Indoor Air Assessment: A Review of Indoor Air Quality Risk Characterization Studies

                                       EPA/600/8-90/044
                                            March 1991
    Indoor Air - Assessment

A Review of Indoor Air Quality
 Risk Characterization Studies

    United States:  1989-1990
  Environmental Criteria and Assessment Office
 Office of Health and Environmental Assessment
      Office of Research and Development
     U.S. Environmental Protection Agency
      Research Triangle Park, NC 27711
                                ) Printed on Recycled Paper

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                                   DISCLAIMER                              :

     This document has been reviewed in accordance with U.S. Environmental Protection
Agency policy and approved for publication. Mention of trade names or commercial
products does not constitute endorsement or recommendation for use.
                                       11

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. ' PREFACE ... . j.

In October of 1986 Congress passed the Superfund Amendments and Reauthorization
Act (SARA, PL 99-499) that includes Title IV—the Radon Gas and Indoor Air Quality
Research Act. The act directs that EPA undertake a comprehensive(indoor'air research
program. ' - ' '' '-• V:.,r,..-;•-;-<^;;! •.-.:•-,;•.:••;,-:• i?<-,-..-;—'•.•.•.•-.-/
Research program requirements under Superfund Title IV are specific. THey include
identifying, characterizing, and monitoring (measuring) the sources and levels of indoor air
pollution; developing instruments for indoor air quality data collection; and studying high-risk
building types. The statute also requires research directed at identifying effects of indoor air
pollution on human health. In the area of mitigation and control the following are required:
development of measures to prevent or abate indoor air pollution; demonstration of methods
to reduce or eliminate indoor air pollution; development of methods to assess the potential for
contamination of new construction from soil gas; and examination of design measures for
preventing indoor air pollution. EP A's indoor air research program is designed to be
responsive in every way to the legislation.
In responding to the requirements of Title IV of the Superfund Amendments, EPA-ORD
has organized the Indoor Air Research Program around the following categories of research:
(A) Sources of Indoor Air Pollution, (B) Building Diagnosis and Measurement Methods,
(C) Health Effects, (D) Exposure and Risk (Health Impact) Assessment, and (E) Building
Systems and Indoor Air Quality Control Options.
EPA is directed to undertake this comprehensive research and development effort not
only through in-house work but also in coordination with other Federal agencies, state and
local governments, and private sector organizations having an interest in indoor air pollution.
The ultimate goal of SARA Title IV is the dissemination of information to the public.
This activity includes the publication of scientific and technical reports in the areas described
above. To support these research activities and other interests as well, EPA publishes its
results in the INDOOR AIR report series. This series consists of five subject categories:
Sources, Measurement, Health, Assessment, and Control. Each report is printed in a limited
quantity. Copies may be ordered while supplies last from:
in

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     U.S. Environmental Protection Agency
     Center for Environmental Research Information
     26 West Martin Luther King Drive
     Cincinnati, OH 45268

When EPA supplies are depleted, copies may be ordered from:

     National Technical Information Service
     U.S. Department of Commerce
     5285 Port Royal Road
     Springfield, VA 22161
                                        IV

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•' -. . - •'. ;: ', . :' ABSTRACT ';• ;: -:'" .:;''-"'^"'*''' ;/V;:,— -
. • . • • ' . • L'_' ; .;. •. - -"• ";".. ; _ • = _; •>,- • -^ ' i , '• • *- •

Risk assessment methodologies provide a.mechanism for incorporating scientific
evidence and judgments into the risk management decision process. A risk characterization
framework has been developed to provide a systematic approach for analysis and presentation
of risk characterization study results. This framework was used as a tool, to review published
studies that provide quantitative risk estimates associated with exposure to indoor air
pollutants. Comparisons of both the methods and the resulting risk estimates are presented.
Critical assumptions concerning risk estimates and exposure estimates for each study are
recorded on the framework.
Fourteen risk characterization studies were reviewed, including three studies for radon,
six for environmental tobacco smoke, three for volatile organics, one for formaldehyde, and
one for asbestos. The quality and rigor of analysis varied greatly among the studies
reviewed. Some of the studies clearly state that they are intended to be preliminary analyses
or screening studies, others are reported as sensitivity analyses, and others are detailed risk
assessments. Studies which are technically rigorous in some risk components (e.g.,
dose-response relationships) are often less rigorous in other components (e.g., exposure
assessment).
Summary figures are presented which compare individual lifetime cancer risks estimated
for each pollutant category and the annual cancer mortality attributable to each pollutant
category.
This report is the second in a series of EPA/Environmental Criteria and Assessment
Office monographs: \
I. DEVELOPMENT OF A RISK CHARACTERIZATION FRAMEWORK
II. A REVIEW OF INDOOR AIR QUALITY RISK CHARACTERIZATION STUDIES
IH. USE OF BENZENE MEASUREMENT DATA IN RISK CHARACTERIZATION
ESTIMATES: A PRELIMINARY APPROACH
IV. INDOOR CONCENTRATIONS OF ENVniONMENTAL CARCINOGENS
V. METHODS OF ANALYSIS FOR ENVIRONMENTAL CARCINOGENS

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                             CONTENTS
DISCLAIMER	

PREFACE 	

ABSTRACT	

FIGURES  	

AUTHORS, CONTRIBUTORS, AND REVIEWERS

INTRODUCTION	

SELECTION OF LITERATURE FOR REVIEW .  .

METHODOLOGY  	

OVERVIEW OF RESULTS	

REVIEW AND ANALYSIS OF STUDIES  	
   Radon  	
   Environmental Tobacco Smoke (ETS)   	
   Volatile Organics  	
   Formaldehyde	
   Asbestos	

CONCLUSIONS/RECOMMENDATIONS	
   Conclusions  	
   Recommendations   	

REFERENCES  	

APPENDIX A	

APPENDIX B		
 11

 iii

 v

 vii

viii

 1

 3

 4
 68
 68
 73
 82
 89
 91

 93
 93
 94

 96

101

105
                                 VI

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FIGURES
Number   _.'.'.~; ;".;'-....'..."/ '. .'.'..'.:.._ ^U. "l:^:-.'; _.~;V '.""''--•  :  ~--'  :  '""'

   1     Risk characterization framework ....... . . . . .  ...'. .  .

   2     Comparison of individual lifetime cancer risk due to indoor
         air pollutants   ...... . . . . . .  . ...... . . ......

   3     Comparison of annual cancer cases due to indoor air
         pollutants ... . . . . .  . . . , .... .  . .... .'".. • . • • • •  • • •  •

   4     Radon: U.S. Environmental Protection Agency (1987)  .  .

   5     Radon: National Research Council (1988)  . . . . .  . . .  .

   6     Radon: National Council on Radiation Protection and  ;
         Measurements (1984)   .	

   7     ETS:  Robins (1986) .  . ... .........  . . .... .'.  .

   8     ETS:  Repaceand Lowrey (1985)  . ... . .  . . . . . ... .  .

   9^    ETS:  Russell et al. (1986)  ..>..	

   10     ETS:  Wigleet al. (1987)	 . .'.,.,

   11     ETS:  WeUs(1986)  ... ........ ...-.  .... . ,-....,

   12     ETS:  Pong (1982)  . ... ..... . . ... .... .... . .

   13     VOC: Tancrede et al. (1987)  .................

   14     VOC: Wallace (1986)   , .... ,. . . . .... ..... . .

   15     VOC: McCann et al. (1986) ..................

   16     Formaldehyde: Interagency Regulatory Liaison Group
          (1981)  ...	 . . ... • • ... •	•

   17      Formaldehyde: Tancrede et al. (1987)	

   18      Formaldehyde: McCann et al. (1986)  . ..... ... . ...

   19.    Asbestos:  Mauskopf (1987)   ......... . . .	
                                  Page
                                     8

                                     9

                                    12


                                    15

                                    18

                                    21

                                    24



                                    30

                                    33

                                    36

                                    44
                                    53

                                    55

                                    57

                                    62
    Vll

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                  AUTHORS, CONTRIBUTORS, AND REVIEWERS

     This report was written by Dr. Dennis F. Naugle, Dr. Terrence K. Pierson, and
Dr. Margaret E. Layne of Research Triangle Institute.  The technical input and direction of
Dr. Michael A. Berry (EPA/ECAO) is gratefully acknowledged.  Others who contributed to
the completion of this monograph include Dr. Max Peterson, Dr. Josephine Mauskopf, and
Robert Hetes of Research Triangle Institute. Reviewers were Mr. Charles Ris,  Human
Health Assessment Group, U.S. Environmental Protection Agency, Washington, DC; and
Dr. Christopher De Rosa, Environmental Criteria and Assessment Office,
U.S. Environmental Protection Agency, Cincinnati, OH.
                                       vm

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Risk management has emerged as a principal analytical activity for environmental
managers within the federal government, the private sector, and local governments. Risk
assessment methodologies provide a mechanism for incorporating scientific evidence and
judgments into the risk management decision process. There are four main uses of risk
assessments as input for risk management decision making:
1. To estimate an ambient concentration for a specific chemical that will be
protective of human health and the environment; ••---- =t •:
2. To estimate the human health and environmental effects associated with
current ambient concentrations of a particular chemical;
3. To estimate the human health and environmental effects associated with
releases of one or more chemicals from one or more sources; and
4. To compare estimated risks from specific chemicals in order to set
priorities for regulatory actions or motivation for source mitigation.
The objectives of this monograph are to review publications that provide quantitative
risk estimates associated with exposure to indoor air pollutants and to compare both the
methods and estimates presented in these publications. The focus is to pull together the risk
characterization work for indoor air pollutants or pollutant categories. Analyses that included
exposure assessments, the assessment of dose-response relationships, and the quantitative
characterization of risk were of primary interest. The more qualitative health effects and,
hazard assessments of the numerous studies that have focused on measuring the indopr
pollutant concentration were not reviewed in detail and are not discussed in this monograph.
The Risk Characterization Framework presented in Figure 1 has been developed to
provide a systematic approach for analysis and presentation of risk characterization study
results. .This framework is described in detail in a companion monograph (Naugle et al.,
1990). It is applied in this monograph as the basis for review and comparison of risk
estimates and risk assessment methodologies associated with indoor air pollutants.

