oEPA
Office of Radiation & Indoor Air EPA 402-R-05-007
Radiation Protection Division (6608J) August 2007
Washington, DC 20460
Technologically Enhanced
Naturally Occurring
Radioactive Materials
From Uranium Mining
Volume 2:
Investigation of Potential
Health, Geographic, and
Environmental Issues of
Abandoned Uranium Mines
TECHMQLQGICALLrEWAMCED KKTUHftU.? DCCUHHHG HMJDftCTIVt MMt:HIM.5
TENOR
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Table of Contents
Introduction v
1.0 Major Studies Supporting This Scoping Risk Analysis 1-1
1.1 1983 EPA Report to Congress 1-1
1.2 1989 EPA Study in Support of NESHAPs 1-6
1.3 Uranium Mines on the National Priorities List 1-6
1.4 DOE Report on Costs of U. S. Uranium Mine Environmental Restoration 1-9
2.0 Geographic Analysis on the Location of Uranium Mines 2-1
2.1 Errors in Mine Locations 2-3
2.2 Number of People Potentially Exposed to Uranium Mine Wastes 2-3
2.3 Mines by Watershed 2-4
3.0 Cancer Risks from On-Site Exposure 3-1
3.1 Potential Scenarios and Exposure Pathways for the General Public 3-1
3.2 Methodology Used in This Analysis 3-3
3.3 Recreational Scenario Risk Calculations 3-5
3.4 Other Recreational Use Scenarios 3-14
3.5 Metals in Uranium Mines 3-14
3.6 Migration of Uranium Waste into Groundwater 3-21
3.7 Mobility of Uranium and Radium through Groundwater 3-22
3.8 Consideration of Multiple Exposure Pathways 3-28
4.0 Risk from Uranium Mining Waste in Building Materials 4-1
4.1 Building Materials Analysis 4-4
4.2 Risk of Exposure of On-site Residents to Uranium Mining Waste 4-9
5.0 Potential Ecological Impacts from Uranium Mines 5-1
5.1 Other Metals 5-3
6.0 Uncertainties 6-1
7.0 Conclusions 7-1
7.1 Summary 7-1
7.2 Potential Considerations for Site Prioritization 7-2
8.0 Bibliography 8-1
Appendices
Appendix I Swimming Risk AI-1
Appendix II Calculation of Slope Factors for Naturally Occurring Radionuclides AII-1
Appendix III Occupational and Public Risks Associated with In-Situ Leaching AIII-1
Appendix IV Risks Associated with Conventional Uranium Milling Operations AIV-1
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List of Tables
1.1 Sources of Contamination at Uranium Mines 1-2
1.2 Estimated Lifetime Fatal Cancer Risks from 1 Year of Exposure to Airborne
Uranium Mine Emissions 1-4
1.3 Estimated Lifetime Fatal Cancer Risks from Lifetime Exposure to Airborne
Uranium Mine Emissions 1-4
1.4 Annual Exposure from Radon Decay Product Emissions from Model Uranium
Mines 1-5
1.5 Estimated Individual Lifetime Fatal Cancer Risks for Various Exposures to Radon
Decay Products 1-6
1.6 Potential Cancer Risks from the White King/Lucky Lass and Midnite Mine Sites 1-9
2.1 Mine Sizes for Four-Corners States 2-1
2.2 Comparison of Data Compiled from Uranium Mine Records 2-3
2.3 Estimated Number of People within 1 Mile (1.6 km) and 5 Miles (8 km) of a
Recorded Mine 2-5
2.4 Number of Mines on Federal Lands in Selected States 2-7
3.1 Selected Radionuclide Toxicity and Preliminary Remediation Goals for Superfund
for Comparison with the SSG Pathway-Specific Approach 3-5
3.2 Soil Screening Levels for External Exposure to Ra-226 3-6
3.3 Soil Screening Levels for External Exposure to Th-232 3-7
3.4 Soil Screening Levels for External Exposure to Natural Uranium 3-8
3.5 Soil Screening Levels for Ingestion of Ra-226 in Soil 3-10
3.6 Soil Screening Levels for Ingestion of Th-232 in Soil 3-10
3.7 Soil Screening Levels for Ingestion of Natural Uranium in Soil 3-11
3.8 Target Lifetime Cancer Risk for Ingestion of Arsenic by Children up to
6 Years Old 3-16
3.9 Radionuclide Maximum Contaminant Levels for Public Water Supplies 3-17
3.10 Radionuclide Mortality and Morbidity Risk Coefficients 3-18
3.11 Lifetime Risks Estimated from Drinking Unremediated Yazzie-312 Mine Pit Water.. 3-20
3.12 Look-up Table for Estimated Range of Kd Values for Uranium Based on pH 3-24
3.13 Soil Screening Values for Uranium as a Function of Kd 3-24
