumonitis, pulmonary fibrosis, and tuberculosis (Boice et al. 2008; Fox et al. 1981; Lundin et al. 1971; Muller et al. 1985; Roscoe 1997; Roscoe et al. 1989, 1995; Samet et al. 1991; Tirmarche et al. 1993; Waxweiler et al. 1981). However, potential confounding by smoking and respirable dust, especially crystalline silica dust, were likely major contributors to mortalities from noncancer respiratory diseases. Excess mortality associated with exposure to silica dust was reported in California diatomaceous earth miners for which the standard mortality ratio (SMR) was 2.01 (95% CI 1.56–2.55). The mortality rate increased sharply with exposure using a 15-year latency, which is longer than the 5-year latency used for most miner and residential radon studies (Checkoway et al. 1997). Chronic lung disease was reported to increase with increasing cumulative exposure to radiation and with cigarette smoking (Archer 1980). In addition, nonsmoking uranium miners were also reported to have increased deaths from nonmalignant respiratory disease compared to a nonsmoking U.S. veteran cohort (Roscoe et al. 1989). For non-miners, the rate of chronic obstructive pulmonary disease (COPD) was reported to increase with increasing radon exposure (Turner et al. 2012).

Alterations in respiratory function have been reported in studies of U.S. uranium miners (Archer et al. 1964; Samet et al. 1984a; Trapp et al. 1970). Archer et al. (1964) reported decrements in pulmonary function with increasing cumulative exposure; however, the study also noted that pulmonary disability was affected by age and smoking more than by radiation exposure. Samet at al. (1984a) reported significantly increased prevalence of dyspnea with increasing duration of underground mining. Evaluations of these respiratory end points did not include assessment of the effects of each of the other possible mine pollutants, such as ore dust, silica, or diesel engine exhaust.

No information was located regarding respiratory effects in animals following exposure to radon and its progeny at concentrations considered relevant to human health.

Cardiovascular Effects. No significant association was observed between cumulative exposure to radon progeny and death from cardiovascular diseases in cohorts of German uranium miners (Kreuzer et al. 2010) or Newfoundland fluorspar miners (Villeneueve et al. 2007a). A significant relationship was noted between cumulative radon exposure and prevalence of mortality from cerebrovascular disease (excess relative risk of 0.49 [95% CI 0.07–1.23] per 100 WLM) in a French cohort study of uranium miners between 1946 and 1999 (Nusinovici et al. 2010). However, the study authors cautioned that a lack of data limited the ability to assess possible confounding by cardiovascular risk factors.

No information was located regarding cardiovascular effects in animals following exposure to radon and its progeny at concentrations considered relevant to human health.

Hematological Effects. No studies were located regarding hematological effects after inhalation exposure to radon at concentrations considered relevant to human health.

Renal Effects. Although a statistically significant increase in mortality due to kidney disease, characterized by chronic nephritis and renal sclerosis, was reported among U.S. uranium miners (Waxweiler et al. 1981) and in Canadian miners at the Eldorado mines (Muller et al. 1985), this finding is not generally considered to be related to radon exposure per se.

No information was located regarding renal effects in animals following exposure to radon and its progeny.

Ocular Effects. Abdelkawi et al. (2008) reported significantly increased refractory index, decreased protein concentration, and increased protein molecular weight after 6 weeks of exposure in the lens and cornea of mice exposed to radon at a mean concentration of 55.8 kBq/m³ (attached fraction of 0.62) for 6 hours/day, 5 days/week for up to 8 weeks. Effective radon lung doses ranged from 20.92 to 83.68 mSv. The study investigators considered the effects on the lens and cornea to have resulted from systemic distribution of inhaled radioactivity. Although the dermal route of exposure was not considered, the permeability of corneal epithelium to atmospheric gases could allow radon at the high concentrations used in the study to diffuse toward and expose both the cornea and lens to alpha radiation.

3.2.1.3. Immunological and Lymphoreticular Effects

No information was located regarding immunological effects after inhalation exposure to radon at concentrations considered relevant to human health.

3.2.1.4. Neurological Effects

No studies were located regarding neurological effects after inhalation exposure to radon at concentrations considered relevant to human health.

3.2.1.5. Reproductive Effects

No maternal or fetal reproductive effects in humans have been attributed to exposure to radon and its progeny. However, a decrease in the secondary sex ratio (males:females) of the children of male underground miners may be related to exposure to radon and its progeny (Dean 1981; Muller et al. 1967; Wiese and Skipper 1986). Ismail and Jaafar (2010) assessed possible relationships between radiation dose to the lungs from radon and radon progeny and rates of infertility within various locations in Iraqi Kurdistan. Radon levels were measured in homes; annual effective lung doses by inhalation of radon and radon progeny were estimated to range from approximately 2 to 6 mSv/yr. The dose to the gonads was not estimated, but would have been lower than that for the lungs and orders of magnitude lower than the protracted external radiation doses stated to induce temporary sterility (400 mSv/yr) or permanent sterility (2000 mSv/year). The study authors reported an exponential relationship between annual effective lung dose and the rate of male infertility that is not supported by the reported low estimates of radiation dose. The authors also stated that there were increases in blood and prostate cancers (which influence male fertility), but did not provide quantitative data. Overall, the scientific value of this study is questionable.

No information was located regarding reproductive effects in animals following exposure to radon and its progeny at concentrations considered relevant to human health.

3.2.1.6. Developmental Effects

No studies were located regarding developmental effects in humans following inhalation exposure to radon and its progeny.

No information was located regarding developmental effects in animals following exposure to radon and its progeny at concentrations considered relevant to human health.

3.2.1.7. Cancer

Associations between exposure to radon and lung cancer mortality have been examined in studies of underground miners at facilities in the United States (Archer et al. 1973, 1976, 1979; Boice et al. 2008; Checkoway et al. 1985; Gottlieb and Husen 1982; Hornung and Meinhardt 1987; Hornung et al. 1998; Lane et al. 2010; Lubin et al. 1995a, 1995b; Luebeck et al. 1999; Lundin et al. 1971; Moolgavkar et al. 1993; NIH 1994; Roscoe 1997; Roscoe et al. 1989, 1995; Samet et al. 1984b, 1989, 1991, 1994; Schubauer-Berigan et al. 2009; Stayner et al. 1985; Stram et al. 1999; Thomas et al. 1994; Wagoner et al. 1963, 1964; Waxweiler et al. 1981), Australia (Woodward et al. 1991); Brazil (Veiga et al. 2006), Canada (Howe and Stager 1996; Howe et al. 1986, 1987; Kusiak et al. 1993; L’Abbé et al. 1991; Morrison et al. 1985, 1998; Muller et al. 1985), China (Qiao et al. 1989, 1997; Yao et al. 1994), the Czech Republic (Kulich et al. 2011; Ševc et al. 1988, 1993; Tomášek 2002, 2011; Tomášek and Darby 1995; Tomášek and Plaček 1999; Tomášek and Žárská 2004; Tomášek et al. 1993, 1994a, 1994b, 2008), England (Fox et al. 1981; Hodgson and Jones 1990b), France (Amabile et al. 2009; Laurier et al. 2004; Leuraud et al. 2007; Rogel et al. 2002; Tirmarche et al. 1993; Vacquier et al. 2007, 2009), Germany (Brüske-Hohlfeld et al. 2006; Kreuzer et al. 2000, 2010; Schnelzer et al. 2010; Taeger et al. 2006, 2008, 2009; Walsh et al. 2010), Italy (Carta et al. 1994), Norway (Solli et al. 1985), and Sweden (Axelson and Sundell 1978; Bergdahl et al. 2010; Damber and Larsson 1982; Edling and Axelson 1983; Jonsson et al. 2010; Jorgensen 1984; Radford and Renard 1984; Snihs 1974). The mining cohorts were primarily uranium miners, but included some cohorts mining other metals, hard rock, or coal.

Lubin et al. (1997) provide combined results from eleven of these cohorts; the results demonstrate increased risk of mortality from lung cancer with increasing WLM (see Table 3-2). The combined data included 115 lung cancer deaths among workers without known occupational exposure to radon and 2,674 lung cancer deaths among exposed miners. Some of these miners had been exposed to more than 10,000 WLM; the mean exposure among the pooled miner data was 162 WLM. In order to make more meaningful comparisons between radon exposure among mining cohorts and residential radon exposure, Lubin et al. (1997) assessed mortality from lung cancer among two groups of workers with the lowest exposure levels (<50 and <100 WLM, respectively). Even in these groups of miners with relatively low-level exposure, relative risk of lung cancer mortality exhibited an apparent linear and statistically significant increasing trend with WLM (in decile categories). RRs for the two highest exposure categories (142.3–250.8 and >250.8 Bq/m³; equivalent to 3.84–6.77 and >6.77 pCi/L, respectively) were 1.28 (95% CI 1.0–1.6) and 1.20 (95% CI 1.0–1.5), respectively. ERRs per WLM were estimated to be 0.0117/WLM (95% CI 0.002–0.025) for exposures <50 WLM and 0.0080/WLM (95% CI 0.003–0.014) for exposures <100 WLM. General patterns of declining excess relative risk per WLM with attained age, time since exposure, and exposure rate were observed in both the unrestricted pooled data and in those restricted to <50 and <100 WLM. Lubin et al. (1997) noted an apparent inadequacy in the fit of the ERR model, but this was not improved using nonlinear or linear threshold models. Lubin et al. (1997) also noted that results for the mining cohorts with relative low-level exposure might not be applicable to residential radon exposure due to uncertainties in estimating miner exposure to radon, radon progeny, and other carcinogens (e.g., silica dust) during the earliest years of mining when ventilation was poor.