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        SELECTION  OF LITERATURE FOR REVIEW


     Extensive literature exists on the health effects associated with indoor air pollutants.  In
order to identify the most relevant information for the purposes of this literature review, the
following resources were utilized:      .                    '  •
         Recent summary documents on the health effects associated with exposure to
         indoor air pollutants (i.e., the Samet et al. [1987] articles entitled "Health
         Effects and Sources of Indoor Air Pollution, Parts I and n", the U.S.
         Environmental Protection Agency report [i987b] EPA Indoor Air Quality
         Implementation Plan and the appendices, and the Proceedings of the Fourth
         International Conference on Indoor Air Quality and Climate held in Berlin,
         West Germany, on August~l7-2i\ 1987 [Seifertet al.; 1987]);         ' '"•'.-;;^;

         The indoor air quality bibliography data base established by the U.S. EPA's
         Environmental Criteria and Assessment Office (ECAO);
       •  On-line search using DIALOG MEDLINE and Environmental Abstracts
         using the key words "indoor air quality and risk or exposure or health
        /effects";.;   .   -,  ,'..-:,, ._'''.,.,..-,•..•,' ....     •  /,':-\, "   ';;:'. ^  .:. >

       •  The bibliographies of documents identified in the previous steps, especially
         review articles;        _                 ;;  -.-. ,.=• *-,   . .../•: •-.'• .-'•• ''-•-.-:: -:. ,.•

       •  Personal contact with individuals in several Federal agencies, particularly
         .the Centers fp.r. Disease Control (CDC) and the National institute of      .
         Occupational Safety and Health (NIOSH), to identify any recent or ongoing
         research that might be relevant to this project; and             :
         Individuals who reviewed an earlier draft of this monograph identified
         additional studies that could be included in this review.
     The identified articles were obtained from RTI files, ECAO, of the librariies of local
universities and reviewed.                                     •  -;      ^ r        ::

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METHODOLOGY
The Risk Characterization Framework was used as a tool for review and analysis of
each document. The components of the risk characterization process identified in the
Framework served as a guide in the literature review and subsequent analyses. Relevant
information on each of these components and the critical assumptions concerning risk
estimates and exposure estimates were abstracted from each document and recorded on forms
provided to each reviewer.
The use of the Risk Characterization Framework as the guide for the review of each
study required a detailed understanding of the risk characterizations presented by the authors,
the critical assumptions used in the analysis, and the equations used to estimate risk. During
the review process it became clear that authors frequently do not provide key information on
all components of the Risk Characterization Framework. For example, the dosimetry factors
used to convert exposure to dose are often not explicitly stated. Also, some risk
characterizations lack detail in one or more of the ten elements in Figure 1 required to fully
evaluate the study. In these instances, the Risk Characterization Framework served as a tool
for identifying studies found to be inadequately documented.
For several of the indoor air pollutants, there are numerous studies in the literature that
report risk estimates, while for others there are few or no such studies. Radon, for example,
is the mostly widely studied of the indoor air pollutants with regard to the characterization of
risk. Since many of these radon studies rely on the same basic data for developing each
component of the risk characterization process, this report presents a detailed review of three
that were viewed as representative of the best available studies. This is also true of the
studies reviewed for environmental tobacco smoke (ETS).
The quality and rigor of analysis varied greatly among the studies reviewed. Some
studies clearly state that they are intended to be preliminary analyses or screening studies,
others are reported as sensitivity analyses, and still others are presented as detailed risk
assessments. Even refined risk estimation studies cannot account for all variances in source
parameters, exposure duration, and concentrations for various populations, human activities,
and complex human physiological factors. Studies which are technically rigorous for some
risk components (such as dose-response relationships) are often very weak in other areas

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(such as exposure assessments).  Although various levels of analysis presented by different
authors were recognized, the three "Quality Levels" suggested in the Risk Characterization
Framework (Figure 1, column A) were not reported since the criteria for uniform reporting
of such levels have not yet been developed.  However, from the detailed analysis of the
components of the risk characterizations provided, the reader will be able to see the strengths
and weaknesses of each study.

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OVERVIEW OF RESULTS
Figure 2 summarizes estimates of lifetime risk of cancer or mortality obtained from the
risk characterization literature reviewed in this work. For many of the studies included in
this review, lifetime risk estimates refer exclusively to lung cancer. However, three of the
articles addressing risk attributable to exposure to environmental tobacco smoke (ETS)
included endpoints other than lung cancer. The Russell et al. (1986) and Wells (1988)
articles include endpoints such as heart disease, bronchitis, and emphysema in addition to
lung cancer in the risk estimates. The Fong (1982) article includes emphysema as well as
lung cancer in the risk estimates. Also, Mauskopf (1987) provides risk estimates for
mesothelioma as well as lung cancer that are attributable to exposure to asbestos. All data
shown in Figure 2 are from column I of each Risk Characterization Framework described
later in Figures 4 through 19. The ranges shown in this figure are for different point
estimates of risk and do not represent a statistical confidence interval or range of uncertainty.
Lifetime risk estimates typically refer to the probability of developing cancer over a lifetime
for the average individual in a defined population and for a specified exposure scenario over
the entire lifetime. Assumptions often are required concerning the exposure scenario and the
, dose-response relationship. The risk estimates refer only to the probability of developing
cancer over a lifetime that is attributable to exposure from the pollutant shown. As can be
seen, such estimates have been developed for radon and its decay products, ETS, a number of
volatile organics (including formaldehyde), and asbestos. The individual studies from which
these estimates are taken are described more fully in the next section.
Figure 3 presents the range of estimates of population risk of cancer obtained from the
literature for each pollutant or pollutant category for which such estimates exist. Population
risk estimates typically refer to the number of cancer cases that are projected for a defined
population and for a specified exposure scenario during one year. Such an estimate is based
on one of two sources of data: epidemiological studies (as in the case of ETS), or animal
studies which are extrapolated to give an estimate of individual lifetime risk to humans and
then combined with some estimate of the exposed population. Assumptions often are
required concerning the exposed population, the exposure scenario, and the dose-response
relationship. The pollutants shown in Figure 3 were the only ones for which, nationwide :

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                                               66
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population risks were reported in column K of Figures 4 through 19. All assumptions and
caveats associated with these population risks are stated in the footnotes and comments of the
referenced figures (Figures 4 through 19).
     Figures 2 and 3 should be interpreted by  considering both their uses and their
limitations as follows:                  '    ;           ,            -,-,.-

       •   Lifetime risks to individuals as shown" in Figure 2 are highly dependent on     "
          local exposure conditions and activity patterns.  The results of broadly
          focused studies by others are intended for comparison purposes and do not
          cover the full range of all local exposure conditions. Despite efforts to
          equally compare the indoor poilutants, differences are bound to exist based
          on varying assumptions by different researchers.

       •   Data on nationwide risks as presented in Figure 3 are important for broad
          policy considerations but are  not widely .reported by researchers.  Additional
          work needs to be done to expand beyond the relatively few reported studies
          and to perform sensitivity analyses to look at the range of outputs given
          various input assumptions.
         Figures 2 and 3 should be used in combination to draw conclusions.  For
         example,  the 5,000 to 20,000 estimated deaths due to radon (Figure 3) are
         greater than reported estimates for other pollutants, yet individual lifetime
         risks due  to environmental tobacco Asmoke and some organic compounds can
         be very high as shown in Figure 2.                                  .'"  .

         For comparisons of the assumptions implicit in the summary data shown,
         readers are encouraged to follow the pathway: leading to greater and greater
         detail. Summary data in Figures 2 and 3 come from column I and
         column K of Figures 4 through 19. These figures clearly show key
         measurements or assumptions in Columns A through K.  They are further
         described in comments below each risk characterization figure.  When
         additional explanation is  required, it is provided in the text near each figure.
         Finally, since no figure and brief analysis can cover all aspects of the
         complex studies analyzed, readers are- encouraged to  refer to the original
         references cited and  listed at the end of this report.
                                          67
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             REVIEW AND ANALYSIS OF  STUDIES

      The following sections present a detailed review of risk characterization studies for
 radon, ETS, volatile organics, formaldehyde, and asbestos.  Each of the studies reviewed has
 a corresponding figure which presents the components of the risk characterizations. The
 discussion is organized by pollutant category.

 RADON
      There are numerous studies in the literature that report risk estimates associated with
 exposure to radon and radon decay products. Most of the risk estimates are based on
 epidemiological studies of underground  miners. Miners are typically healthy males, working
 in an occupation that requires a much-higher-than-average amount of physical activity and
 breathing air containing a pollutant mix (airborne participates, chemical vapors, and a
 relatively high concentration of radon and radon decay products) that is quite different from
 most work or home environments.  As a result, the uncertainty associated with extrapolation
 from the dose-response factors typically derived for miners to a dose-response factor for the
 general population is high (Ellett and Nelson, 1985; Nazaroff and Nero, 1988).
      The information presented in this section is centered around an attempt to compare risk
 estimates from three recent studies:  the U.S. Environmental Protection Agency's Radon
 Reference Manual (1987c), the most recent report from the National Academy of Sciences'
. Committee on the Biological Effects of Ionizing Radiations (National Research Council,
 1988), and the National Council on Radiation Protection  and Measurement's Report 78
 (NCRP, 1984).  Although other risk estimates have been reviewed, these are presented as
 representative of the best available risk estimates.  Some  of the terms used in this discussion
 are defined in Appendix A.