3.14 Relationship Between pH Levels and Strontium Mobility as a Surrogate for
Radium 3-26
3.15 Soil Screening Values for Radium as aFunction of Kd 3-26
3.16 Multi-pathway Soil Screening Levels for Ra-226 3-29
3.17 Multi-pathway Soil Screening Levels for Th-232 3-29
3.18 Multi-pathway Soil Screening Levels for Natural Uranium 3-29
11
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List of Tables (Continued)
4.1 Doses from 30 Years of External Exposure toU-238 in aNavajo Hogan 4-6
4.2 Doses from 30 Years of External Exposure toRa-226 in aNavajo Hogan 4-8
4.3 Estimated Lifetime Risk of Fatal Lung Cancer from Living on Contaminated Land ... 4-10
5.1 Biota Concentration Guides (BCGs) for Water and Sediment for Evaluation of an
Aquatic System 5-2
5.2 Biota Concentration Guides (BCGs) for Water and Soil for Evaluation of a
Terrestrial System 5-2
5.3 Mineral Commodities with Uranium Associations 5-4
in
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List of Figures
1.1 Aerial Image of Midnite Mine, Washington State 1-7
1.2 Aerial Image of White King and Lucky Lass Mines, Oregon 1-8
2.1 Mines and Other Locations with Uranium in the Western U.S 2-2
2.2 Uranium Locations from EPA Database and Federal Lands 2-6
2.3 Western Uranium Mine Density by 8 Digit Hydrologic Unit Code 2-7
2.4 Uranium Locations in Southwest Colorado and Southeast Utah 2-8
3.1 Uranium Mine Pit Lake 3-2
3.2 External Exposure - Relationship between Exposure Frequency, Radium
Concentration, and Target Lifetime Cancer Risk 3-7
3.3 External Exposure - Relationship between Exposure Frequency, Thorium
Concentration, and Target Lifetime Cancer Risk 3-8
3.4 External Exposure - Relationship between Exposure Frequency, Uranium
Concentration, and Target Lifetime Cancer Risk 3-9
3.5 Relationship between Exposure Frequency, Radium Concentration, and Target
Lifetime Cancer Risk from Soil Ingestion 3-10
3.6 Relationship between Exposure Frequency, Thorium Concentration, and Target
Lifetime Cancer Risk from Soil Ingestion 3-11
3.7 Relationship between Exposure Frequency, Uranium Concentration, and Target
Lifetime Cancer Risk from Soil Ingestion 3-12
3.8 Cancer Risks from Lifetime and Recreational Exposures to Radium in Drinking
Water: 70 Years, 365 Days/Year & 10 Years, 14 Days/Year Exposure 3-19
3.9 Cancer Risks from Lifetime and Recreational Exposures to Gross Alpha in
Drinking Water: 70 Years, 365 Days/Year & 10 Years, 14 Days/Year Exposure 3-19
3.10 Cancer Risks from Lifetime and Recreational Exposures to Uranium in Drinking
Water: 70 Years, 365 Days/Year & 10 Years, 14 Days/Year Exposure 3-20
3.11 Average Precipitation (inches/year) for the Western United States 3-27
3.12 Multi-pathway Soil Screening Levels forRa-226 3-30
3.13 Multi-pathway Soil Screening Levels for Th-232 3-30
3.14 Multi-pathway Soil Screening Levels for U-238 3-31
4.1 Locations of Building Gamma Anomalies Due to Uranium Ore from 1973 EPA-
AEC Study 4-2
4.2 Monument Valley Navajo Hogan 4-3
4.3 Navajo Home in Proximity to Uranium Mine 4-4
4.4 Uranium Mine Debris Pile 4-4
4.5 Navajo Hogan Building Model 4-6
4.6 Doses from 30 Years of External Exposure to U-23 8 in a Navajo Hogan 4-7
4.7 Doses from 30 Years of External Exposure toRa-226 in a Navajo Hogan 4-8
IV
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Introduction
Uranium is a common element in nature, and has been used for centuries as a coloring agent in
decorative glass and ceramics. Today, uranium has uses that range from metal alloys to aircraft
counterweights. The most significant modern uses of uranium, however, have been for national
defense and electric power generation. The advent of nuclear weapons and nuclear power in the
United States resulted in a full-blown exploration and mining boom, starting immediately after
World War II and making uranium the most important commodity in the mining industry. The
greatest period of uranium production spanned from approximately 1948 to the early 1980s (U.S.