Leuraud et al. (2011) assessed the effects of exposure to radon and radon decay products and smoking status on the risk of lung cancer in a combined analysis of 1,046 lung cancer cases and 2,492 controls with detailed radon exposure data and smoking status selected from three major minor cohorts in the Czech Republic (Tomášek et al. 2003), France (Laurier et al. 2004), and Germany (Kreuzer et al. 2010). The combined analysis resulted in an ERR/WLM of 0.010 (95% CI 0.006–0.018) unadjusted for smoking and an ERR/WLM of 0.008 (95% CI 0.004–0.014) after adjustment for smoking, which was based on four categories: never smoker, ex-smoker for ≥10 years, ex-smoker for <10 years, and current smoker. The results of Leuraud et al. (2011) suggest a sub-multiplicative interaction between radon exposure and smoking.

The results of the miner studies consistently demonstrate significant positive associations between lung cancer and exposure to radon. However, most miner studies were performed prior to the identification of other substances as known human carcinogens (silica dust and arsenic; NTP 2011) or probable human carcinogens (diesel exhaust particulates; Attfield et al. 2012; NTP 2011) in the mining air. Accounting for these carcinogens would likely reduce the calculated impact of radon on lung cancer mortality in the mining cohorts. For example, Xuan et al. (1993) estimated a 75% reduction in the lung cancer risk to a cohort of Chinese tin miners after adjusting for arsenic exposure. Bergdahl et al. (2010) reported decreased lung cancer risk from radon after adjusting for silica exposure within the highest exposure group from a cohort of Swedish iron ore miners. Statistically significant excess lung cancer mortality has been associated with average cumulative exposures to radon progeny as low as 36–39 WLM in Czech and French cohorts of uranium miners (Ševc et al. 1988; Vacquier et al. 2007); exposure levels were higher among many of the other uranium miner cohorts. Vacquier et al. (2009) reported an ERR per 100 WLM of 0.58 (p<0.01) for lung cancer within a cohort of 5,086 French uranium miners (4,133 with positive radon exposure; mean cumulative exposure of 36.6 WLM) and 159 lung cancer cases during 30 years of follow up and noted that higher risk persisted when the effects of hard labor and period of exposure were taken into account. An inverse exposure rate effect (i.e., lower exposure rates for long periods are more hazardous than equivalent cumulative exposure received at higher exposure rates over a shorter time) was evident at relatively high exposure levels (WL) (Hornung et al. 1998; Lubin et al. 1995a, 1997; Luebeck et al. 1999; Moolgavkar et al. 1993; NIH 1994); however, this effect appeared to be attenuated or absent at relatively low exposure levels (Lubin et al. 1995a; NIH 1994; Tomášek et al. 2008). The apparent inverse exposure rate effect could have been associated with using no-threshold models, a restriction that was not required in the European residential assessment (HPA 2009).

Among smoking and nonsmoking uranium miners, the most frequently reported type of lung cancer was small cell lung carcinoma (SCLC) in the early phase of follow-up (Archer et al. 1974; Auerbach et al. 1978; Butler et al. 1986; Gottlieb and Husen 1982; Saccomanno et al. 1971, 1988; Samet 1989). Archer et al. (1974) also noted relatively high rates of epidermoid and adenocarcinomas, while large-cell undifferentiated and other morphological types of lung cancer were seen less frequently. A report on the German uranium mining cohort identified squamous cell carcinoma as the predominant lung tumor cell type, followed by adenocarcinoma and SCLC (Kreuzer et al. 2000). Jonsson et al. (2010) reported lung cancer risk in a cohort of 5,449 male iron ore miners in Sweden; the follow-up period spanned the years 1958–2000. A total of 3,597 of the miners had been exposed to radon; the average cumulative radon exposure was 65 WLM over an average of 14.6 years of employment. For all lung cancers, the ERR per kBq-year/m³ was 0.046 (95% CI 0.015–0.077), which equals an ERR/WLM of 0.022WLM (95% CI 0.007–0.038). For small cell cancer (55 cases), squamous cell cancer (51 cases), and adenocarcinoma (12 cases), ERRs per kBq-year/m³ were 0.072 (95% CI −0.003–0.147), 0.049 (95% CI −0.003–0.102), and 0.000 (95% CI 0.017–0.017), respectively. After adjusting for cumulative quartz (silica) together with attained age and calendar period, the ERR for all lung cancers was 0.031 per kBq-year/m³ (95% CI 0.009– 0.070), which equals an ERR of 0.015 per WLM (95% CI 0.04–0.034).

In a subcohort of 516 white nonsmoking uranium miners (drawn from a larger cohort of U.S. uranium miners), mean exposure was reported as 720 WLM. For this cohort, the mortality risk for lung cancer was found to be 12-fold greater than that of nonsmoking, nonmining U.S. veterans; the SMR was 12.7 (95% CI 8.0–20.1) for lung cancer in nonsmoking mining cohort. No lung cancer deaths were found in nonsmoking miners (Colorado Plateau cohort) who had exposure <465 WLM (Roscoe et al. 1989); continued follow-up of this cohort (1960–2005) revealed SMRs of 2.9 (95% CI 1.8–4.4) for never smokers with cumulative exposure to radon progeny in the range of 400–1,000 WLM and 6.3 (95% CI 4.6–8.5) for never smokers with cumulative exposure to radon progeny ≥1,000 WLM (Schubauer-Berigan et al. (2009). These results for nonsmokers follow a dose rate rather than an inverse dose rate response curve.

Some studies of mining cohorts included assessments of mortality from cancers other than lung cancer. Kreuzer et al. (2010) reported a slight, but statistically significant excess of cancers of the extrathoracic airways and trachea (ERR/WLM = 0.062%; 95%CI 0.002–0.121%) within a cohort of 58,987 male uranium miners at the Wismut mine in Germany with follow-up from 1946 to 2003. Řeřicha et al. (2006) reported significant positive associations between cumulative radon exposures and incidences of leukemia (RR=1.75; 95% CI 1.10–2.78) and incidences of chronic lymphocytic leukemia (CLL) (RR=1.98; 95% CI 1.10–3.59) in a cohort of Czech uranium miners at 110 WLM. In apparent contrast, this same group did not find this relationship to hold in a longer-term follow-up of the same population (Kulich et al. 2011). Möhner et al. (2006, 2010) observed no significant associations between leukemia risk (from acute myeloid leukemia [AML], CLL, or all types) and exposure to radon progeny (from <50 to >1,500 WLM) among 377 leukemia cases and 980 individually matched controls from a cohort of 360,000 uranium miners at the Wismut mine in the former East Germany that supported the Soviet nuclear weapons program. However, the leukemia risk became significant (OR 2.64, 95% CI 1.60–2.35) in the highest dose category when doses were lagged 20 years (Möhner et al. 2010), the same period used by Pelucchi et al. (2006) for the silica dust study, but much longer than 5 years used in other studies in this section. This indicates that the 5-year lag period might be too short. Results of a few other studies indicate statistically significant excessive mortalities from laryngeal, liver, kidney, stomach, and/or gall bladder cancers (Kreuzer et al. 2004, 2008; Tirmarche et al. 1992; Tomášek et al. 1993; Vacquier et al. 2007); however, the excess mortalities from these cancers did not appear to be related to cumulative exposure to radon and were not supported by results from other studies of mining cohorts (Kreuzer et al. 2010; Kulich et al. 2011; Laurier et al. 2004; Möhner et al. 2008).

Numerous residential case-control studies of lung cancer have been performed in the United States and other countries, including Brazil, Canada, China, Croatia, the Czech Republic, Finland, France, Germany, Israel, Italy, Japan, Romania, Spain, Sweden, and the United Kingdom. Some of these studies reported positive or weakly positive associations between lung cancer risk and residential radon concentrations, whereas no significant associations were observed in others. As discussed earlier, recent assessment of available residential case-control studies includes analyses of pooled data from 2 China case-control studies (Lubin et al. 2004), a combined analysis of 7 North American case-control studies (Krewski et al. 2005, 2006), a combined analysis of 13 European case-control studies (Darby et al. 2005, 2006), and a comparative assessment of the three combined analyses (HPA 2009). Pooling resulted in much larger numbers of lung cancer cases and controls than were achieved in individual case-control studies. The results of these analyses of pooled data provide evidence of increased risk for lung cancer with increasing residential levels of radon (Tables 3-3, 3-4, and 3-5) for cigarette smokers, including a statistically significant relative risk of lung cancer at mean radon concentrations ≥542 Bq/m³ (14.65 pCi/L) reported by Darby et al. (2006) (Table 3-5). The HPA (2009) assessment identified flaws that resulted in the China and North American studies overestimating the radon cancer risk. According the HPA (2009), the European study properly assessed radon risk by using more accurate dosimetry and including more homes, and the resulting radon risk estimate paralleled that of the whole China and North American database. However, the China and North American studies selected only a portion of the database, limiting inclusion to those living in one or two homes with more complete dosimetry, but the manner in which homes were selected would have introduced socioeconomic bias, which in addition to lower dosimetry approach, caused the China and North American studies to overestimate the radon risk. Nevertheless, all three combined analyses support the conclusion that residential radon is carcinogenic. HPA (2009) estimated that the cumulative (absolute) radon-induced lung cancer risk to age 75 for long-term exposure to radon at 0, 100, 200, 400, and 800 Bq/m³ (0, 2.7, 5.41, 10.81, and 21.62 pCi/L) is 15, 17, 19, 23, and 30%, respectively, for lifetime smokers, and 0.4, 0.5, 0.5, 0.7, and 0.9%, respectively, for lifetime nonsmokers. Overall, the risk for the lifetime nonsmokers was 25 times lower than for the lifetime smokers (Darby et al. 2006; HPA 2009). Figure 3-1 shows the cumulative absolute risk of death from lung cancer by age 75 years relative to long-term average in-home radon concentration for continuing smokers, ex-smokers, and lifelong nonsmokers in the United Kingdom, as reported by Darby et al. (2006) and HPA (2009).