 U.S. Environmental Protection Agency (1987c)
      Figure 4 presents a review of the characterization of risk attributable to exposure to
 radon and radon decay products. The derivation of EPA's risk estimates involves
 assumptions relating to four steps in the estimation process:  (1) determination of radon decay
 product concentrations from radon concentrations;  (2) estimation of cumulative radon decay

                                         68
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product exposure; (3) conversion of individual cumulative exposure to lifetime risk; and
(4) projection of individual lifetime risks to the entire population.  With regard to the
relationship between radon and radon decay product concentration, it has been found that the
equilibrium fraction (defined in Appendix A) ranges from 0.3 to 0.7, with an average of
approximately 0.5.  Using the average equilibrium fraction of 0.5, a ratio of 200 picocuries
per liter of radon to  1 working level (WL) of decay products is considered to be fairly typical
for residential environments.  EPA used this ratio to convert radon concentrations to the
average working level of 0.004 shown in column C of Figure 4.
     The estimation of average lifetime dose of radon and radon decay products requires
several assumptions in order to convert cumulative residential exposures to working level
months (WLMs), defined as the integrated exposure a miner receives during 170 hours in a
one-working-level (WL) environment.  First, an adjustment factor for exposure duration is
needed.  EPA assumed that a resident is exposed to a given radon level 75 percent of the
time (shown in Figure 4,  column D), as opposed to 170 hours per month.  Second, an
adjustment factor for inhalation rate is required  since, on average, the inhalation rate of a
miner is considerable higher than trie rate for the general population as a result of the miner's
increased physical activity.  EPA assumed the average breathing rate of an adult is  15.3 liters
per minute while the rate for a miner is about 30 liters per minute.  This ratio  (15.3 to 30)
was used by EPA to correct the estimate of cumulative residential exposure expressed in
WLM (shown in Figure 4, column F).  EPA's estimates of cumulative exposure do not
account for potential variations in lung cell sensitivity or differences in sensitivity of different
individuals.^
      Estimates of the risk of lung cancer associated with exposure to radon and radon decay
products are obtained primarily from epidemiological studies of underground miners. Since
 miners in the study are still living, the risk observed  during part of  a lifetime must be
 extrapolated to a whole lifetime.  EPA has adopted a relative risk model to estimate an
 increase in risk per WLM of one percent to four percent.  EPA's calculations are based on an
 assumed linear dose-response relationship  and a minimum latent period of 10 years.
      EPA has projected lifetime risk to the entire population resulting in predictions of 5,000
 to 20,000 deaths per year from radon exposures.  The following  equation was used:
                                           69
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Total LCDs from Indoor Radon = CR X T X
                                                                x TCR X POP
 where:
            CR      =  average (mean) lifetime indoor radon decay product concentration
                    =  0.004 WL-life

            T       =  average interval of lifetime exposure in hours, following a 10 year
                       minimum induction period during which no lung cancer will be
                       observed, assuming 75% occupancy and 73.88 year life span (1980
                       vital statistics)
                    =  .75 X (73.88-10) x 365  X 24 = 419,691.6 hours/life

            FWLM   =  factor converting average cumulative indoor exposure in WL hours
                       to working level months (WLM) for a miner (since risk estimates are
                       based on miner data), accounting for 170 hours per month exposure
                       period per WLM (by definition), and the difference in breathing rate
                       between the average adult (15.3 liters per minute) and a miner
                       (30 liters per minute)
                    =  1/170  x 15.3/30 = 0.003 WLM per hour

            RRRM   =  relative lung cancer risk for lifetime exposure to radon, per WLM,
                       using relative risk model
                    =  1% to 4% per WLM                                        \

            TCR     =  underlying annual average of U.S. lifetime lung cancer risk (1980
                       vital statistics)
                    =  4.584 x  10'4 per person.

            POP     =  1980 U.S. population
                    =  226,545,805
National Research CouncU (1988)

     A relative-risk, time-since-exposure model is used by the National Research Council's
Committee on the Effects of Ionizing Radiation (National Research Council, 1988) for

computing an age-specific lung cancer mortality rate attributable to exposure to radon and

radon decay products.  Components, with appropriate conversion of units to fit the generic
Risk Characterization Framework, are shown in Figure 5.

     The National Research Council (1988) report uses the time-since-exposure (TSE) model

to generate tables of gender-specific lifetime risks of lung cancer mortality by age exposure
                                         70
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started and age exposure ended and at annual exposures ranging from 0.10 WLM/year to
20.00 WLM/year.
     The mathematical form of the TSE model is:
                                   •  .                             «       •
                          r(a) = r0(a) [I+/8«y(a) (^ + O.SW^]
where:
     r(a)  =  age-specific lung cancer mortality rate from all causes
     r0(a), =  age-specific background lung cancer mortality rate from all causes other than
              radon                                                            •    -• •
     /}     =  the slope of the dose-response relation
     (a)   =  effect of age on risk (1.2 for ages  < 55, 1.0 for ages 55 to 64, and 0.4 for
              ages > 65 years)                                                 ,
     Wt   =  WLM of exposure incurred between 5 and  15 years before current age

                                                                      f
     The steps used in applying the TSE model are:
     1.  Each year of the period of interest is placed in the appropriate interval:
         a.  5-15 years before current age,
         b.  > 15 years before current age;
     2.  The total annual  risk for the person's age is calculated using the TSE risk equation
         shown above (using an appropriate background age-specific risk r0(a));
     3.  The calculated value of r(a) is multiplied by the chance of surviving all causes of
         death to that age (including the increased risk due to exposure).
The choice of an appropriate age-specific background rate is dependent on proper treatment
of smoking as well as gender and calendar time.
     National Research Council (1988) estimates the unit lifetime risk (response factor) of
lung cancer mortality due to a lifetime exposure to radon progeny for males at 5.06 X 10"4
per WLM of exposure and for females at 1.86  x 10"4 per WLM of exposure. This estimate
is based on a lifetime (69.7 years for males  and 76.4 years for females) exposure to
0.1 WLM/year of radon decay products.  Conversion of unit lifetime risk to average lifetime
individual risk is shown in Figure 5, column I.
                                           71
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National Council on Radiation Protection and Measurements (1984)

     The NCRP risk model relies on the Harley and Pasternack (1981) Model B of lung

cancer excess due to radon progeny.  Components, with appropriate conversion of units to fit
the generic Risk Characterization Framework, are shown in Figure 6.


     The mathematical form "of the model is:
                                              85
                                   LR(t0) = 2A(t|t0)
where:

           LR(to)  = the lifetime risk from a single annual exposure at time t0

           A(t| y = attributable annual tumor- rate at age t (t > 40) due to a single annual
                      exposure at t0.  If exposure occurs after age 35, risk commences at
                      t0 + 5 years, if exposure occurs before age 35, risk commences at
                      age 40.

The equation used to calculate A(t| t0) is the following:
where:
                             A(t|g  = RC (Pt/Pto)
           RC    =
risk coefficient; assumed 10  x 10"4 cases/yr/WLM, which is an
average value of all epidemiology studies reviewed
                 = life-table correction to account for death from other causes
                 = probability that an individual will be alive at age t0.
                 = probability that an individual will be alive at age t.

                 = decrease in rate of risk expression due to repair, cell death, or
                    unspecified mechanisms.
                                          12
-------
    The model includes the following assumptions:
      1. Based on a review of epidemiologic data on underground miners, estimates
         of attributable risk range from l.f> to 45 X 10"6 with a rounded average
         value of 10  x 10"6 cases per year per WLM of exposure. This average
         value was used in the calculations in the National Council on Radiation
         Protection and Measurements (1984) report.

      2. Following a latent period,  the tumor rate is an exponentially decreasing
         function of the time since exposure.

      3. Disease rate excess associated witira single exposure increases with age at
         exposure.

      4. Lung cancer is rare before the age of 40 years.

      5. Median age of lung cancer among miners is about 60 years in nonsmokers
         and 50 years or older in smokers,

      6. The minimal time for tumor growth, from initial cell transformation to
         clinical detection,  is 5.years.  :,.:-^:v:!>;;
The model specifies a 5-year latency period for persons first exposed at age 35 or older and a

(40 - u)-year latency period for persons under the age of 35 years.  To obtain lifetime risk

due to a single exposure at age u, the equation is integrated over t from age 40 to maximal

assumed life (age 85 years).  To obtain the excess risk at t due to all previous exposure, the

equation is integrated over years of exposure,-uj,...,un.  To obtain lifetime excess risk from

all exposures, the equation is integrated over t and u.            , „ •
     The NCRP model predicts 130 lung-cancer/deaths per 106 persons per WLM of

exposure.  The estimate'is based on an average lifespan of 70 years and a lifetime risk for

lifetime exposure starting at age 1 of 9.1 X 10"3. Conversions to average lifetime;individual

risk estimates are shown in Figure 6, column I.


ENVIRONMENTAL TOBACCO SMOKE (ETS)

      A number of studies have been completed which estimate the annual number of deaths

attributable  to exposure to environmental tobacco smoke (ETS) and the lifetime individual

risks  attributed to such exposures.  Six of these studies are reviewed using the Risk
                                          73
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 Characterization Framework.  These reviews are presented in Figures 7 through 12.
 Although different methodologies and source data are employed by each of the authors, most
 of the estimates for both lifetime individual risk and annual death attributable to exposure to
 ETS are similar.  The predominant health effect that has been assessed is lung cancer.
 However, Fong (1982) includes lung cancer and emphysema together in his population risk
 estimates; Russell et al. (1986) include heart disease, bronchitis,  emphysema, and lung cancer
 together in their risk estimates; and Wells (1988) provides separate risk estimates for other
 cancers and for heart disease.