DOE/EIA 1992). Through 2005, the industry had generated over 420,000 metric tons (MTs) of
uranium to foster U.S. dominance in nuclear weapons technology, and later to feed the growing
number of commercial power plants utilizing the enormous energy contained in the uranium
nucleus (U.S. DOE/EIA 2003a, 2003b, 2006).
Another legacy of uranium exploration, mining, and ore processing was the creation of
unreclaimed land workings wherever the uranium concentration in rock was either found or
thought to be economically viable. Thousands of miners and prospectors, as well as large
mining companies, searched the United States in search of veins, lenses, sedimentary deposits,
and breccia pipes concentrating the valuable metal, echoing the California gold rush 100 years
earlier. In many instances, they left behind unreclaimed and exposed wastes elevated in
naturally occurring radioactive materials (uranium and its radioactive decay progeny), exposing
people and the environment to its hazards.
In this report, Naturally Occurring Radioactive Material (NORM) is defined as: Materials
which may contain any of the primordial radionuclides or radioactive elements as they
occur in nature, such as radium, uranium, thorium, potassium, and their radioactive decay
products, that are undisturbed as a result of human activities. Radiation levels presented by
NORM are generally referred to as a component of "natural background radiation."
The term Technologically Enhanced Naturally Occurring Radioactive Material (TENORM) is
defined as: Naturally occurring radioactive materials that have been concentrated or
exposed to the accessible environment as a result of human activities such as
manufacturing, mineral extraction, or water processing. "Technologically Enhanced" means
that the radiological, physical, and chemical properties of the radioactive material have been
altered by having been processed (beneficiated) or disturbed in a way that increases the potential
for human and/or environmental exposures. This definition differs somewhat from other
definitions provided by the National Academy of Sciences (NAS 1999a) and the Conference of
Radiation Control Protection Directors (CRCPD 2004) in that it further amplifies the need to
include materials which have not been modified by human activities, yet have been disturbed in
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such ways that they can be misused by humans, or affect the environment1; it does not include a
reference to Atomic Energy Act materials, as the definitions are changing (see Volume I and its
Appendix VI). Uranium TENORM includes the succession of radioactive decay progeny of the
parent uranium.
Under the Atomic Energy Act, the U.S. Nuclear Regulatory Commission (NRC) regulates
operations that produce and concentrate uranium and thorium. In accordance with terminology
of the Act, the NRC has defined in 10 CFR 40.4 "source materials" as (1) uranium or thorium,
or any combination thereof, in any physical or chemical form, or (2) ores which contain by
weight one-twentieth of one percent (0.05%) or more of: (i) uranium, (ii) thorium or (iii)
any combination thereof. Source material does not include special nuclear material. It also
defines the "by-product materials" (wastes) of those operations as tailings or wastes produced
by the extraction or concentration of uranium or thorium from any ore processed
primarily for its source material content, including discrete surface wastes resulting from
uranium solution extraction processes. Byproduct materials are also regulated by the NRC.
Underground ore bodies depleted by such solution extraction operations do not constitute
"byproduct material" within this definition. Wastes from conventional uranium mining (both
surface and underground) are not subject to NRC regulation, but are considered to be TENORM,
and thus subject to U.S. Environmental Protection Agency (EPA) and State agency oversight.