Figure 3-1

Cumulative Absolute Risk of Death from Lung Cancer by Age 75 Years Versus Long-Term Average Radon Concentration at Home for Continuing Smokers, Ex-Smokers, and Lifelong Nonsmokers in the United Kingdom. Source: adapted from HPA 2009

Assessment of the results of residential case-control studies and comparisons between the presently-available pooled results of the China case-control studies (Lubin et al. 2004), North American case-control studies (Krewski et al. 2005, 2006), and European case-control studies (Darby et al. 2005, 2006) must take into account the effects of exposure measurement error and methodological differences in final analyses. Estimates based on measured radon concentrations will likely underestimate the true risks associated with residential radon, due to misclassification of exposure from detector measurement error, spatial radon variations within a home, temporal radon variation, missing data from previously occupied homes that currently are inaccessible, failure to link radon concentrations with subject mobility, and measuring radon gas concentration as a surrogate for radon progeny exposure (Field et al. 1996, 2002). Generally, if exposure misclassification does not differ systematically between cases and controls, the observed results tend to be biased toward the null (for example, the true effect is actually underestimated). In fact, Field et al. (2002) demonstrated that empirical models with improved retrospective radon exposure estimates were more likely to detect an association between prolonged residential radon exposure and lung cancer. Direct comparisons between the pooled results of the China case-control studies (Lubin et al. 2004), North American case-control studies (Krewski et al. 2005, 2006), and those of the European case-control studies (Darby et al. 2005, 2006) are problematic because only the pooled results of the European case-control studies included regression calibration in an attempt to adjust for some of the measurement error.

Turner et al. (2011) recently reported the results of a large prospective study that found positive associations between ecological indicators of residential radon and lung cancer. The study included a cohort of 811,961 individuals from nearly 1.2 million participants recruited in 1982 for the American Cancer Society Cancer Prevention Study-II. The cohort encompasses 2,754 U.S. counties and 3,493 observed lung cancer deaths as of 1988. Ecological estimates of residential radon concentrations were obtained from the Lawrence Berkeley National Laboratory and were intended to represent the average annual radon concentrations in the main living areas of homes by primary county of residence. The study authors reported a significant positive trend between categories of radon concentrations and lung cancer mortality (p=0.02), a 15% (95% CI 1–31) increase in risk of lung cancer mortality per 100 Bq/m³ (2.7 pCi/L) increase in radon, and a 34% (95% CI 7–68) increase in risk of lung cancer mortality among residents with estimated radon concentrations above the EPA guideline value of 148 Bq/m³ (4 pCi/L).

Information regarding radon-induced lung cancer in animals exposed to radon and its progeny at concentrations considered relevant to human health includes significantly increased incidences of lung tumors in rats repeatedly exposed to radon and its progeny at cumulative exposures as low as 20–50 WLM (Chameaud et al. 1984; Morlier et al. 1994). These results are consistent with the demonstrated associations between lung cancer risk and exposure to radon and radon progeny in occupationally-exposed miners and residentially-exposed individuals.

3.2.2.1. Death

3.2.2.2. Systemic Effects

3.2.2.3. Immunological and Lymphoreticular Effects

3.2.2.4. Neurological Effects

3.2.2.5. Reproductive Effects

3.2.2.6. Developmental Effects

3.2.2.7. Cancer

Information regarding cancer in humans after exposure to radon and its progeny in water is limited to ecological studies. As noted earlier, ecological studies are limited by several factors that may include bias in estimated indoor radon levels, inadequate control of confounding, model misspecification, and misclassification. Radon levels were measured in 2,000 public and private wells in 14 counties in Maine (Hess et al. 1983). The county averages were compared to cancer rate by county to determine any degree of correlation. Significant correlation was reported for all lung cancer and all cancers combined, when both sexes were combined, and for lung tumors in females. Confounding factors (e.g., smoking) were not considered in this analysis. In addition, exposure from radon in these water supplies could have been by the inhalation route as well as the oral route. Results of some ecological studies suggest positive associations between radon levels in ground water sources and incidences of cancers, including lung cancer (Hess et al. 1983), all cancers combined (Mose et al. 1990), and childhood cancer (leukemias and all cancers combined) (Collman et al. 1990). In another study, Collman et al. (1988) found no consistent associations between radon concentrations in ground water and cancer mortality. More recent case-cohort studies in Finland found no significant associations between mean concentrations of radon in well water and cases of stomach cancer (Auvinen et al. 2005) or bladder or kidney cancer (Kurttio et al. 2006).

No studies were located regarding cancer in animals after oral exposure to radon and its progeny.

3.2.3.1. Death

3.2.3.2. Systemic Effects

No studies were located regarding respiratory, cardiovascular, gastrointestinal, musculoskeletal, hepatic, renal, endocrine, dermal, body weight, or metabolic effects in humans or animals following exposure to radon and radon progeny.

Ocular Effects. Abdelkawi et al. (2008) reported increased refractory index, increased molecular weight, and decreased concentration of soluble proteins in the cornea and lens of mice exposed to radon at a mean concentration of 55.8 kBq/m³for 6 hours/day, 5 days/week for up to 8 weeks. Effective radon lung doses ranged from 20.92 to 83.68 mSv for combined dermal contact and inhaled radioactivity. Corneal effects were observed after 2 weeks, followed by lenticular effects (after 6 weeks). The study authors suggested that the corneal effects resulted from direct radon exposure and plating out of radon progeny on the cornea; thus, effects on both the cornea and lens might be the result of combined external and internal exposures.

No studies were located regarding the following health effects in humans or animals after dermal exposure to radon and radon progeny:

3.2.3.3. Immunological and Lymphoreticular Effects

3.2.3.4. Neurological Effects

3.2.3.5. Reproductive Effects

3.2.3.6. Developmental Effects

3.2.3.7. Cancer

A statistically significant increase in the incidence of basal cell skin cancers (103.8 observed vs. 13.0 expected) was observed in uranium miners in the Czech Republic who were exposed to radon and radon progeny for >10 years (Ševcová et al. 1978). The study authors used the permissible concentration of radon progeny in the workplace air (41,000 MeV/L; approximately equal to 31.9 pCi/L) at the time of the study to calculate a skin contamination of 2.3±0.7 kBq/m² (6.22 pCi/cm²), based on total alpha activity per unit area from the radon progeny, for an experimental group of the miners. Assuming 1,700 working hours per year and using a quality factor of 10 for alpha radiation, the study authors calculated a mean dose equivalent to the basal layer of the epidermis of 60 rem/year. The standard quality factor (now termed radiation weighting factor) for alpha particles is 20, which would yield an equivalent dose of 120 rem/year. Additionally, the study authors noted that during earlier times, concentrations of radon progeny in the uranium mines were up to 10 times higher than the permissible concentration in the late 1970s. Exposure to other agents (such as arsenic, a known dermal carcinogen (see Agency for Toxic Substances and Disease Registry [2007a])) in the uranium mining environment, as well as minor traumas of the skin, may also have contributed to the observed incidence of skin cancer. Increased incidences of skin cancer have not been reported in other uranium miner cohorts or for workers in other types of mining, such as metal or coal mines; these end points were not examined in most of these studies.

Eatough and coworkers performed a series of studies designed to estimate the average dose of alpha radiation from radon and radon progeny to the skin under normal environmental exposure conditions in which airborne particles containing radon and radon progeny make dermal contact (Eatough 1997; Eatough and Henshaw 1992; Eatough et al. 1999). In the study of Eatough et al. (1999), results of personnel monitoring of airborne radon progeny ²¹⁸Po and ²¹⁴Po in which individuals were immersed under normal environmental conditions in the United Kingdom indicated that at the average radon concentration of 20 Bq/m³ (0.54 pCi/L) to continuously exposed skin would result in 3,500– 28,000 decays/cm²/year from ²¹⁸Po and 7,000–21,000 decays/cm²/year from ²¹⁴Po. The results of Eatough and coworkers indicate a potential for radon and radon progeny to elicit skin cancer under normal environmental conditions. Few of the alpha particles from those decays would actually reach and expose the dermis since fewer than half would be released toward the body (reducing the values by more than 50%) and most of their energy would transfer to the nonliving epidermis, leaving little to expose the underlying dermis.

No studies were located regarding cancer in animals after dermal exposure to radon and its progeny.

3.4.1.1. Inhalation Exposure

Inhalation exposures to radon deliver the gas and its progeny (e.g., ²¹⁴Bi, ²¹⁴Pb, ²¹⁰Pb, ²¹⁸Po, and ²¹⁰Po; some attached to atmospheric particles and the rest unattached) into the respiratory tract (Marsh and Birchall 2000). Longer-lived radon progeny (e.g., ²¹⁰Pb and ²¹⁰Po) contribute little to the radiation dose to lung tissue because they have a greater likelihood of being physically cleared from the lung by mucociliary or cellular transport mechanisms before they can decay and deliver a significant radiation dose.

Progeny aerosol formation involves distinct physical-chemical processes (Butterweck et al. 2002; El-Hussein et al. 1998; Ishikawa et al. 2003b): (1) immediately after formation, progeny react with gases and vapors and form clusters, referred to as unattached particles, having diameters of approximately 0.5– 3 nm or (2) unattached particles form complexes with other aerosols or particles in air to form attached particles, which can undergo hygroscopic growth to achieve diameters ranging from approximately 50 to 1,500 nm. The magnitude of the unattached fraction in inhaled air depends on the concentration and size distribution of aerosols in the ambient environment, and will vary with the exposure conditions (e.g., indoor, outdoor) and activities of the individual (e.g., sleeping, activities that release particulates into the air) (Marsh and Birchall 2000). The sizes and size distributions of attached particles in smoker homes and mines can be approximately the same, while the particle sizes are smaller in nonsmoker residences (HPA 2009). The unattached fraction for typical indoor environments has been estimated to be 5–20% (Porstendörfer 1994, 2001), while EPA assumes 50%. Smoking and other aerosol-generating activities (e.g., vacuum cleaning, cooking, fireplace and circulating fan usage) will increase the attached fraction, which correspondingly increases the dose (Sun 2008).