 Robins (1986)
     The most mathematical analysis of risk attributable to exposure to ETS  is Robins'
 (1986) appendix to the National Research Council's (1986) report entitled "Environmental
 Tobacco Smoke: Measuring Exposures and Assessing Health Effects",  which presents a
 sensitivity analysis approach to risk assessment that considers both the epidemiological data
 and some measures of exposure to the constituents  of ETS.  Study components are shown in
 the Risk Characterization Framework in Figure 7.  Robins' estimates of the lifetime risk of
 lung cancer attributable to ETS in a nonsmoker with moderate ETS exposure are developed
 from an analysis that considers 30 different "exposure histories", three different estimates of
 the coefficients of the dose-response model, and two different estimates of the overall
 summary rate ratio (i.e.,  average relative risk) which is assumed to be the ratio of the "true"
 relative risk in exposed subjects to that in unexposed  subjects.  ("True"  is used here to mean
 that this is, in fact, an assumed true value and not the mean of a sample.)
     Two different average relative risk ratios are used in this analysis.  A weighted average
 of the average relative risk ratios from  13 epidemiological studies presented in the National
 Research Council report is calculated to be roughly 1.3.  In the analysis, the  author assumes
 that a weighted average of 1.3 is causally related to differences in ETS exposure between
 "exposed" and "unexposed" individuals and not to bias.  A second average relative risk ratio
 of 1.14 was calculated based on the subset of these epidemiological studies that were
identified as U.S. studies. The assumption underlying the use of these two different average
relative risk ratios is that the true average relative risk ratio is most likely 1.3 and at least
 1.14.
                                          74
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    : In this risk assessment, the most important assumption was that the average relative risk
was caused by ETS exposure.  This assumption is basic to the equations used for estimating
the true relative risk, the dose (expressed as the carcinogen-equivalent number of actively
smoked cigarettes inhaled daily), and the.subsequent estimates of lifetime risk of death from
lung cancer.  The following is a presentation of the general equations used by the author:
where:
           EXCESS^)       =  tf16 excess of lung cancer deaths at age t due to a specific
                              exposure history                                    : ...,..;;
           70(t)            =  the incidence of lung cancer death at age t in the absence of
                              all exposure to ETS                     '••-'•''
           RREXCESS(t)    =  the excess relative risk for lung cancer due to a particular
                              exposure history.  Mathematically, it can be expressed in the
                              following way:
                                                    - 1
where RR(t) is the true relative risk at age t for the exposed group compared to a completely
unexposed group.  The terms in this model are described mathematically by the author in his
article (see National Research Council, 1986, pages 325-326),
     One method the author employs in the analysis. is to estimate the coefficients of a
five-stage cancer process under the assumption that cigarette smoke influences the first and
fourth stages of the process. The author further assumes that the ratio of the effect (on a
multiplicative  scale) on stage four to that of stage one is the same for ETS and mainstream
smoke.  Thus, an estimate of this ratio can be obtained by fitting a five-stage cancer model to
data on  the lung cancer experience of active smokers. : The author  calculates this ratio based
on two different data sets and calculates a third ratio based  on' an adjustment to one of the
data sets.                                                      .
     In addition to the two average relative risk ratios (i.e., 1.3 and  1.14) and the three
different sets of values for the coefficients of the multistage model, the author developed

                                           75
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30 exposure histories for the exposed and unexposed individuals.  The 30 exposure histories
describe the percentage of time individuals in various age groups were exposed to various
ETS dose rates (relative to the current ETS exposure of an adult nonsmoker without a
smoking spouse).  The author states that smaller differences postulated between the lifetime
ETS exposure of "exposed" and "unexposed" individuals will be associated with larger
estimates of the true relative risk.  Thus, some exposure histories were selected that would
modestly underestimate the true difference in exposure between the "exposed" and
"unexposed" study subjects, and others were selected that would modestly overestimate this
difference. Shown in Figure 7 are the maximum and minimum estimates from the range of
lifetime risk estimates for males and females covering the 30 exposure histories and three
response factors for the relative risk ratio of 1.3.  Also presented by the author but not shown
in this figure were the same calculations using a relative risk ratio of 1.14. The range of
estimates of lifetime individual risk (shown in Figure 7,  column I) for nonsmokers  is based '
on three separate estimates of the coefficients in the multistage model and the range in
exposure calculated for the 30 different exposure histories. Population risks expressed as the
total number of lung cancer deaths per year among nonsmoking women (or men) (AN)
attributable to  ETS in 1985 were calculated by Robins (1986)  using the following equation:
                            AN = 2 AF (t)  X I0(t) X N(t)
                                   t        '
where:
           N(t)   =  the number of nonsmoking women (or men) at risk at age t in  1985
           I0(t)   =  the age-specific death rate among nonsmoking women (or men) in
                     1985
           AF(t)  =  the age-specific fraction of lung cancer deaths due to ETS exposure in
                     nonsmoking women (i.e.,  the average relative risk minus 1, divided by
                     the age-specific relative risk)
                                         76
-------
..;:,  In order to estimate age-specific relative risk among nonsmoking women (or men), age-
specific estimates of the probability of being married to a smoker and the true relative risk in
exposed and unexposed subjects are needed. The age-specific estimates of the probability of
being exposed are taken from Garfinkel et al. (1985).  Relative risk estimates were .calculated
in the same way as described for lifetime individual risk estimates.
     Robins (1986) provides estimates of ETS attributable lung cancer deaths among U.S.
nonsmokers in 1985 based on epidemiologic data, as shown in column K of Figure 7.
Estimates are presented for females and males separately with a maximum and minimum
across all exposure histories and five choices of dose-response coefficients. As shown in,
Figure 7, the maximum and minimum estimates calculated based on epidemiological data are
3,220 and 1,768 for females and 1,942 and 721 for males respectively.
     Similar sensitivity analyses are presented by Robins (1986)' for ex-smokers andr   :    ,>
continuing smokers for an alternative  true relative risk ratio of 1.1.4 and for dosimetrie
measurements rather than epidemiological studies,: For example, the author estimates the;
individual lifetime risk of lung cancer attributable to other people's cigarette smoke for an.
ex-smoker who smoked one pack per  day from age 18 to 45 and was moderately exposed to
other people's cigarette smoke  lies between 5.2 x 10"3 and 2,03 x 10"2. Note these risks
are about 30 percent higher than the estimates shown in Figure 7, column I, for lifetime
nonsmokers.  Similarly, use of a relative risk ratio of 1.14 rather than 1.3 provides a range
of estimates of annual lung cancer deaths for female nonsmokers of 935 to 1,730 and for
male nonsmokers  of 360 to 980.  These estimates are approximately 50 percent lower than
the estimates shown in Figure 7, column K.

Repace and Lowrey
     The Repace  and Lowrey (1985)  estimates of annual risk of limg cancers attributable to
exposure to ETS are based on estimates that U.S. nonsmokers are exposed to between 0 and
 14 mg of tobacco tar per day with the typical nonsmoker exposed to 1.4 mg i>er:day.  A
calculation based on age standardized differences in lung cancer .mortality rates between
Seventh-Day Adventists (SDAs) who  never smoked and demographically comparable
non-SDAs who never smoked (from studies by Phillips et al.,  1980) yields a passive smoking
risk rate of 7.4 X  10"5 per year. This was computed from an estimate that about 4,700 lung
                                          77
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cancer deaths per year were attributable to exposure to ETS among the 62.4 million U.S.
                                                                                 i
nonsmokers. Results are shown in Figure 8.  Note that data from the authors' reference are
in columns I, J, K, and E.  Column H was back-calculated in this work.
     Seventh-Day Adventists have a lifestyle which avoids smoking and is oriented toward
socialization with co-religionists.  Moreover, 40 percent of the study group worked for
church-run organizations;  so the SDA lifestyle involves less exposure to ETS than
demographically comparable nonsmokers from the general population.
     The phenomenological exposure-response relationship derived by Repace and Lowrey
(1985) was shown by them to predict to within 5 percent the lung cancer mortality rate and
mortality ratio reported in the American Cancer Society cohort studied by Garfinkel (1981) in
his study of passive smoking and lung cancer, as well as to explain the differences between
the study of Garfinkel (1981) and the cohort studied by Hirayama (1981). Repace and
                                                                             t
Lowrey (1986), in a refinement of their estimates to adjust for sex and standardized to the
total nonsmoking population, calculated a total of about 4,891 lung cancer deaths per year
from passive smoking, of which 1,441  were estimated to be male and 3,450 female.  These
estimates are shown in parentheses in Figure 8, column K. These refined estimates are
within 5 percent of their earlier estimates and are consistent with the results of a later
case-control study by Garfinkel et al. (19.85) to within 5 percent and  to within 5 percent for
the weighted lung cancer mortality ratios of the 13 epidemiologic studies of passive smoking
and lung cancer as analyzed by the National Research Council report in  1986 (Repace  and
Lowrey, 1987).
     The methodology for calculating these adjusted estimates of lung cancer deaths
attributable to passive smoking is to start with the age-specific estimates for both sexes
combined.  These age-specific  estimates are then  separated into male and female estimates of
lung cancer deaths, assuming 28.6 percent occur in men as presented in Arundel  et al.
(1986).   The percent of over- or under-estimation for each sex in each age group is derived
from Arundel et al. (1986).  These percentages are used to adjust the estimates of lung cancer
deaths for each sex in each age group.  For three of the age groups (e.g., 35-44, 45-54, and
55-64) both male and female estimates of lung cancer deaths were adjusted downward  while
for the remaining two age  groups (65-74 and 74+) estimates of lung cancer deaths for both
sexes were adjusted upward.
                                          78
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Russell et al. (1980)
     Urinary nicotine concentrations in smokers and. nonsmokers are used to estimate
exposure and risk to nonsmokers from ETS in Russell et al. (1986). This analysis is
summarized in Figure 9.  Urinary nicotine levels were chosen as a chemical marker for
smoke intake because nicotine is specific to tobacco smoke, and urinary nicotine is stable and
thus is suitable as a measure of tobacco smoke exposure over several hours. ,The levels easily
differentiate "exposed" and "non-exposed" nonsmokers.. The average concentration in
nonsmokers who reported some exposure was 15.5 ng/ml, three times the value of 5.2 ng/ml
for those who reported no recent exposure.  The average urinary nicotine concentration of
188 nonsmokers from four studies was 10.8 ng/ml compared with 1,471 ng/ml in a combined
sample of 229 smokers.  On the basis of these measurements they estimated that nonsmokers
receive on average about 0.7 percent of the nicotine dose of active smokers.
     The authors assume that the dose-response effect at low levels is linear.  This leads to a
further assumption that a fair .estimate of deaths due to passive smoking can be based on a
proportion of deaths attributed to active smoking.  In other words, the risk of death from
passive smoking was assumed by the authors to be approximately 0.7 percent of that due to
active  smoking.  The authors calculate the number of premature deaths per year among
nonsmokers is estimated to be about 1,000 in Britain and over 4,000 in the U.S. as shown in
column K of Figure 9.                                    .             (      r