Under the Energy Policy Act of 2005, the Atomic Energy Act was amended to place additional
discrete (highly radioactive in small, defined volumes) sources of TENORM which had the
potential to pose a threat to public health and safety or the common defense and security under
NRC jurisdiction. The definition of byproduct materials was further modified to include discrete
sources of radium-226, any material made radioactive by use of a particle accelerator for use in a
commercial, medical or research activity, or materials which might pose a similar threat to public
health and safety or the common defense and security. Specific requirements were provided for
determining the appropriate waste disposal methods for these materials. The NRC regulatory
definitions of byproduct materials to accommodate these amendments are expected to be
finalized in the summer of 2007, to reflect the recent amendments as of this writing. These
products and wastes are not the subject of this report.
The U.S. Environmental Protection Agency (EPA) has previously issued reports on the uranium
mining industry in response to congressional mandates and programmatic needs. In 1983, EPA
1 The National Academy of Sciences (NAS 1999a) defined TENORM as "... any naturally occurring
radioactive materials not subject to regulation under the Atomic Energy Act whose radionuclide concentrations or
potential for human exposure have been increased above levels encountered in the natural state by human activities."
The International Atomic Energy Agency (2003), although referring to this class of wastes and products as
"NORMs", defined them as encompassing "all naturally occurring radioactive materials where human activities
have increased the potential for exposure in comparison with the unaltered situation. Concentrations of
radionuclides (i.e. TE-NORM) may or may not have been increased." Alternatively, the Conference of Radiation
Control Program Directors (CRCPD 2004) has defined them as a naturally occurring radioactive material whose
radionuclide concentrations are increased by or as a result of past or present human practices. TENORM does not
include background radiation or the natural radioactivity of rocks or soils. TENORM does not include "source
material" or "byproduct material" as both are defined in the Atomic Energy Act of 1954, as amended (AEA 42 USC
§2011 et seq.) and relevant regulations implemented by the NRC. EPA believes the definition should include
materials which were disturbed, but not further concentrated by human activities, so that the full scope of hazards
from TENORM materials can be considered.
VI
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published its Report to Congress on the Potential Health and Environmental Hazards of
Uranium Mine Wastes (U.S. EPA 1983a, b, c), as required by the Uranium Mill Tailings
Radiation Control Act of 1978. This study provided an important overview of the characteristics
and generation of uranium mining TENORM wastes during a period when the uranium mining
industry was still near its production peak. A subsequent 1985 Report to Congress on Wastes
from the Extraction and Beneficiation of Metallic Ores, Phosphate Rock, Asbestos, Overburden
from Uranium Mining, and Oil Shale (U.S. EPA 1985), carried out pursuant to requirements of
the Resource Conservation and Recovery Act of 1976 (RCRA), as amended, provided additional
risk information and characterization of uranium mining waste. In 1995, EPA issued the
Technical Resource Document Extraction and Beneficiation of Ores and Minerals: Uranium as a
technical update to provide a means of evaluating wastes that were exempt from or subject to
regulation under RCRA (U.S. EPA 1995).
During the period 1989 to 1993, EPA worked on a draft scoping report (SC&A 1993) which
compiled information on TENORM in several industries, including uranium mining. A
preliminary risk assessment was also developed for certain public and occupational exposure
scenarios involving the known radiation levels in those industries. Comments received on the
draft from industry, as well as EPA's Science Advisory Board (SAB) (U.S. EPA 1994), resulted
in further revisions of the scoping draft, though it was ultimately decided that a final report
would not be issued.
Following a review of EPA's guidance for TENORM by the National Academy of Sciences,
EPA's response to the NAS study, and discussions with EPA's Science Advisory Board, EPA's
Radiation Protection Division decided that a further review of the current hazards associated
with uranium mining TENORM was warranted. The SAB (U.S. EPA 200la) agreed with EPA's
intent to make TENORM documents useful to a broad audience, but also recommended that the
whole life cycle of a TENORM source—in this case uranium extraction—be considered beyond
regulatory or inter-agency considerations, and that the impacts of non-radiological contaminants
also be examined in the Agency's technical reports. In addition to most sources of TENORM,
EPA is responsible for setting environmental standards under the Uranium Mill Tailings
Radiation Control Act, cleaning up hazardous waste sites that include some former uranium
mines, and assisting Native Americans, including assisting in environmental reviews of proposed
in situ leach (ISL) facilities. While this report focuses on the impacts associated with
conventional surface and underground uranium mines, it provides limited background materials,
in appendicies, on risks associated with uranium milling and ISL operations and wastes
generated by those processes, even though they may not be considered TENORM by virtue of
their regulation by the NRC and its Agreement States under the Atomic Energy Act and its
amendments.