Deposition and the subsequent absorption of inhaled radon and radioactive decay progeny are influenced by physiological factors as well as chemical and physical characteristics of the radionuclides, carrier aerosols, and atmospheric particles. Radon is a nonreactive noble gas, and deposition and absorption will be determined by Brownian motion, its solubility in lung fluids, its permeability at the lung:circulating blood interface, and blood flow to the lungs. The blood:air partition coefficient for radon has been estimated to be approximately 0.4 (Nussbaum and Hursh 1957; Sharma et al. 1997); therefore, at steady-state, the blood concentration of radon will be approximately 0.4 times the concentration of radon in lung air. Assuming rapid (i.e., near-instantaneous) partitioning of radon between air and blood, the absorption clearance of radon gas from air in the lung will be governed by the blood flow to the lung (i.e., absorption rate will be flow-limited). At a blood flow to the lung of 5.3 L/minute in an adult, and lung air volume of 2.82 L, the t_1/2 for absorption of radon from lung to blood would be approximately 0.4 min (rate constant=113 hours⁻¹) (Peterman and Perkins 1988). A similar value was estimated for the t_1/2 for clearance of radon gas from the lung air to external air (t_1/2=0.4 minute; rate constant=115 hours⁻¹) (Peterman and Perkins 1988). Rapid clearance of radon gas from the lung by absorption and exhalation will result in steady-state concentrations of radon in blood within 2–3 minutes of initiating exposure to radon gas. Clearance of radon from the blood following removal from exposure will be governed by blood flow rates to major tissue depots for radon (see Section 3.4.2).

Exposures to radon in air occur along with exposures to aerosols of radon progeny, which will deposit on the lung epithelia. The amounts and location of deposition of radon progeny will be determined by factors that influence convection, diffusion, sedimentation, and interception of particles in the airways. These factors include air flow velocities, which are affected by breathing rate and tidal volume; airway geometry; and aerosol particle size (Cohen 1996; James et al. 1994; Kinsara et al. 1995; Marsh and Birchall 2000; Yu et al. 2006). Radon progeny consist of a mixed distribution of unattached and attached particles. Assuming activity median aerodynamic diameters (AMAD) of approximately 1–3 nm for the unattached fraction and 100–200 nm for attached particles (Butterweck et al. 2002; Ishikawa et al. 2003b), deposition fractions of inhaled radon progeny can be estimated from models of particle deposition in the human respiratory tract (ICRP 1994b). The deposition fraction (i.e., percent of total number of inhaled particles that deposit) for unattached particles is predicted to be approximately 97–99%, with most of the deposition (70–80%) occurring in the extrathoracic region of the respiratory tract. The deposition fraction for attached particles is predicted to be approximately 20–40% with most of the deposition occurring in the alveolar region. Deposition will occur more predominantly in the nasal airways when breathing occurs through the nose. These predictions are based on the ICRP (1994b) human respiratory tract model assuming recommended values for deposition fractions for adult members of the general public exposed to homogeneous aerosols, and will vary with different assumptions for breathing rate and ratio of nose-to-mouth breathing (ICRP 2001). Predictions that deposition will be higher for the unattached fraction compared to the attached fraction and higher for nose-breathing exposures are in reasonable agreement with experiments conducted in humans exposed to heterogeneous distributions of aerosols of radon progeny (Booker et al. 1969; George and Breslin 1967, 1969; Holleman et al. 1969; Hursh and Mercer 1970; Hursh et al. 1969a; Ishikawa et al. 2003b; Pillai et al. 1994; Swift and Strong 1996) and with experiments conducted using casts of the human respiratory tract (Chamberlain and Dyson 1956; Cohen 1996; Kinsara et al. 1995; Martin and Jacobi 1972). The deposition fraction in subjects who inhaled (nose-only) 0.5–0.6-nm particles of ²¹⁸Po was estimated to be approximately 94–99% (Swift and Strong 1996). Near complete deposition (>99%) was observed in an adult subject who inhaled (mouth-only) unattached particles of ²¹²Pb formed from decay of ²²⁰Rn in a low ambient aerosol environment, whereas the deposition fraction was 34–60% when the exposure was to aerosols formed in room air and having a particle size range of 50–500 nm, more typical of attached particles (Booker et al. 1969). Deposition fractions for radon progeny have been measured during exposures to aerosols in underground uranium mines (George and Breslin 1969; Holleman et al. 1969). Deposition fractions increased with increasing tidal volume, and decreased with increasing aerosol aerodynamic diameter, from 50–70% for diameters <10 nm to 30–40% for diameters >70 nm. Hursh and Mercer (1970) estimated thoracic deposition (i.e., total of bronchi, bronchioles, and alveolar region) based on external gamma counting of the chest area of ²¹²Pb produced from decay of ²²⁰Rn and inhaled (mouth-only) as aerosols having AMADs of 20–25 or 200–230 nm. The deposition fractions in adult subjects were approximately 50–62% for the smaller particles and 27–38% for the larger particles. When adult subjects inhaled (mouth-only) natural ²¹²Pb aerosols generated from ²²⁰Rn decay in room air, the measured deposition fractions ranged from 14–45% (Hursh et al. 1969a). Pillai et al. (1994) made chest gamma measurements on four subjects who were exposed to ²¹²Pb aerosols for 10–60 minutes in a thorium hydroxide storage facility. The particle size of the ²¹²Pb aerosol was approximately 90 nm. Deposition fraction was estimated to have been 55–76%.

Particles containing radon progeny that deposit in the respiratory tract are subject to three general clearance processes: (1) mucociliary transport to the gastrointestinal tract for progeny deposited in the ciliated airways (i.e., trachea, bronchi, and bronchioles); (2) phagocytosis by lung macrophages and cellular transport to lymph nodes (e.g., lung, tracheobronchial, mediastinal); or (3) absorption and transfer by blood and/or lymph to other tissues. The above processes apply to all forms of deposited radon progeny, although the relative contributions of each pathway and rates associated with each pathway may vary with the physical characteristics (e.g., particle size), chemical form (degree of water solubility), and radiological characteristics (e.g., specific activity).

Absorption half-times (t_1/2) have been estimated for radon decay progeny in adults who inhaled aerosols of lead and bismuth isotopes generated from decay of ²²⁰Rn or ²²²Rn. Values for ²¹²Pb and ²¹²Bi in an aerosol having an activity median particle diameter of approximately 160 nm (range 50–500 nm), a value typical of attached radon progeny particles, were estimated to be approximately 10 and 13 hours, respectively (Marsh and Birchall 1999). The latter estimates were based on an analysis of data from human inhalation exposures to ²¹²Pb and ²¹²Bi progeny of ²²⁰Rn (Booker et al. 1969; Hursh and Mercer 1970; Hursh et al. 1969a; Jacobi 1964; Pillai et al. 1994). However, absorption of unattached radon progeny may be faster than that of attached progeny. Butterweck et al. (2002) exposed nose- or mouth-breathing human subjects to ²²²Rn-derived aerosols that had diameters of approximately 0.3–3 nm, typical of unattached progeny particles. Absorption half-times were estimated to be approximately 68 minutes (range 56–86) for ²¹⁸Po/²¹⁴Pb and 18 minutes (range 17–21) for ²¹⁴Bi. Binding of unattached radon progeny in the respiratory tract may result in slower absorption kinetics. Butterweck et al. (2002) proposed that a 10-hour t_1/2 would apply to the unattached fraction after binding in the respiratory tract, and that the unbound fraction may have an absorption t_1/2<10 minutes. This behavior would be consistent with dissolution of deposited particles being the rate-limiting step in absorption and smaller particles dissolving faster than larger particles.

3.4.1.2. Oral Exposure

Exposure to radon by the oral route can occur as a result of radon gas dissolving in water. At equilibrium, the concentration of radon dissolved in water will be approximately 0.25 of that in air (i.e., Henry’s law constant=4.08 at 20 °C) (NAS 1999b). Radioactive decay of radon in water produces radon progeny; therefore, ingestion of water containing dissolved radon will also result in ingestion of radon progeny. Absorption of radon is thought to occur primarily in the stomach and small intestine, although some absorption may also occur in the large intestine (Ishikawa et al. 2003a; Khursheed 2000; NAS 1999b). Radon is relatively nonreactive and its absorption from the stomach will be determined largely by rates of diffusion of radon from stomach contents to vascularized mucosa; its solubility in the stomach tissues and blood; blood flow to the stomach; and rates of transfer of stomach contents into the intestine (Ishikawa et al. 2003a; NAS 1999b). Diffusion of radon from stomach contents to stomach tissues may be rate-limiting in absorption (NAS 1999b). However, assuming rapid (i.e., near-instantaneous) partitioning of radon from vascularized mucosa to blood, the absorption clearance of radon from stomach mucosa will be governed by the blood flow rate to the stomach (i.e., absorption rate will be flow-limited). At a stomach blood flow of 1% of cardiac output (1% of 6.5 L/minute in an adult), and stomach wall volume of approximately 0.15 L (NAS 1999b), the t_1/2 for absorption of radon from the stomach wall to blood would be approximately 1.6 minutes (rate constant=0.43 minute⁻¹). An absorption t_1/2 of 1–2 minutes is consistent with observations of peak blood radon concentrations and peak radon concentrations in exhaled air within 5 minutes following ingestion of radon in water by adults (Brown and Hess 1992; Hursh et al. 1965; Sharma et al. 1997).