Wigle et al. (1987)         ;
     Information on the proportion of people who had never smoked among victims of lung
cancer m Canada is compiled in Wigle et al. (1987).  This analysis is summarized in Figure,
10. The total number of lung cancer deaths per year attributable to ETS in Canada was.   .
estimated to be 330.  This number was derived by applying the age- and sex-specific rates of
death from lung cancer attributable to ETS that were estimated by Repace and Lowrey to the
Canadian population at risk as determined by the 1983 survey on  smoking habits of .
Canadians.  The U.S. population is approximately 10 times the population, of Canada; ,
therefore 3,300 deaths from lung cancer were estimated in the U.S. based on .this study. ,
                                         79
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Wells (1986)
     The epidemiological literature on ETS and adult lung cancer, other cancers, and heart
disease is reviewed in Wells (1986).  The primary model that was used by the author to
estimate risk combined relative risks from the various studies that pertained to a given sex
and disease and assumed that the combined relative risk was constant with age.  A combined
relative risk was calculated using the following equation:
                 Rcb = exp
              Wco In Rco +
                                                  In Rcc
                                      W  + W
                                       "co ~ vvcc
where:
           w    =
             co
           w    =
             cc
=  the combined relative risk
=  the relative risk for cohort studies
=  the relative risk for case control studies
   the weight for cohort study
   the weight for case control study
     The weights for the cohort and case control studies are the inverses of the respective
variances.  The excess death rates attributable to lung cancer for never smokers for passive
smoking (Dpx) for each sex are shown in Figure 11, column K.  These estimates were
calculated for each sex and five-year age range from never smoker death rates (Dns) by the ,
following equation:
where:
           Fp  =  the fraction of the population that is exposed
           R  =  the combined relative risk (Rcb from above)
                                          80
-------
Deaths were then calculated by multiplying the passive smoking excess death rate by the
exposed population for each sex and five-year age interval, then summed. The sum for males
and females is shown in Figure 11, column K.
     For female lung cancer three cohort studies and 14 case control studies were combined.
The overall combined relative risk is 1.44 with a 95 percent confidence limit of 1.3 to 1.7.
The male lung cancer relative risk was based on a combination of two cohort studies and
seven case  control studies. The overall combined relative risk is 2.1 with a 95 percent
confidence  limit of 1.3 to 3.2.
     In addition to lung cancer risk, the author estimates passive smoking related deaths for
other cancers and for heart disease.  By disease, the total consists of 700 lung  cancer deaths,
11,000 other cancer deaths, and 32,000 deaths due  to heart disease.  For each million of the
total population, the deaths by disease are 13 for lung cancer, 46 for other cancers, and 134
for heart disease.                                     ,        ,

Fong(1982)
     The hazard of ambient cigarette smoke to nonsmokers is compared to the hazard of
primary smoke to smokers in Fong (1982).  Figure 12 provides a summary of his analysis.
Fong assumed the nonsmoker worked in an office for 11 hours per day where 37 percent of
the workers smoke and commuted on an unventilated bus  1 hour each day where 37 percent
of the occupants smoke.  The remaining 12 hours of each day was assumed to be unexposed.
The ratio of the exposure of nonsmokers to the exposure of smokers is represented by the
following equation:
                                R = r1 X r2 x r3 X r4
 where:
            R  =  the ratio of exposure of a nonsmoker to a smoker
            rj  =  the ratio of the density of secondary smoke to primary smoke
            r2  =  the ratio of the times of exposure of the two cases
                                           81
-------
           r3 =  the ratio of toxicities of the two types of smoke
           r4 =  the ratio of the amounts of the ambient smoke (secondary plus side stream)
                 a nonsmoker is exposed to and the primary smoke a smoker is exposed to
The maximum value of the ratio R calculated for this scenario is 1/13.
     Since the effect of smoking in the U.S. has been stated as an excess of deaths of
350,000 per year, the maximum risk of ambient smoke to nonsmokers in the U.S. was
estimated at 50,000 excess deaths per year (based on a smoker to nonsmoker ratio of 37:63).
Fong (1982) assumes that 16 percent of the total  deaths due to smoking are caused by lung
cancer and 4 percent are due to emphysema. The remaining 80 percent are caused by heart
disease and other cancers. The author notes, however, that other factors such as diet
confound the effects of ETS on heart disease and cancers at other sites.  He, therefore,
assumes that only 20 percent of the risk of death  from all causes attributed to ETS exposure
as calculated by the above methodology is the realistic state-of-the-art estimate of the excess
risk of death attributable to ETS. Thus the excess risk of death from ETS exposure is 1/60
that of primary smoke.  This corresponds to an excess of 10,000 deaths per year.

VOLATILE ORGANICS
     This section includes risk estimates of indoor exposure to volatile organic compounds
other than formaldehyde,  which  is addressed separately in the next section.  Three studies  are
included in this review:  Tancrede et al. (1987), Wallace (1986), and McCann et al. (1986).
EPA's Total Exposure Assessment Methodology  (TEAM) Study provides indoor exposure
concentration data used in all three studies.  However, the first two studies use TEAM Study
data exclusively to estimate exposure, whereas the McCann et al. (1986) study uses multiple
literature sources as the basis for determining exposure concentrations  to be used in the risk
analyses.
     A major difference among  the three studies  is the methodology employed to estimate
carcinogenic potency.  Wallace (1986) uses unit risk values developed by EPA's Carcinogen
Assessment Group (CAG). As part of a sensitivity analysis, he compares these unit risk
estimates with those developed by the "Harvard Group" which explicitly incorporates
uncertainty into the estimate of potency.  The potencies estimated by the Harvard Group can
                                         82
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 be several orders of magnitude higher than the CAG estimates. Tancrede et al. (1987) use
 the Harvard Group methodology to estimate carcinogenic potency. Because this methodology
 is different from the methodology employed by GAG, Tancrede et al. estimate potencies for
 52 chemicals, of which only nine have corresponding CAG values.  McCann et al. (1986)
 provide a sensitivity analysis on carcinogenic potency for 24 chemicals, based on four
 methodologies.  Nine of the chemicals have corresponding GAG values.  -Figures 13, 14, and
 15 provide a detailed review of these three studies..                     .,

 Tancrede et al. (1987)                       ,                         —
     This study by Tancrede et al. uses indoor air pollutant exposure data from the Total
 Exposure Assessment Methodology (TEAM) Study conducted in Bayonne and Elizabeth, NJ,
 and Los Angeles, CA, as well as data from a study by Lebret et al. (1984) in the Netherlands
 to predict cancer risk from exposure to 52 organic chemicals.  Median estimates of cancer
 potency were calculated from human data (benzene) or animal  bioassay data (20 chemicals)
 when such information was available.  Cancer potency was estimated from toxicity and  .
 mutagenicity data and other information on biological activity for the remaining 31 chemicals
 using the analogical theoretical methodology developed by Fiering and Wilson (1983). For
 most of the 31 chemicals,  studies of comparative toxicities and activities in promotion and
 co-carcinogenesis experiments of the chemicals and related compounds were used to estimate
 potencies. For a few of these chemicals, potencies were estimated from an  inhalatipn or oral
 LD50 taken from the literature.        ,r        \                      .""".,-..
     The review presented in Figure 13 focuses on Tancrede's risk estimates using exposure
 data from the Los Angeles, CA, TEAM study.  These data are based on personal exposure
 monitoring of 200 people in the Los Angeles area during the winters of 1983 and 1984.
 Although not presented in  Figure 13, the authors present a similar analysis based on the New
Jersey TEAM study and the study by Lebret. et  al.  (1984) of organic chemicals in four Dutch
homes over a six-month period. Figure 13 also presents only those risk estimates for which
the potency factor was based on human or animal bioassay data, and excludes all; risk
estimates for which the potency factor  was based on the  analogical theoretical methodology
since this methodology is not in accord with the EPA Risk Assessment Guidelines (U.S.  -
Environmental Protection Agency,  1987a). Tancrede et  al. 's (1987) analysis of lifetime..
                                          83
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median, mean, and 98th-percentile risk to residents of Los Angeles, CA, for individual
organic chemicals is presented on separate framework pages in Figure 13.
     The median risk estimate (pages 1 to 3 of Figure 13) is based on 'median exposure and
potency estimates,  without incorporating uncertainty, using the following equation:

                                   R = 3 x KT4 ~fi  ~d

where:
              R  =  the median estimate of lifetime risk
              ft  =  the median estimate of carcinogenic potency
              d  =  the median estimate of exposure dose

3 X 10"4 is a dosimetry factor which accounts for a breathing rate of the exposed individual
of 20m3  per day; a 100 percent absorption rate; a 70 kg average body weight; and the
conversion of micrograms to milligrams.
     The mean risk estimates (pages 4 to 6 of Figure 13) incorporate uncertainty information
into the risk calculation by using median  estimates of the potency factor and dose along with
an estimate of a2 in the following way:

                             R =  3 X KT4 ~0 'd exp(a2/2)
where:
              )8,  d and 3 x 10"4 are defined above.

The authors assume that potency and dose are lognormally distributed, and thus risk is
assumed to be lognormally distributed with a variance a2 given by adding the variance of
each of the following four factors:
                                          84
-------
where:
a  =
               =  the standard deviation of the risk estimate
                  the uncertainty in measuring potency in animal experiments
ff  =
                  the uncertainty in extrapolating from animals to humans, assumed to be
                  loge(4.5) = 1.5
                  the uncertainty due to different routes of administration, is assumed to
                  beloge(5)=1.6
           ffz  =  uncertainty (variability) in dose

This results in a mean risk estimate which is exp(a2/2) greater than the median estimate.
     The 98th-percentile risk estimate (pages 7 to 8 of Figure 13) incorporates uncertainty
into the risk calculation in a similar manner using the following equation:
                               = 3 X 10"4 ~p H exp(2.0537o)  j
Thus, the 98th-perceritile risk estimate is exp(2.0537a) greater than the median' estimate.  A
result of this methodology is that estimates with the most uncertainty (largest standard
deviation) in the input parameters have the highest estimated risk.
     Two key assumptions are made in this analysis.  First, exposure is assumed to be
continuous for a 70-year lifetime.  This assumption is based on evidence that individuals
spend 80 to 90 percent of their time indoors.  However, the exposure data used in the
analysis are from overnight personal air samples taken during the winter months  in each of
the three locations.  Thus, a second assumption is that the overnight personal air sample data
for the home are reasonably representative of air concentrations from all indoor micro-
environments,  over long time periods.
     Tancrede et al.  (1987) also provide estimates of the mean annual total risk for the
mixture of contaminants measured in the TEAM Study at each of three locations: Bayonne,
NJ; Elizabeth, NJ; and Los Angeles, CA.  The basic assumption used to derive these
estimates is that the risk from the mixture of organics presented in the analysis is the sum of
the lifetime individual risk posed by each of the chemicals separately.  Thus, risks are
                                          85
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assumed to be additive.  These risk estimates are not presented in Figure 13 since they
include many compounds other than those shown here.