This is the second of two reports on uranium mining TENORM. The first report,
Technologically Enhanced Naturally Occurring Radioactive Materials from Uranium Mining,
Volume 1: Mining and Reclamation Background(U.S. EPA2006a), provides background
information on the occurrence of uranium, mining techniques, and reclamation of uranium
mines. This report investigates the potential radiogenic cancer risks from abandoned uranium
mines and evaluates which may pose the greatest hazards to members of the public and to the
environment. The intent of this report is to identify who may be most likely to be exposed to
wastes at small abandoned uranium mines, and where the greatest risks may lie. The specific
Vll
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wastes of EPA concern from this report and study are from abandoned conventional open-pit and
underground uranium mines, and include overburden, unreclaimed sub-economic ores (protore),
waste rock, core hole and drill cuttings, and mine and pit (or pit lake) water. All are described in
Volume I of this study. In addition, EPA has compiled and published a uranium location
database (U.S. EPA 2006b).
A first draft of this report underwent an outside peer review following the Agency's peer review
process. Using the comments obtained, the report has been updated and revised. Appendices
have been added to this version of the report providing references and information on the risks
associated with uranium mill operations and ISL operations. While some of the thousands of
conventional open surface and underground uranium mines in the United States have been
reclaimed, many have not. Any mine may pose such hazards as open shafts and unstable
supports (rock and wood), and contain gases, such as carbon dioxide and methane, that displace
oxygen and could lead to asphyxiation. In addition to the immediate physical threats that
abandoned mines may pose, exposure to radiation from uranium and radium and other
contaminants in abandoned mine waste can increase a person's risk of cancer.
People are exposed to naturally occurring radioactive materials in soils, as well as natural
occurrences of uranium in rock outcrops. However, the primary focus in this report is on
exposures to those naturally occurring radioactive materials that have been enhanced by human
activities. In examining the radiological risks due to mining, the focus is on those concentrations
above natural background, as recommended in the EPA Abandoned Mine Site Characterization
and Cleanup Handbook (U.S. EPA 2000a), with emphasis on uranium and radium. Abandoned
conventional uranium mines may also contain other hazardous contaminants, such as metals.
For example, the carcinogen arsenic may be a problem at some uranium mines, contributing to
increased risks.
This scoping report describes in Chapter 1 several previous studies supporting the risk analysis,
while Chapter 2 provides a geographic location analysis of uranium mines in the western United
States. Chapter 3 discusses potential scenarios and exposure pathways for the general public to
hazards from uranium mines, describes the methodologies used in the analysis, and assesses
cancer risks posed by human exposure to the various hazards from the mines. Chapter 4
examines the use of uranium risks in building materials, and Chapter 5 briefly discusses the
potential for ecological impacts from the mines. Uncertainties and conclusions are presented in
Chapters 6 and 7.
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1.0 MAJOR STUDIES SUPPORTING THIS SCOPING RISK
ANALYSIS
The most important period of past U.S. uranium production spanned from approximately 1948
to the early 1980s (U.S. DOE/EIA 1992). Through 2005 the industry had generated over
420,000 metric tons (MTs) of uranium for nuclear weapons and commercial power plants (U.S.
DOE/EIA 2003a, 2003b, 2006). Uranium exploration, mining, and ore processing left a legacy
of unreclaimed land workings wherever the uranium concentration in rock was either found or
thought to be economically viable. This report investigates some potential health, geographic,
and environmental issues of abandoned uranium mines.