Kinetics of absorption of radon progeny are more complex, reflecting different mechanisms (e.g., membrane cation transport proteins and channels) and sites of absorption for radon and progeny. Absorption of radon progeny following oral exposure is thought to occur largely in the small intestine (Agency for Toxic Substances and Disease Registry 2007b; ICRP 1994c). As a result, absorption of ingested progeny, and progeny formed from radon after ingestion, will be influenced by rates of transfer of stomach contents into the small intestine, as well as rates of absorption of progeny from the small intestine. Ishikawa et al. (2003a) used external gamma counting to measure the kinetics of elimination of ²¹⁴Pb and ²¹⁴Bi from the stomach following ingestion of water containing radon. Elimination kinetics from the stomach exhibited multiple components, with a fast phase (40–50% of ingested activity) having a t_1/2 value of approximately 10 minutes and two slower phases having t_1/2 values of 150 and 240 minutes. The presence of food in the stomach delays stomach emptying and may alter the absorption kinetics of radon and progeny (Brown and Hess 1992; Hursh et al. 1965; Suomela and Kahlos 1972). ICRP (1995c, 2001) recommends values of 0.05 and 0.1 as gastrointestinal absorption fractions for bismuth and polonium, respectively. The absorption fraction for ingested inorganic lead varies with age; from 40 to 50% in infants and children to approximately 8–15% in adults (Agency for Toxic Substances and Disease Registry 2007b; Leggett 1993; O’Flaherty 1993).

3.4.1.3. Dermal Exposure

Data regarding the absorption of radon following dermal exposure are very limited. Dermal absorption of radon has been measured in subjects after bathing in a radon-water spa (Furuno 1979; Pohl 1965) or after application of a radon-containing ointment to the intact skin (Lange and Evans 1947). After bathing for 5–15 minutes, radon concentrations in expired air reached approximately 0.9% of that in the water and ranged from 17.9 to 49.1 pCi/L of air (662–1,817 Bq/m³) compared to pre-bath levels of <1 pCi/L of air (37 Bq/m³). Radon concentrations in the water were reported by the authors as 5,800 pCi (215 Bq)/kg. However, the relative contributions of the dermal and inhalation routes of absorption cannot be determined in these studies (Furuno 1979). Radon concentrations in blood reached 0.85–1% of the radon concentration in the bath water, which was 1.8×10⁵ pCi (4.9×10⁶Bq)/L of water after 30–40 minutes of bathing while breathing compressed air (Pohl 1965). Approximately 4.5% of the radon applied in ointment to intact skin was measured in expired air within 24 hours following application (Lange and Evans 1947).

Peterman and Perkins (1988) proposed a model for simulating the absorption of radon, based on a model largely parameterized to simulate absorption of xenon gas through the skin. Although parameter values for radon were not reported and skin penetration of radon was not modeled, the general structure is potentially relevant to estimating radon absorption rates. In the Peterman and Perkins (1988) model, the rate-limiting step in dermal absorption was considered to be the diffusion of xenon through the skin to the subcutaneous fat. Transfer from subcutaneous fat to blood was assumed to be flow-limited and determined by blood flow to subcutaneous fat. Peterman and Perkins (1988) estimated the dermal diffusion rate of xenon to be approximately 0.18 hour⁻¹. This rate would be equivalent to a t_1/2 value of approximately 4 hours and is substantially slower than the t_1/2 for absorption from lung (t_1/2=0.4 minutes; rate constant=115 hours⁻¹) (Peterman and Perkins 1988). The corresponding t_1/2 value for absorption from subcutaneous fat was approximately 38 minutes (rate=0.018 minute⁻¹), assuming a blood flow of 0.16 L/minute and a tissue volume of 8.2 L.

3.4.2.1. Inhalation Exposure

Based on studies conducted in animals, the distribution of absorbed radon appears to reflect its solubility in water and fat. Nussbaum and Hursh (1957) exposed rats to radon gas in an enclosed exposure chamber (whole body) for periods of 30 minutes to 48 hours and measured tissue radon levels at the conclusion of the exposure. The highest radon concentrations were observed in fat. Tissue:air concentration ratios were as follows (mean±standard error [SE]): omental fat 4.83±0.07, venous blood 0.405±0.016, brain 0.309±0.008, liver 0.306±0.004, kidney 0.285±0.012, heart 0.221±0.013, testis 0.184±0.007, and skeletal muscle 0.154±0.005. Tissue:air ratios for soft tissues reported by Nussbaum and Hursh (1957) are close to those expected for a Henry’s law constant of 4 (i.e., water:air=0.25) and a lipid:air partition coefficient of 6 (Nussbaum and Hursh 1957). For example, assuming fat and water contents of soft tissue of 5 and 70%, respectively, in the rat (Davies and Morris 1993), the tissue:air ratio for soft tissue would be approximately 0.36 if solubility in water and fat were the only determinants of tissue radon levels. The corresponding fat:air ratio would be 5.85 for a lipid content of adipose tissue of 80% and water content of 15%. The tissue:air ratios reported in Nussbaum and Hursh (1957) are the bases of tissue:blood partition coefficients that have been used in various biokinetics models (e.g., Khursheed 2000; NAS 1999b; Peterman and Perkins 1988; Sharma et al. 1997).

Nussbaum and Hursh (1957) also reported information on the kinetics of uptake of inhaled radon in tissues. In all tissues studied, except fat, steady-state levels of radon were achieved within 1 hour of initiating a continuous inhalation exposure. Uptake into omental fat was slower and exhibited fast and slow components having t_1/2 values of 21 and 138 minutes, respectively. The slower uptake kinetics of fat may reflect, in part, the relatively slower blood perfusion of adipose tissue (per unit mass of tissue) compared to other soft tissues. Similarly, relatively slow perfusion of fat should contribute a slower component to total body elimination kinetics following cessation of exposure to radon (see Section 3.4.3).

Information about the distribution of absorbed radon progeny, bismuth, lead, and polonium can be found in reviews of these subjects (Agency for Toxic Substances and Disease Registry 2007b; ICRP 1980, 1994c, 1995c). A relatively large fraction of inhaled ²¹²Pb (inhaled as natural ²¹²Pb aerosols generated from ²²⁰Rn decay in room air) distributes to red blood cells (Booker et al. 1969; Hursh et al. 1969a). Red cell ²¹²Pb burdens, expressed as percent of the lead initially deposited in the respiratory tract, increased from approximately 5% within 1–2 hours following exposure to approximately 50% at times >24 hours following exposure (Hursh et al. 1969a). Long-lived (²¹⁰Pb) and stable progeny (²⁰⁶Pb, ²⁰⁷Pb, and ²⁰⁸Pb), can be expected to deposit and be retained in bone, where approximately 90% of the total lead body burden resides (Agency for Toxic Substances and Disease Registry 2007b). Following chronic exposure in humans, ²¹⁰Pb has been found in bone (Black et al. 1968; Blanchard et al. 1969; Cohen et al. 1973; Fry et al. 1983) and teeth (Clemente et al. 1982, 1984). ICRP (1980, 2001) recommends, for the purpose of modeling bismuth-derived radiation doses, that 40% of absorbed bismuth distributes to kidneys and 30% to other tissues; the remaining 30% is assumed to be excreted rapidly and does not contribute to distribution beyond the central compartment. Retention in kidneys and other tissues are assumed to be the same (elimination t_1/2 values of 0.6 and 5 days for fast and slow phases); therefore, approximately 40% of the body burden of bismuth would be in the kidneys. ICRP (1994c, 2001) recommends the following values for percentages of absorbed polonium distributed to tissues: 30% liver, 10% kidney, 10% red marrow, 5% spleen, and 45% other tissues. Retention in all tissues is assumed to be the same (elimination t_1/2=50 days); therefore, the latter percent distributions will reflect the distribution of the body burden of polonium (e.g., 30% in liver).

3.4.2.2. Oral Exposure

Measurements of the tissue distribution of radon or progeny following ingestion of radon have not been reported. However, as discussed in Section 3.4.2.1, the distribution of absorbed radon appears to reflect its solubility in water and fat; therefore, steady-state distribution following absorption from the gastrointestinal tract would be determined by tissue:blood partition coefficients and the rate of approach to steady state would be determined by tissue blood flows. Based on tissue:air ratios reported by Nussbaum and Hursh (1957) during inhalation exposures of rats (see Section 3.4.2.1 for further discussion), the following tissue:blood ratios (i.e., tissue:blood=tissue:air/blood:air) can be estimated for radon in the rat: omental fat 12, brain 0.76, liver 0.76, kidney 0.70, heart 0.55, testes 0.45, and skeletal muscle 0.38. Therefore, the highest concentrations of radon would be predicted for adipose tissues.

Distribution of absorbed radon progeny would be expected to be similar to the distribution following inhalation exposures, although, first-pass delivery to the liver from the gastrointestinal tract may influence the tissue distribution. As discussed in Section 3.4.2.1, the largest fractions of the body burdens for radon progeny would be expected to be found in bone for lead, kidney for bismuth, and liver for polonium (Agency for Toxic Substances and Disease Registry 2007b; ICRP 1980, 1994c, 2001).

3.4.2.3. Dermal Exposure

No studies were located regarding distribution in humans or laboratory animals after dermal exposure to radon or its progeny. However, as discussed in Sections 3.4.2.1 and 3.4.2.2, the distribution of absorbed radon appears to reflect its solubility in water and fat; therefore, steady-state distribution following absorption from the skin would be determined by tissue:blood partition coefficients and the rate of approach to steady state would be determined by tissue blood flows.