Wallace (1986)
     This study uses EPA's TEAM Study data to construct a rough nationwide risk
assessment for the following six chemicals:   benzene, chloroform (both air and water),
carbon tetrachlbride, trichloroethylene, tetrachloroethylene, and ^ara-dichlorobenzene.  The
TEAM study measured personal air and drinking water exposures to these six chemicals and
approximately 15 other chemicals for 600 residents of New Jersey, North Carolina, North
Dakota, and California between 1981 and 1984.  This sample represented a total population
of ~ 700,000  in seven cities. The author cautions that  although this is the largest personal
                                                                                   t
exposure data set available, its use in providing rough estimates of population risks should be
considered in  light of the many and great uncertainties in these estimates.
     The review presented in Figure 14 focuses on the author's estimates of individual
lifetime risk and population risk for the six chemicals for both metropolitan and
nonmetropolitan residents.  The equation used by Wallace (1986) for calculating the risk of
excess cancer incidence (y^ for an individual from exposure to a carcinogen is the following:
                                          = aX
where:
            a  =
potency of the carcinogen, in units of excess cases per unit concentration
normalized to a 70 kg male with a 70-year lifetime (i.e., unit risk factor)
            X;  =  mean lifetime concentration
 The equation used to estimate aggregate or population risk (Y) where the sample size is N
 and the sum of the weights equals a target population (P) is the following:
                                       Y = Pa wx
                                           86
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where:
           a    =  potency as defined above
           wx .==  the arithmetic mean of the population-weighted exposure distribution

The author notes that these equations ignore differences in individual susceptibility; the
possible multistage nature of cancer; the different effects of early exposure versus late
exposure; and the different actions of initiators, promoters, and carcinogens, among others.
     The potency estimates used by,the author for five of the chemicals were unit risk
estimates [expressed as (/fg/m3)"1] developed by EPA's Carcinogen Assessment Group
(GAG).,  For benzene, the unit risk estimate is based on human epidemiological data and are
a best (maximum likelihood) estimate.  For tetra chloroethylene, trichloroethylene,
chloroform, and carbon tetrachloride, the unit risk estimates are based  on animal studies and
are calculated from the 95-percent upper confidence limit on the potency factor.  The unit
risk estimate for ^ara-dichlorobenzene is nota'CAG estimate but was calculated by the
author based on a National Toxicology Program mouse bioassay.   For  comparative purposes,
the author also presents mean and 95-percent upper confidence limit potency estimates
developed by a group at Harvard. Their methodology incorporates uncertainties of various
types into the calculation,  A large variance exists between the CAG 95-percent upper
confidence limit estimates and the Harvard estimates.
     The author calculates  both lifetime individual risk and population.risk for those living in
U.S. metropolitan areas and nonmetropolitan areas.  The author assumed that the TEAM
Study data for New Jersey  and California are representative exposures  for 178 x  106
metropolitan residents,  and the North Carolina and North Dakota data are representative
exposures of 56 x 106 nonmetropolitan residents.
     In addition to what  is shown in Figure 14, the author presents similar estimates for
outdoor exposures to these  chemicals and comparative risk estimates using the potency factors
developed by the Harvard group.              '
                                          87
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McCann et al. (1986)
     The literature for indoor air concentrations for 140 organic compounds is reviewed in
McCann et al. (1986).  In Figure 15 the risk estimates calculated using both maximum and
mean concentrations combined with the maximum likelihood (MLE) and 95-percent upper
confidence level (UCL) unit risk factors are shown for eight carcinogens.  These eight
chemicals (plus formaldehyde shown in the next section) were included in Figure 15 because
they are considered carcinogens by EPA and have unit risk factors estimated by EPA's
Carcinogen Assessment Group.  The MLE estimate is shown because in most cases it is the
lowest or similar to the lowest unit risk estimate of the four.  Similarly, the 95-percent UCL
is similar to the highest unit risk estimate, in most cases.  However, EPA unit risk estimates
are the highest for several compounds including benzene, dichloromethane,
dimethylnitrosoamine, and vinylidenechloride.  Figure 15 gives the general range of risk
estimates made by the authors.
     The primary sources for the full list of compounds examined in this study were the
published literature and presentations made at the  1984 Indoor Air Quality meetings  in
Sweden. Based on this literature review, the authors performed preliminary assessments of
cancer risks for 24 chemicals.  All concentration data presented by the authors are direct field
measurements in homes and public buildings that the authors judged to reflect everyday
exposure in normal, noncomplaint homes and offices.  They did not include, for example,
concentrations of formaldehyde in UFFI homes or concentrations measured in traditional
occupational settings.  When available, maximum  and median or mean concentration
measurements were recorded by the authors.
     Four  different approaches were employed to  estimate unit risk values for the 24
chemicals.  Maximum likelihood estimates and 95-percent upper confidence level values were
calculated using the multistage model based on the dose-response data given by Gold et al.
(1984, 1986).  The authors do not state which animal study these estimates are based upon
when multiple studies are reported for the same chemical.  Risks were also calculated using
the most potent TD50 estimated by Gold et al. (1984, 1986).  The authors assumed this value
is a point on a linear dose-response curve  and  have divided 0.5 by the TD50 (estimated as the
                      (•>
equivalent dose in ^g/nr) to obtain the risk per unit dose.  When Gold et al.  (1984, 1986)
reported that curves were nonlinear, the authors modified the risk estimates by using a less
                                          88
-------
than (<) or greater than (>) sign.  Also, unit risk factors reported in EPA's Health
Assessment Documents (U.S. Environmental Protection Agency, 1985a-f, 1986) were used.
Nine of the 24 chemicals have EPA unit risk numbers and eight of these chemicals are shown
in Figure 15; formaldehyde is shown separately in Figure 18.

FORMALDEHYDE
Interagency Regulatory Liaison Group (1981)
     The Interagency Regulatory Liaison Group's Task Group on Formaldehyde (1981)
published an extensive summary of studies on the exposure and resulting health effects,
environmental effects, and health risks attributable to formaldehyde. A summary is presented
in Figure 16.  The report provides a detailed breakdown for exposure of subpopulations
including four occupational categories, three consumer categories (e.g., residents of mobile
homes, conventional homes), and ambient air and water.  The risk estimates shown in
Figure 16,  column I have been calculated by the Interagency Regulatory Liaison Group
(1981) with potency  factors that were estimated from an animal bioassay.  No  scaling factors
were used in estimating human potency; therefore, it is difficult to present an accurate     •
relationship between columns G, H, and I in Figure 16.

Tancrede et al. (1987)
     Tancrede et al. (1987) present median, mean, and 98th-percentile individual  risk
estimates attributable to exposure to formaldehyde in the average U.S. home.  The authors
assume formaldehyde exposure follows a lognormal distribution with a median indoor
concentration of 0.05 ppm and a logarithmic standard deviation of 1.8. This is an assumed
value  in agreement with several studies for all U.S. houses.  The authors use a carcinogenic
potency factor derived  by Zeise et al.  (1984) of 6a = 0.11 (mg/kg-day)"1 with a standard
deviation of 0.18. The authors also assume the standard deviation of human response
associated with this potency factor to be 1.5.  This analysis is shown in Figure 17.
                                          89
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McCann et al. (1986)
     Formaldehyde was included in the McCann et al. (1986) study on indoor air
concentrations and preliminary carcinogenic risk assessment for 24 organic compounds. This
study has been previously mentioned in the section on volatile organic compounds.
     The authors used four different approaches to obtain unit risk values.  Maximum
likelihood estimates (MLE) and 95-percent upper confidence level values (UCL) were
estimated from experimental results and tumor data used by Gold et al. (1984, 1986) to
calculate the most potent TD50.  A response factor was also calculated from the TD50  value,
which is assumed to be a point on a linear dose-response curve.  The fourth method (EPA)
was to use the unit risk factor reported in EPA Health Assessment Documents (Environmental
Protection Agency, 1986). The unit risk factors were applied to both average and maximum
concentrations reported in literature to obtain the eight lifetime risk estimates  which are
shown in Figure 18.  Lifetime individual risk using maximum measured concentrations
ranged from a high of 9.0 X  10"2 to a low of  3.1  x  10"3.  (Figure 18, page  1, column I).
Corresponding values for the mean measured concentrations (page 3, column  I) ranged from
a high of 9.5 X  10~3 to a low of 3.7  X 10~6.  Due to the nonlinearity of the  dose-response
curve for formaldehyde, the potency factor estimates  are also dependent on the concentration.
Thus, corresponding values in column H on pages 1 and 3 of Figure 18 are not identical. At
lower concentrations there is a nonlinear change in risk (column I) for MLE and UCL
estimates.  Linearity is assumed in EPA and TD50 potency estimates.
     Additionally, the authors estimated the percentage of population at "high risk" from
formaldehyde exposure. "High risk"  was defined as  lifetime individual risks  of greater than
one in 1000 (10"3).  Populations at risk were calculated based on geometric mean and
geometric standard deviation of concentration data by DeBortoli et al. (1985). Those at high
risk ranged from less than 0.01 percent of the  total U.S. population for the MLE to greater
than 99 percent for TD50. These values have  been included in Figure 18, column K, page 1,
for estimates of carcinogenic risk for  lifetime exposures  to maximum measured
concentrations.  However, no attempt was made to quantify the number of excess cancers
expected.
                                          90
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ASBESTOS
     Several risk assessments for occupational exposures to asbestos have been performed
(Hogari and Hod, 1981; Nicholson et al., 1982; and Nicholson, 1983).  These assessments
focus on occupational exposures that have occurred in the past. The recent analysis by
Mauskopf (1987), which is reviewed  in this monograph," focuses on occupational and
nonoccupational exposures to asbestos from products manufactured under current regulations
for asbestos.and asbestos products. This analysis draws on the work of Nicholson and others
and was viewed as one of the few analyses addressing nonoccupational exposure.  Risk
characterization components are shown in Figure 19.

Mauskopf (1987)
     This study presents an analysis of cancer risk attributable to exposure to asbestos from
products manufactured during the years 1985 to 2000-for nine product categories (shown in
column B of Figure 19).  This analysis takes into account current government regulation of
asbestos and projects cancer risk for the next 100 years attributable to  exposure to  products
manufactured during the 16-year period, 1985 to 2000.  Five occupational and four
nonoccupational exposures identified in the lifecycle of each of the nine product categories
are included in  the analysis:  occupational and nonoccupational exposure during primary
manufacture, secondary manufacture (occupational only), installation, use,  and .
repair/disposal.
     Exposure data used by the author vary in duration from 1 to 30 years among the
categories and therefore cannot be presented directly in Figure 19.  A  weighted yearly
lifetime average exposure was calculated and is presented in Figure 19, column E in units of
fibers per year.' These exposure values must be converted, using the dosimetry factor in
Figure 19, column F, to units applicable to  the risk models used (fibers per ml), normalized
for an occupational work exposure (Figure 19, column G). The risk models used were
developed from occupational studies.  The unit of measure for dose-in the model is in fibers
per ml over a typical work year. Since data from occupational exposure studies indicate that
excess mortality is proportional to both the level and duration of exposure, it is reasonable to
normalize indoor air exposure to an occupational equivalent.  The total exposure in fibers per
year were normalized to an occupationally equivalent concentration, as if that exposure
                                          91
-------
occurred during a typical work year of 40 hours per week, 50 weeks per year with a
                    o
breathing rate of 1 nr/hr.
     Two health effects were evaluated, lung cancer and mesothelioma.  The author used a
life table risk assessment model which utilized actual data on the time, level, and duration of
exposure and which provided estimates of individual lifetime risk (Figure 19, column I) and
expected cancer cases (column K) among total exposed population (column J).  Estimates
were also generated for when in the future those cancer cases would occur and the mean ages
of those persons contracting these diseases as a result of asbestos exposures.  These types of
estimates are not shown in Figure 19.
     Cancer cure rates and annual death rates after diagnosis for the terminal patients were
assumed to be constant across demographic groups, exposure categories,  and products.
Values estimated arid used in the analysis were the following: cure rates of 8 percent and
2 percent for lung cancer and mesothelioma respectively, and, for those dying from the
diseases, annual death rates after diagnosis of 81 percent and 71 percent for lung cancer and"
mesothelioma respectively.  The following dose response factors from Selikoff et al. (1979)
were used:  KL = 1.0 x  10"2 (fibers/milliliter)"1 for the lung cancer relative risk model and
Km = 1.5 x 10"8 for the mesothelioma absolute risk model.  These are not the same as the
response factors shown in Figure 19, column H.  The  calculated response factors shown in
Figure 19, column H represent the average  lifetime risk from asbestos exposure of an
occupationally normalized exposure of  1 fiber/ml. These response factors were
back-calculated by dividing the lifetime individual risk (Figure 19, column I) by the
calculated yearly average lifetime dose  (Figure 19, column G).
     The author conducted a sensitivity analysis using  alternative dose-response factor.
Upper bound estimates were from Seidman  et al. (1979) and obtained from a study  of
                                                                    f\
workers exposed less than a year. Seidman values were KL = 6.8 x 10  and Km  =
5.7  X 10'8.  Values of KL = 3:1 X  10'3 (Hughes and Weif, 1980) and  K^ =  0.7  X  10'9
(Peto, 1980) were used to obtain lower bound estimates.  The author predicted a total of
1,541 cancer cases, and the sensitivity analysis revealed a range of 268 to 8,090 total cases.
                                          92
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            GONCLUSIONS/REGOMMENDATIONS

CONCLUSIONS
• , Few complete quantitative risk characterizations appear in the literature.
   While.many scientific studies were found which deal with individual components of risk
   assessment, only a few  studies combine exposure assessment data with dose-response
   evaluations to'produce quantitative risk characterizations (as defined in the U.S.
   Environmental Protection Agency Risk Assessments Guidelines of 1986).

•  Most of the risk characterization studies which were found and reviewed focus on
   cancer risk.                                     - \   ...-'-
   For radon, formaldehyde, and asbestos, the cancer risk estimates are for lung cancer.  The
   analysis of cancer risk from exposure to asbestos includes mesothelioma as a separate
   health effect in addition to lung cancer. The analysis of cancer risk from exposure to
   specific volatile organics does not specify any particular type of cancer. Three of the ETS
  .analyses address diseases other than cancer including heart and lung, diseases.

•  Few published risk characterizations are strong hi all components of risk assessment.
   Often a rigorous dose-response evaluation has been combined with a very cursory
   exposure assessment (or vice versa).

•  Pollutants may be ranked for comparison by individual lifetime risk or by risk to
   exposed populations.
   The two approaches (Figures 2 and 3) give different results and are best analyzed in
   combination and with regard to weight-of-evidence evaluations.  These summary figures
   provide useful insights into the relative significance of indoor air pollutants such as radon,:
   asbestos, organic compounds, and environmental tobacco smoke.
                                        93
-------
•  A large quantity of data essential to future risk characterizations is emerging from
                                         %
   the current work.
   While few completed risk characterization studies were found, a large number of studies
   containing one or more essential components of risk characterization were identified.

•  The review of risk characterization studies and analyses presented in this work must
   be considered preliminary.
   While considerable effort was expended to identify risk characterization studies (from EPA
   and RTI data bases, technical libraries, and computerized library resources), the authors
   cannot be sure that the universe of all relevant studies has been captured.

RECOMMENDATIONS
•  Continued work should focus on:
   —  refining the systematic review and presentation of the complex components of risk
      assessment.
      This should be done so that critical assumptions and data strengths and weaknesses can
      be further understood by more people.  The Risk Characterization Framework
      (presented in Figure 1 and utilized in Figures 4 through 19) provides a start in that
      direction.

   —  combining existing data on  exposure and dose-response relationships to make new
      risk estimates.
      Such estimates should be based on a methodology arid assumptions designed to provide
      risk estimates that are as comparable as possible between pollutants.  A best currently
      available risk characterization for a particular pollutant could  be developed by
      combining the most scientifically defensible information on hazard identification,
      exposure assessment,  and dose-response relationships.
                                         94
-------
performing sensitivity analyses.
This would provide an understanding of the importance of input risk parameters to the
resulting risk characterizations.  Careful study of the uncertainties in the risk
characterization process will lead to improved risk estimates and greater confidence in
those estimates.
                                     95
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       EPA-600/8-87-031. Available from: NTIS, Springfield, VA; PB87-210720/REB.              ,

U. S. Environmental Protection Agency. (1987b) Radon reference manual.. Washington, DC: Office of Radiation
       Programs; EPA report no. EPA-520/1-87-20.  Available from: NTIS, Springfield,  VA; PB88-196654.

U. S. Environmental Protection Agency. (1987c) The risk  assessment guidelines of 1986.  Washington, DC:
       Office of Health and Environmental Assessment; EPA report no. EPA/600/8-87-045. Available from:
       NTIS, Springfield, VA; PB88-123997/AS.

Wallace, L. A. (1986) Cancer risks from organic chemicals in the home. In: Environmental risk management: is
       analysis useful? Proceedings of an APCA international specialty conference; April; Chicago, IL.
       Pittsburgh, PA: Air Pollution Control Association; publication no. SP-55; pp. 14-24.
                                                      *>'•"     •       .     -  -       '    •
Wells, A. J. (1986) Misclassification as a factor in passive smoking risk. Lancet 8507: 638.
                                                  99
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Wells, A. J. (1988) An estimate of adult mortality in the United States from passive smoking. Environ. Int.  14:
       249-265.

Wigle, D. T.; Collishaw, N. E.; Kirkbride, J.; Mao, Y. (1987) Deaths in Canada from lung cancer due to
       involuntary smoking. Can. Med. Assoc. J. 136: 945-951.

Zeise, L.; Wilson, R.; Crouch, E. (1984) Use of acute toxicity to estimate carcinogenic risk. Risk Anal. 4:
       187-199.
                                                  100
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                          .APPENDIX A
                            GLOSSARY
Absolute risk projection model:  a model which estimates the risk of exposure beyond
the years of observation of a studied population by projecting the average observed
number of excess cancers per unit dose into the future years of risk (U.S.  ,
Environmental Protection Agency,  1987).

Activity:  the mean number of decays per unit time of a radioactive nuclide. Units:
becquerel (Bq), curie (Gi) (National Research Council, 1988).

Alpha energy:  the energy released when an alpha particle emitted during radioactive
decay is halted by collision with a  substance (e.g., lung tissue). The amount of energy
depends on the velocity of the alpha particle, which in turn depends on the source of
radioactive decay (e.g., decay of U-238 versus Ra-226) (U.S. Environmental Protection
Agency,  1987).

Alpha particle: two neutrons and two protons bound as a single particle that is emitted
from the nucleus of certain radioactive isotopes in the process of decay or disintegration
(National Research Council, 1988).

Attached radon decay product:   a radon decay product that is attached to a particle of
dust or other material in air (U.S.  Environmental Protection Agency,  1987).

Bequerel (Bq):  the SI unit of radioactivity equal to  one disintegration per second (U.S.
Environmental Protection Agency, 1987).

Beta particle:  a negatively charged subatomic particle (electron) emitted from a
nucleus during some types of radioactive decay (U.S. Environmental Protection
Agency, 1987).

 Cohort:  a large homogeneous group of people tested in epidemiological or
 socioeconomic studies.  EPA's lung cancer estimates are based on calculations  for a
 cohort of 100,000 people (U.S. Environmental Protection Agency,  1987).