3.4.4.1. Inhalation Exposure

Measurements of exhaled radon following ingestion of radon dissolved in water indicate that absorbed radon is rapidly excreted in exhaled air (see Section 3.4.4.2). Inhaled ²¹²Pb is excreted in urine and feces. Hursh et al. (1969a) estimated that, following inhalation of natural ²¹²Pb aerosols generated from ²²⁰Rn decay in room air, 3% of the amount initially deposited in the respiratory tract was excreted in urine per day and approximately 3%/day was excreted in feces. Longer-term kinetics of excretion of ²¹⁰Pb following chronic exposures to radon progeny may be contributed from slow release of ²¹⁰Pb accumulated in bone (Black et al. 1968). Additional information on the elimination of inhaled radon progeny can be found in reviews of the biokinetics of bismuth, lead, and polonium (Agency for Toxic Substances and Disease Registry 2007b; ICRP 1980, 1994c, 2001). ICRP (1995c, 2001) recommends the following values for the purpose of modeling bismuth-derived radiation doses: a urine:fecal excretion ratio of 1:1 and elimination t_1/2 values of 0.6 (60% of issue burden) and 5 days (40% of tissue burden) for fast and slow phases, respectively. ICRP (1995c, 2001) recommends the following values for elimination of polonium from tissues into urine and feces: a urine:fecal excretion ratio of 1:2 and an elimination t_1/2 value of 50 days.

3.4.4.2. Oral Exposure

Measurements of exhaled radon following ingestion of radon dissolved in water indicate that exhaled air is the dominant route of excretion of ingested radon (Brown and Hess 1992; Gosink et al. 1990; Hursh et al. 1965). Biological elimination kinetics of absorbed radon in exhaled air exhibit multiple phases, with the first half-time ranging from 15 to 80 minutes (Brown and Hess 1992; Gosink et al. 1990; Hursh et al. 1965). Hursh et al. (1965) estimated the following t_1/2 values for fast, moderate and slow phases of biological elimination: approximately 13 minutes (61% of body burden), 19 minutes (34%), and 207 minutes (5%), respectively; 95% of the dose was eliminated within 100 minutes. The slow phase of elimination is consistent with observations made in rats of relatively slow accumulation of radon in adipose tissue during continuous inhalation exposures to radon (Nussbaum and Hursh 1957). The latter t_1/2 values were estimated for subjects who ingested radon in water during fasting. In a subject who ingested radon in water with a meal, moderate and slow phases of elimination appeared to be delayed, with approximate t_1/2 values of 12 minutes (39% of body burden), 60 minutes (51%), and 300 minutes (10%), respectively. Slowing of elimination when radon is ingested with a meal or with lipid has been observed in several studies and may be related to a delay in stomach emptying that alters the absorption kinetics of radon and progeny (Brown and Hess 1992; Hursh et al. 1965; Meyer 1937; Suomela and Kahlos 1972; Vaternahm 1922).

Suomela and Kahlos (1972) estimated radon elimination kinetics in adults who ingested radon in water by monitoring external gamma-radiation from ²¹⁴Bi (i.e., assuming ²¹⁴Bi:²²²Rn disequilibrium ratios ranging from 0.4 to 1). Biological elimination t_1/2 values ranged from 30 to 50 minutes; these are consistent with estimates based on exhaled radon as described above. Out of 10 subjects, ²¹⁴Bi was detected in urine in two subjects (0.4 and 1.8% of ingested ²¹⁴Bi dose; duration of urine collection was not reported). Additional information on the elimination of ingested radon progeny can be found in reviews of the biokinetics of bismuth, lead, and polonium (Agency for Toxic Substances and Disease Registry 2007b; ICRP 1980, 1994c, 2001). In general, the rates and routes of elimination of each progeny absorbed from the gastrointestinal and respiratory tracts are likely to be similar. Information on elimination of inhaled radon progeny is discussed in Section 3.4.4.1.

3.4.4.3. Dermal Exposure

Information on the excretion of radon and its progeny following dermal exposure is very limited. Within 24 hours, 4.5% of the radon, which was applied as a salve to intact human skin, was eliminated by exhalation, while 10% was exhaled after application of the radon to an open wound (Lange and Evans 1947). Bathers breathing compressed air while immersed in radon-containing water had exhaled approximately one-third of radon measured in blood immediately after bathing (Pohl 1965). By 6– 8 minutes after bathing, these persons were exhaling one-half of the amounts exhaled immediately after bathing. The author stated that the remaining radon which distributed to fatty tissue was excreted more slowly.

3.4.4.4. Other Routes of Exposure

Experiments in animals have reported the retention of radon after exposure by the intraperitoneal and intravenous routes. Following intravenous administration, 1.6–5.0% of the administered activity was retained in the animals after 120 minutes (Hollcroft and Lorenz 1949). Retention was greatest at 120 minutes following intraperitoneal administration, but by 240 minutes, it was nearly the same for both routes of administration. These authors also reported that the amount of radon retained in tissues was greater in obese mice than in normal mice, especially after intraperitoneal administration (Hollcroft and Lorenz 1949). Radon retention has also been studied in dogs following intravenous administration of ²²⁶Ra. The amount of radon in bone was found to increase with increasing time after injection (Mays et al. 1975).

Human Respiratory Tract Model for Radiological Protection (ICRP 1994b)

Deposition. The ICRP (1994b) has developed a deposition model for behavior of aerosols and vapors in the respiratory tract. It was developed to estimate the fractions of radioactivity in breathing air that are deposited in each anatomical region of the respiratory tract. ICRP (1994b) provides information used in calculating effective doses of radiation to organs and tissues throughout the body based on a unit intake of radioactive materials. The model applies to three levels of particle solubility, a wide range of particle sizes (approximately 0.0005–100 μm in diameter), and parameter values that can be adjusted for various segments of the population (e.g., sex, age, and level of physical exertion). This model also allows one to evaluate the bounds of uncertainty in deposition estimates. Uncertainties arise from natural biological variability among individuals and the need to interpret some experimental evidence that remains inconclusive. The model has been used for estimating radiation doses from inhalation of radon gas and aerosols of radon progeny; however, it was developed to be applied to a wide variety of radionuclides and their chemical and physical forms.

The ICRP deposition model estimates the fraction of inhaled material initially retained in each compartment (see Figure 3-3). The model was developed with five compartments: (1) the anterior nasal passages (ET₁); (2) all other extrathoracic airways (ET₂) (posterior nasal passages, the naso- and oropharynx, and the larynx); (3) the bronchi (BB); (4) the bronchioles (bb); and (5) the alveolar interstitium (AI). Particles deposited in each of the regions may be removed and redistributed either upward into the respiratory tree or to the lymphatic system and blood by different particle removal mechanisms.

Figure 3-3

Compartment Model to Represent Particle Deposition and Time-Dependent Particle Transport in the Respiratory Tract*. *Compartment numbers shown in lower right corners are used to define clearance pathways. The clearance rates, half-lives, and fractions (more...)

For extrathoracic deposition of particles, the model uses measured airway diameters and experimental data, where deposition is related to particle size and airflow parameters, and scales deposition for women and children from adult male data. Similar to the extrathoracic region, experimental data served as the basis for lung (bronchi, bronchioles, and alveoli) aerosol transport and deposition. A theoretical model of gas transport and particle deposition was used to interpret data and to predict deposition for compartments and subpopulations other than adult males. Table 3-8 provides reference respiratory values for the general Caucasian population during various intensities of physical exertion.

Table 3-8

Reference Respiratory Values for a General Caucasian Population at Different Levels of Activity.

Deposition of inhaled gases and vapors is modeled as a partitioning process that depends on the physiological parameters noted above as well as the solubility and reactivity of a compound in the respiratory tract (see Figure 3-4). The ICRP (1994b) model defines three categories of solubility and reactivity: SR-0, SR-1, and SR-2:

• Type SR-0 compounds include insoluble and nonreactive gases (e.g., inert gases such as H₂, He). These compounds do not significantly interact with the respiratory tract tissues, and essentially all compound inhaled is exhaled. Radiation doses from inhalation exposure of SR-0 compounds are assumed to result from the irradiation of the respiratory tract from the air spaces.
• Type SR-1 compounds include soluble or reactive gases and vapors which are expected to be taken up by the respiratory tract tissues and may deposit in any or all of the regions of the respiratory tract, depending on the dynamics of the airways and properties of the surface mucous and airway tissues, as well as the solubility and reactivity of the compound.
• Type SR-2 compounds include soluble and reactive gases and vapors which are completely retained in the extrathoracic regions of the respiratory tract. SR-2 compounds include sulfur dioxide (SO₂) and hydrogen fluoride (HF).

Figure 3-4

Reaction of Gases or Vapors at Various Levels of the Gas-Blood Interface. Source: ICRP 1994b

Radon gas is categorized by ICRP (1994b) as SR-1, because, even though it has a low reactivity, it is sufficiently soluble to be taken up in the alveolar region where it can be absorbed into blood. ICRP (1994b) recommended default values for regional distribution of inhaled gases (except for those having low solubility) as follows: 10% ET₁, 20% ET₂, 10% BB, 20% bb, and 40% AI. Radon progeny, such as ²¹⁸Po, ²¹⁴Pb, and ²¹⁴Bi are sufficiently reactive to attach to aerosols in the respiratory tract (and external air) and deposit in the respiratory tract according to factors that determine particulate deposition (e.g., sedimentation, inertial impaction, diffusion, and interception). Radon progeny are represented in the ICRP (1994b) model and in extensions of the model (e.g., RADEP) as a mixed distribution of unattached particles (i.e., products of the initial reactions between progeny with gases and vapors) and attached particles (i.e., products of hygroscopic growth of complexes between unattached particles and aerosols in air). AMADs for the two fractions are typically represented in the ICRP model as 1 nm for unattached particles and 200 nm for attached particles (Butterweck et al. 2002; Ishikawa et al. 2003b), although the use of more complex mixed distributions for attached particles has also been used (Marsh and Birchall 2000; Porstendörfer 1994, 2001).