 Constant-relative-risk model: a risk model which assumes that, after a certain time,
 the ratio of the risk at a specific dose to the risk in the absence of the dose does not
 change with time (National Research Council, 1988).

 Cumulative working level month (CWLM): a unit of cumulative radon exposure.
 the sum of lifetime exposure to radon working levels in total working level months
 (U.S. Environmental Protection Agency, 1987).
                                     101
-------
 Curie (Ci):  a unit of radioactivity, defined as that quantity of any radioactive nuclide
 which spontaneously undergoes 3.7 x 1010 disintegrations per second. One gram of
 radium-226 has an activity of one curie (U.S. Environmental Protection Agency, 1987).

 Decay series: the consecutive members of a family of radioactive isotopes formed by
 sequential radioactive decay.  A complete series commences with a long-lived parent
 such as U-238 and ends with a stable element such as Pb-206 (U.S. Environmental
 Protection Agency, 1987).
Electron volt (eV):  a unit of energy = 1.6 x 10"12 ergs =  1.6 x 10~19 J;  1 eV is
equivalent to the energy gained by an electron in passing through a potential  difference
of 1 V; 1 KeV =  1,000 eV;  1 MeV = 1,000,000 eV (National Research Council,
1988).

Equilibrium factor:  an adjustment used in converting from picocuries per liter (pCi/L)
to working-level concentration (WL), which takes into account the possible absence of
radioactive equilibrium between radon and its decay products  (U.S. Environmental
Protection Agency, 1987).

Equilibrium fraction: ratio of the concentration of radon decay products to
concentration of radon in the  same sample of air at radioactive equilibrium.  Typical
measured values range from 0.3 to 0.7, with an average of about 0.5 (U.S.
Environmental Protection Agency, 1987).

Gamma ray:  short-wavelength electromagnetic radiation of nuclear origin (range of
energy, 10 keV to 9 MeV)  (National Research Council,  1988).

Ionizing radiation:  subatomic particles or photons that have  sufficient energy to
produce ionization directly in  their passage through a substance (U.S. Environmental
Protection Agency, 1987).

Latent period:  the minimum period of time between exposure and expression of the
disease.  After exposure to a dose of radiation,  there is delay  of several years (the latent
period) before any cancers are seen (National Research Council, 1988).

Lifetime risk:  the lifetime probability of contracting a specific disease (National
Research Council,  1988).

Lifetime risk ratio:  the ratio of the lifetime risk (Rg) of an exposed person to the
lifetime risk of an unexposed person (R0).  This number minus 1 is the proportional
increased risk associated with  exposure (Re/R0 - 1) (National Research Council,  1988).

Linear dose model:  this model postulates that  the excess risk is linearly proportional  to
the dose (National Research Council, 1988).
                                    102
-------
Linear energy transfer (LET):  average amount of energy lost per unit track length
(National Research Council, 1988).

Kcocuries per'liter (PCi/L):  a unit of measurement of activity concentration.  One
picocurie per liter is equal to 10"12 curies per liter (U.S. Environmental Protection
Agency, 1987).

Progeny:  the decay products resulting after a series  of radioactive decays. Progeny can
also be radioactive, and the chain continues until a stable nuclide is formed.  Radon
progeny of primary concern to the public health community are 218-Pp, 214-Pb,
214-Bi, and  214-Po (National Research Council, 1988).

Radiation dose: the total amount of ionizing  radiation absorbed by material pr tissues,
in the sense  of absorbed dose (expressed in rads), exposure (expressed in roentgens), or
dose equivalent (expressed in rems) (U.S. Environmental Protection Agency, 1987).

Radioactive equilibrium:  a state in which the rate of formation of atoms by decay of a
parent radioactive isotope is equal to its rate of disintegration by radioactive decay, so
that the activity of the parent and the decay product assume a constant proportion.  This
proportion is equal to one, if the parent has only one mode of radioactive decay.
Because radon decay products tend to attach readily to surfaces, equilibrium between
radon and its decay products is seldom reached/Using the average value of the
equilibrium  fraction (0.5), a ratio of about 200 pCi/L of radon to 1 WL of radon decay
products is fairly typical for residences (U.S. Environmental Protection Agency, 1987).

Radioactive secular equilibrium:  ideal steady state situation in which radon decay
products would be formed and would decay at the same rate. Under such ideal
conditions, the ratio of concentrations  (or activities) of radon to decay products would
be such that 100 pCi/L of radon  and 1 WL of decay  products would be present in the
same sample of air. Secular equilibrium is  never achieved because some of each of the
decay products is removed from the, air (due to attachment to walls, floors, etc.) before
undergoing decay (U.S. Environmental Protection Agency, 1987; National Research
Council,  1988).                                            ?      -..-.,;.•

Relative risk:  the ratio of the rate of disease  in exposed to unexposed populations
(U.S. Environmental Protection Agency, 1987).

Relative risk projection model: a model  which estimates the  risk of exposure beyond
the years of observation of a study's population by projecting the currently observed
percentage increase in cancer risk per unit  dose into the future  years (U.S.
Environmental Protection Agency, 1987).
                                     103
-------
Risk coefficient: the increase in the annual incidence or mortality rate per unit dose:
(1) absolute risk coefficient is the observed, minus the expected number of cases per
person-year at risk for a unit dose;  (2) the relative-risk coefficient is the fractional
increase in the baseline incidence or mortality rate for a unit dose (National Research
Council, 1988).

Risk estimate:  the number of cases (or deaths) that are projected to occur in a
specified exposed population per unit dose for a defined exposure regime and expression
period—for radon, the number of cases per person cumulative working-lever month
(National Research Council,  1988).             ;

Synergistic model: a form of cancer causality whereby two Or more carcinogens (for
example, radon and tobacco  smoke) act synergistically to cause cancer with a greater
probability than if each were acting alone (U.S. Environmental Protection Agency,
1987).

Threshold hypothesis:  the assumption that no radiation injury occurs below a specified
dose (National Research Council, 1988).

Time-since-exposure (TSE)  model:  a model in which the relative or absolute risk is
not constant but varies with the time after exposure (National Research Council, 1988).

Unattached radon decay product:  a radon decay product that is not electrostatically
attached to dust or particles in the air. Capable of attaching to lung tissue if inhaled
(U.S. Environmental Protection Agency, 1987).

Working level (WL): any combination of short-lived radon progeny in  1 liter of air
that will result in the ultimate emission 1.3 X 105 MeV of potential alpha energy.  This
number was chosen because it is approximately the alpha energy released from the
decay of progeny in equilibrium with 100 picocuries of 222Ra (National Research
Council, 1988).

Working-level month (WLM): exposure resulting from inhalation of air with a
concentration of 1 working level of radon progeny for 170 working hours (National'
Research Council, 1988).
                                    104
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                                APPENDIXB

      Concentration Units and Conversion Factors for Radon and Radon Progeny
     The quantity of radioactive species present in a given volume of air is typically specified
by activity rather than by mass.  The standard international (SI) unit for activity is the
Becquerel (Bq), which is defined as 1 radioactive decay per second,; The Curie (Ci) and
picocurie (pCi) represent 3.7 X  1010 Bq and 0.037 Bq, respectively, and are often used in
the U.S. to express quantity of radon.         .                       ;         ,
     Concentration of radon in air is typically expressed in Becquerels per cubic meter
(Bq/m3) or in picocuries per liter (pCi/L).  Concentration of radon progeny (218Po, 214Pb,
214Bi, and 214Po) in  air is typically expressed in working levels (WL).  A working level is
any combination of radon progeny which has the potential to release 1.3 x  iO  mega-electron
volts of alpha energy per liter of air (MeV/L). Concentrations of radon and radon progeny in
indoor air are, of course, related,  but the relationship is dependent upon a number of factors.
Such factors include  the rate at which progeny are removed from the air (by attachment to
aerosols, dust particles,  walls, ceilings, and floors).  In indoor environments,'one working
level of progeny corresponds to about  150-300 pCi/L of radon, with 200 pCi/L representing
about the average conversion factor for radon in homes.                ,
     Radon itself, because of its behavior as essentially a chemically inert gas .with no
affinity for attachment to other materials (including aerosols, dust particles,  and lung tissue),
does not appear to pose a serious  threat to human health.  Radon progeny, however, attach
                           :          •' • *          -          '  •   '"•-"'-     - -    "
readily to other materials, including lung tissues.  Because of this, epidemiological studies are
usually concerned with exposure to radon progeny rather than with exposure to radon.   '
Exposure to radon progeny is typically measured in working-level months (WLMs), where
one WLM is equivalent to 170 hours of exposure to one WL of progeny.
                                          105
-------
                            TABLE OF CONVERSIONS
                    Description
                                                              Conversions
Activity:  mean number of radioactive decays per unit
time. Typically expressed in Becquerels (Bq), Curies
(Ci), or picocuries (pCi).
        1 Bq = 1 decay/sec

        1 Ci = 3.7 X 1010 Bq

        1 pCi = 0.037 Bq
Concentration:  activity per unit volume. Typically
expressed in Becquerels per cubic meter (Bq/m3) or
picocuries per liter (pCi/L) for radon and in working
levels (WL) for radon progeny. A WL is defined in
terms of potential alpha energy (from radon progeny).
per liter of air.  Some assumptions regarding relative
concentrations of radon and progeny are required for
conversions between pCi/L (or Bq/m3) and WL.
Progeny concentration may also be expressed in
energy units of mega-electron volts per liter (MeV/L)
or joules per cubic meter (J/m3) with proper
conversion  of units.
        1 pCi/L = 37 Bq/m3

        1 WL = 200 pCi/L
        (typically, in a home)

        1 WL =  3.7 X 103Bq/m3
        (at radioactive equil.)

        1 WL = 1.3 X 105 MeV/L

        1 WL = 2.1 x 10-5J/m3
Exposure:  exposure concentration multiplied by the
length of time exposed.  Exposure to radon progeny is
typically expressed in working-level months (WLM).
A WLM is equivalent to exposure to a concentration
of 1 WL for 170 hours.
       1 WLM = 170 WL-h

       1 WLM = 3.5  X 10"3
       J-h/m3
                                       106
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