The magnitude of the unattached fraction in inhaled air depends on the concentration and size distribution of aerosols in the ambient environment, and will vary with the exposure conditions (e.g., indoor, outdoor) and activities of the individual (e.g., sleeping, activities that release particulates into the air such as smoking) (Marsh and Birchall 2000). The unattached fraction for typical indoor environments has been estimated to be 5–20% (Porstendörfer 1994, 2001). NRC (1991) recommended a default value of 3% for modeling exposures in homes where smoking occurs and 5% for exposures during cooking or vacuum cleaning activities. The Commission of European Communities recommended a default value of 8% (Monchaux et al. 1999).

Respiratory Tract Clearance. This portion of the model identifies the principal clearance pathways within the respiratory tract. The model was developed to predict the retention of various radioactive materials. The compartmental model represents particle deposition and time-dependent particle transport in the respiratory tract (see Figure 3-3) with reference values presented in Table 3-9. This table provides clearance rates, expressed as a fraction per day and also as half-time (Part A), and deposition fractions (Part B) for each compartment for insoluble particles. ICRP (1994b) also developed modifying factors for some of the parameters, such as age, smoking, and disease status. Parameters of the clearance model are based on human evidence for the most part, although particle retention in airway walls is based on experimental data from animal experiments.

Table 3-9

Reference Values of Parameters for the Compartment Model to Represent Time-Dependent Particle Transport from the Human Respiratory Tract.

The clearance of particles from the respiratory tract is a dynamic process. The rate of clearance generally changes with time from each region and by each route. Following deposition of large numbers of particles (acute exposure), transport rates change as particles are cleared from the various regions. Physical and chemical properties of deposited material determine the rate of dissolution and, as particles dissolve, absorption rates tend to change over time. By creating a model with compartments of different clearance rates within each region (e.g., BB₁, BB₂, BB_seq), the ICRP model overcomes problems associated with time-dependent functions. Each compartment clears to other compartments by constant rates for each pathway.

Particle transport from all regions is toward both the lymph nodes and the pharynx, and a majority of deposited particles end up being swallowed. In the front part of the nasal passages (ET₁), nose blowing, sneezing, and wiping remove most of the deposited particles. Particles remain here for about a day. For particles with AMADs of a few micrometers or greater, the ET₁ compartment is probably the largest deposition site. A majority of particles deposited at the back of the nasal passages and in the larynx (ET₂) are removed quickly by the fluids that cover the airways. In this region, particle clearance is completed within 15 minutes.

Ciliary action removes deposited particles from both the bronchi and bronchioles. Though it is generally thought that mucociliary action rapidly transports most particles deposited here toward the pharynx, a fraction of these particles is cleared more slowly. Evidence for this is found in human studies. For humans, retention of particles deposited in the lungs (BB and bb) is apparently biphasic. The “slow” action of the cilia may remove as much as half of the bronchi- and bronchiole-deposited particles. In human bronchi and bronchiole regions, mucus moves more slowly when it is closer to the alveoli. For the faster compartment, it has been estimated that it takes about 2 days for particles to travel from the bronchioles to the bronchi and 10 days from the bronchi to the pharynx. The second (slower) compartment is assumed to have approximately equal fractions deposited between BB₂ and bb₂, with both fractions having clearance half-times estimated at 20 days. Particle size is a primary determinant of the fraction deposited in this slow thoracic compartment. A small fraction of particles deposited in the BB and bb regions is retained in the airway wall for even longer periods (BB_seq and bb_seq).

If particles reach and become deposited in the alveoli, they tend to become imbedded in the fluid on the alveolar surface or move into the lymph nodes. Coughing is the one mechanism by which particles are physically resuspended and removed from the AI region. For modeling purposes, the AI region is divided into three subcompartments to represent different clearance rates, all of which are slow.

In the alveolar-interstitial region, human lung clearance has been measured. The ICRP model uses 2 half-times to represent clearance: about 30% of the particles have a 30-day half-time, and the remaining 70% are assigned a half-time of several hundred days. Over time, AI particle transport falls, and some compounds have been found in lungs 10–50 years after exposure.

Absorption into Blood. The ICRP model assumes that absorption into blood occurs at equivalent rates in all parts of the respiratory tract, except in the anterior nasal passages (ET₁), where no absorption occurs. It is essentially a 2-stage process, as shown in Figure 3-5. First, there is a dissociation (dissolution) of particles; then the dissolved molecules or ions diffuse across capillary walls and are taken up by the blood. Immediately following dissolution, rapid absorption is observed. For some elements, rapid absorption does not occur because of binding to respiratory-tract components. In the absence of specific data for specific compounds, the model uses the following default absorption rate values for those specific compounds that are classified as Types F (fast), M (medium), S (slow), and V (instantaneous):

• For Type F, there is rapid 100% absorption within 10 minutes of the material deposited in the BB, bb, and AI regions, and 50% of material deposited in ET₂. Thus, for nose breathing, there is rapid absorption of approximately 25% of the deposit in ET; for mouth breathing, the value is 50%.
• For Type M, about 70% of the deposit in AI reaches the blood eventually. There is rapid absorption of about 10% of the deposit in BB and bb, and 5% of material deposited in ET₂. Thus, there is rapid absorption of approximately 2.5% of the deposit in ET for nose breathing, and 5% for mouth breathing.
• For Type S, 0.1% is absorbed within 10 minutes and 99.9% is absorbed within 7,000 days, so there is little absorption from ET, BB, or bb, and about 10% of the deposit in AI reaches the blood eventually.
• For Type V, complete absorption (100%) is considered to occur instantaneously.

Figure 3-5

The Human Respiratory Tract Model: Absorption into Blood. Source: ICRP 1994b

ICRP (1994b) assigned gases and vapors to Type F, unless alternative values for absorption rates are available. However, alternatives to this assumption have been explored, including instantaneous partitioning of radon gas into dissolved blood (Butterweck et al. 2002). Radiation doses from exposures to radon have been estimated assuming radon and its progeny behave as Type F or Type M (Kendall and Smith 2002). The difference between the two categories is important for estimating tissue specific radiation dose coefficients (e.g., Sv/Bq inhaled) because of the relatively fast decay of radon (t_1/2=3.8 days) and its short-lived progeny (e.g., ²¹⁸Po, t_1/2=3.05 minutes; ²¹⁴Pb, t_1/2=26.8 minutes; ²¹⁴Bi, t_1/2=19.7 minutes). Type F materials (absorption t_1/2=10 minutes) will have a smaller proportion of progeny formed in the respiratory tract (i.e., prior to clearance) and, as a result, will deliver a smaller internal radiation dose and smaller dose to the respiratory tract relative to systemic tissues. Type M materials (absorption t_1/2=100 days for 90% of deposited material, t_1/2=10 minutes for 10%) will have a larger portion of progeny formed in the respiratory tract, which will deliver a larger internal radiation dose and larger dose to the respiratory tract relative to systemic tissues (Kendall and Smith 2002). Absorption t_1/2 values for ²¹²Pb and ²¹²Bi, in an aerosol having an activity median particle diameter of approximately 160 nm (range 50–500 nm), a value typical of attached radon progeny particles, were estimated to be approximately 10 and 13 hours, respectively (Marsh and Birchall 1999). Use of a t_1/2 value of 10 hours for radon progeny in the ICRP (1994b) model results in predicted radiation dose coefficients that are similar in magnitude to the Type M assumption (Kendall and Smith 2002). However, absorption of unattached radon progeny may be faster than that of attached particles. Absorption halftimes for aerosols having approximately 0.3–3 nm in diameter, typical of unattached progeny particles, were estimated to be approximately 68 minutes (range 56–86) for ²¹⁸Po and ²¹⁴Pb and 18 minutes (range 17–21) for ²¹⁴Bi (Butterweck et al. 2002). Butterweck et al. (2002) proposed that binding of unattached radon progeny in the respiratory tract may result in slower absorption kinetics. They proposed that a 10-hour t_1/2 would apply to the unattached fraction after binding in the respiratory tract and that the unbound fraction may have an absorption t_1/2<10 minutes (see Section 3.4.1.1 for further discussion of absorption estimates).

The Radon Dose Evaluation Program (RADEP) implements a simplified version of the ICRP (1994b) model and is designed to simulate radon and radon progeny radiation dosimetry (Marsh and Birchall 2000; Figure 3-6): (1) the alveolar interstitial compartment is represented as a single compartment that has a particle transport rate of 0.00661 d⁻¹ to the fast bronchiolar compartment, bb₁; (2) sequestered compartments, ET_seq, BB_seq, and bb_seq are not considered; (3) radon progeny are assumed to not bind to the respiratory tract; and (4) hygroscopic growth of unattached particles is simulated.

Figure 3-6

Simplified Version of the Human Respiratory Tract Model (HRTM). Source: Marsh and Birchall 2000

Validation of the Model. ICRP (1994b) and RADEP have been evaluated with data on deposition and clearance of inhaled particulate aerosol and gases in humans and absorption of radon progeny (ICRP 1994b; Ishikawa et al. 2003b; Marsh and Birchall 1999). Sensitivity and uncertainty analyses of model predictions have been reported (Marsh and Birchall 2000; Yu et al. 2006).

Risk Assessment. The model has been used to establish the radiation dose (Sv) per unit of inhaled radon (Bq) for ages 3 months to 70 years (Kendall and Smith 2002).

Target Tissues. The model is designed to calculate radiation dose coefficients (Sv/Bq) corresponding to specific inhalation exposures to radionuclides. Dose coefficients for radon and progeny have been estimated for all major organs, including the bone surfaces, bone marrow, and liver, and other tissues (Kendall and Smith 2002).

Species Extrapolation. The model is based on both human and animal data. However, it is intended for applications to human dosimetry. Applications to other species would require consideration of species-specific adjustments in modal parameters.

Interroute Extrapolation. The ICRP model is designed to simulate kinetics of inhaled radionuclides. (Note: ICRP/NCRP models are for normal lungs, not those of smokers.)

National Research Council Radon PBPK Model (NAS 1999b)

NAS (1999b) developed a PBPK model for simulating absorption and distribution of ingested or inhaled radioactive materials, including radon gas (Figure 3-7). NAS noted that statistically significant excesses of mortality from leukemia; esophagus, stomach, colon, liver, lung, breast, ovary, and urinary tract cancers; and multiple myeloma had been associated with atomic bomb survivors, but other epidemiological studies were used to establish risk models for thyroid and breast (based mainly on medical exposure data), bone (from radium exposures), and liver (from thoratrast) cancers, along with laboratory animal study data. The NAS cancer risk estimates were based on calculations with specific cancer site risk projection models using the computational method described in Federal Guidance Report No. 13 (EPA 1999a). The model simulates absorption of inhaled radon and distribution to tissues as flow-limited processes (i.e., tissue clearance equivalent to tissue blood flow) with parameters for tissue volumes, blood flow, and blood:tissue partition coefficients. Absorption of radon gas from the stomach and small intestine is simulated as diffusion-limited transfer from the lumen to the wall (i.e., vascularized submucosa), and flow-limited exchange between blood and wall. A separate model is described in NAS (1999b) for estimating wall diffusion rate constants, which predicts a time-integrated radon concentration in the stomach wall of approximately 30% of that of the lumen. Parameter values for adults are presented in Table 3-10. Values for blood flows were derived from Leggett and Williams (1991, 1995); volumes and densities from ICRP (1990); and tissue:blood partition coefficients from Nussbaum and Hursh (1957). Parameter values for infants, children, and adolescents are also presented in NAS (1999b).

Figure 3-7

Schematic Diagram of the NAS (1999b) PBPK Model Developed to Describe the Fate of Radon within Systemic Tissues. Source: NAS 1999b

Table 3-10

Parameters in the NAS (1999b) PBPK Model.

Validation of the Model. The NRC model has been evaluated with data on deposition and clearance of inhaled particulate aerosols and gases in humans and absorption of radon progeny (Correia et al. 1988; Crawford-Brown 1989; Harley and Robbins 1994; Harley et al. 1994; Hursh et al. 1965; NAS 1999b).

Risk Assessment. The model has been used to establish the radiation dose (Sv) per unit of inhaled or ingested radon (Bq) for ages 3 months to 70 years (NAS 1999b).

Target Tissues. The model is designed to calculate radiation dose coefficients (Sv/Bq) corresponding to inhalation or ingestion exposures to radon. Dose coefficients for radon and progeny have been estimated for all major organs, including the bone surfaces, bone marrow, and liver, and other tissues (NAS 1999b).

Species Extrapolation. The model is based on both human and animal data. However, it is intended for applications to human dosimetry. Applications to other species would require consideration of species-specific adjustments in modal parameters.

Interroute Extrapolation. The model is designed to simulate kinetics of inhaled or ingested radon. Extrapolation to other routes of external exposure would require modifications of the model to simulate absorption from those routes.

Khursheed (2000) Model

Khursheed (2000) developed a PBPK model for simulating absorption and distribution of ingested or inhaled radon gas (Figure 3-8). The model is similar in structure to the NRC (NAS 1999b) model, with the addition of a tissue compartment representing breast. The model has not had widespread use in risk assessment, relative to that of ICRP (1994b), RADEP, or the NRC (NAS 1999b) models. Absorption of inhaled and ingested radon, and distribution to tissues, are simulated as flow-limited processes (i.e., tissue clearance equivalent to tissue blood flow) with parameters for tissue volumes, blood flow, and blood:tissue partition coefficients (Table 3-11). Values for blood flows were derived from Leggett and Williams (1991, 1995); and tissue volumes were derived from ICRP (1990). Tissue:blood partition coefficients were derived from Nussbaum and Hursh (1957); however, a single value (0.36) was adopted for all soft tissues, with a higher value used for the gastrointestinal tract and stomach to account for higher fat content of these tissues. Values for partition coefficients for breast and red marrow assumed 30 and 40% fat content, respectively. Although age-dependence of radon biokinetics is discussed in Khursheed (2000), age-specific parameter values for the model are not reported.

Figure 3-8

Khursheed (2000) PBPK Model for Inhalation and Ingestion of Radon Gas. Source: Khursheed 2000

Table 3-11

Parameters in Khursheed (2000) PBPK Model for Radon Gas.

Validation of the Model. The model has been evaluated with data on whole body retention kinetics of radon following ingestion of radon in water (Hursh et al. 1965; Khursheed 2000).

Risk Assessment. The model has been used to predict tissue-specific annual radiation doses associated with continuous inhalation exposures to 20 Bq/m³ of radon, or following ingestion of 1 Bq of radon (Khursheed 2000).

Target Tissues. The model is designed to calculate radiation dose coefficients (Sv/Bq) corresponding to inhalation or ingestion exposures to radon. Dose coefficients for radon and progeny have been estimated for major organs, including the bone surfaces, bone marrow, and liver, and other tissues (Khursheed 2000).

Species Extrapolation. The model is based on both human and animal data (e.g., partition coefficients). However, it is intended for applications to human dosimetry. Applications to other species would require consideration of species-specific adjustments in modal parameters.

Interroute Extrapolation. The model is designed to simulate kinetics of inhaled or ingested radon. Extrapolation to other routes of external exposure would require modifications of the model to simulate absorption from those routes.

Biomarkers of Exposure and Effect

Exposure. Potential biomarkers of exposure may include the presence of radon progeny in urine, blood, bone, teeth, or hair. Although the detection of radon progeny in these media is not a direct measurement of an exposure level, estimates may be derived from mathematical models. Quantification of exposure to radon is further complicated by the fact that radon is a ubiquitous substance and background levels of radon and radon progeny are needed to quantify higher than “average” exposures.

Effect. It has been reported that chromosome aberrations in the peripheral blood lymphocytes may be a biological dose-response indicator of radiation exposure (Bilban and Jakopin 2005; Brandom et al. 1978; Pohl-Rüling et al. 1976; Smerkovsky et al. 2001, 2002). In addition, the frequency of abnormalities in sputum cytology has been utilized as an early indicator of radiation damage to lung tissue (Band et al. 1980); this has not been recommended regarding exposure to radon. However, more extensive research is needed in order to correlate these effects with radon exposure levels and subsequent development of lung cancer or other adverse effects.

Absorption, Distribution, Metabolism, and Excretion. The toxicokinetics of inhaled and ingested radon and radon progeny has been fairly well studied, but information regarding the toxicokinetics of radon and radon progeny following dermal exposure is limited. Additional information on the deposition patterns in airways for radon progeny and the relationship of these deposition patterns to the onset of respiratory disease could help to enhance understanding of the disease process and delineate health protective measures to reduce deposition.

Comparative Toxicokinetics. Similar target organs have been identified in both humans and laboratory animals exposed to radon and radon progeny. More information on respiratory physiology, target cells, lung deposition, and absorption of radon and its progeny in different animal species is needed to clarify observed differences in species-sensitivity and tumor types. For example, rats generally develop lung tumors in the bronchioalveolar region of the lung while humans develop lung tumors in higher regions (tracheobronchial area). These studies could identify the appropriate animal model for further study of radon-induced adverse effects, although differences in anatomy and physiology of the respiratory system between animals and humans require careful consideration. Most of the information available on the toxicokinetics of radon and progeny has been obtained from studies of inhalation exposure. Studies on the transport of radon and progeny following oral and dermal exposures would be of use for comparing different routes of exposure, although oral and dermal exposure routes do not appear to be of particular toxicity concern.

Methods for Reducing Toxic Effects. Lung cancer is generally considered to be the only toxicity concern following long-term exposure to radon and radon progeny. The high-energy alpha emissions from radon progeny deposited in the lung are the source of toxicity concern. The sequence of events leading from irradiation of living cells is generally believed to involve ionization that causes cellular damage that includes DNA breakage, accurate or inaccurate repair, apoptosis, gene mutations, chromosomal change, and genetic instability. Cigarette smoke, crystalline silica dust, and arsenic have been reported to interact with radon and its progeny to potentiate their effects. The quality of breathing air in mines was addressed decades ago and the concentrations of radon, its progeny, silica dust, arsenic, and other mine pollutants that contribute to lung cancer were effectively reduced.

Methods for reducing the potential for radon-induced toxic effects consist of reducing or eliminating smoking, as well as periodically testing for radon in indoor air and reducing radon concentrations to below the EPA recommended action level of 4 pCi/L, using active soil depressurization (ASD) in existing homes and radon-reducing features in new home construction. In some cases, an active soil pressurization system (ASP) may be necessary (see Section 6.5 for explanatory information regarding ASD and ASP systems. Laws are being enacted to eliminate or limit smoking in public areas and business locations. Continued research is needed to develop effective smoking reduction and stop smoking campaigns and to develop and implement additional techniques for reducing radon levels in homes and public buildings.

Children’s Susceptibility. If data needs, relating to both prenatal and childhood exposures, and developmental effects expressed either prenatally or during childhood, are identified, they are discussed in detail in the Developmental Toxicity subsection above.

Age-related differences in susceptibility to the effects of exposure to radon and radon progeny have not been demonstrated. Differences in lung morphometry and breathing rates in children may result in higher estimated radiation doses relative to adults (NCRP 1984a; Samet et al. 1989). However, available information from children employed as miners in China does not provide evidence of increased susceptibility to the effects of exposure to radon (Lubin et al. 1990; NIH 1994).

Child health data needs relating to exposure are discussed in Section 6.8.1, Identification of Data Needs: Exposures of Children.