The major studies supporting this scoping analysis include EPA's 1983 Report to Congress on
the Potential Health and Environmental Hazards of Uranium Mine Wastes (U.S. EPA 1983a, b,
c) and EPA's risk assessments for underground and surface uranium mines for Clean Air Act
requirements (U.S. EPA 1989a). Other analyses considered include a report of two uranium
mines on the Superfund National Priorities List (U.S. EPA 2001b) and a U.S. Department of
Energy report (U.S. DOE/EIA 2000). These studies are discussed in this chapter.
1.1 1983 EPA Report to Congress
The Uranium Mill Tailings Radiation Control Act of 1978 directed EPA to conduct a study on
"the location and potential health, safety and environmental hazards of uranium mine wastes,"
and to provide "recommendations, if any, for a program to eliminate these hazards." When
EPA published its 1983 Report to Congress (U.S. EPA 1983a, b, c) (hereafter referred as the
1983 EPA report or study), there were about 340 active uranium mines in the United States. At
the end of 2002, there were no active conventional uranium mining operations in the United
States, and only two active operations using the in situ leaching process (U.S. DOE/EIA 2003a).
However, with an increase in the price of uranium since 2004, additional conventional mines
have begun production or will be coming on line in the near future, and some suspended mine
operations have recommenced. As part of the 1983 study, EPA also made observations at a
number of active and inactive uranium mine sites, collected soil and water samples, and took
some external gamma and radon flux measurements at sites in Colorado, New Mexico, Texas,
and Wyoming.
1.1.1 Sources and Pathways Modeled
In the 1983 report, EPA used the information discussed above to develop models for large and
small mines, including an inactive surface mine hypothetically located in Wyoming and an
inactive underground mine hypothetically located in New Mexico (U.S. EPA 1983b). From
these model mines, which were classified as an average mine or a large mine, EPA estimated
the health effects to populations within 50 miles (80 km) of each mine and on a hypothetical
most exposed individual living about 1 mile from the center of a mine. The pathways
considered were as follows:
• Breathing air containing windblown dust and radon decay products
• Drinking water containing uranium and its decay products
1-1
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• Eating food contaminated by either air or water
• Living in homes on land covered by mine wastes (U.S. EPA 1983b)
With the exception of the last pathway, the focus of the report was on estimating risks to people
who were off site. The home pathway was not explicitly modeled like the other pathways, but
used estimates of indoor radon as a function of radium in the soil. While the 1983 report
produced many analyses, some issues were not explicitly addressed, including the following:
• Drinking groundwater and surface water near a mine. This pathway was considered and
included for the regional population, but was not included for the most exposed
individual due to lack of information on radionuclides in potable water.
• Individuals spending time on mine sites.
• Using mine waste material for buildings.
In its 1983 Report to Congress, EPA identified the sources modeled and those considered, but
not modeled, due to a lack of information (Table 1.1). For groundwater, the report noted that
uranium mines may pose a problem, but the authors did not have enough information to
consider it. The report also noted that spending time at the mine sites and using waste materials
in the buildings would be a health hazard, but did not quantitatively address the issues.
Table 1-1. Sources of Contamination at Uranium Mines
In its 1983 Report to Congress, EPA identified the sources modeled (M)
and those considered (C), but not modeled, due to a lack of information.
Sources of Contamination
Waste Rock (Overburden) Pile
Wind-suspended dust
Radon-222 emanation
Precipitation runoff
Sub-Ore Pile
Wind-suspended dust
Radon-222 emanation
Precipitation runoff
Ore Stockpile
Wind-suspended dust
Radon-222 emanation
Precipitation runoff
Abandoned Mine Area Surfaces
Radon-222 emanation
Mining Activities
Dusts
Combustion products
Radon-222
Wastewater
Surface discharge
Seepage
Underground
Active
M
M
C
M
M
C
M
M
C
M
M
M
M
M
C
Mines
Inactive
M
M
C
M
M
C
M
M
C
M
NA
NA
NA
NA
C
Surface
Active
M
M
C
M
M
C
M
M
C
M
M
M
M
M
C
Mines
Inactive
M
M
C
M
M
C
M
M
C
M
NA
NA
NA
NA
C
Note: NA = not applicable.
Source: USEPA 1983b, Table 2.
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7.7.2 7953 EPA Study Findings
Using the risk methodology of the time (AIRDOS-EPA, DARTAB, and RADRISK), the study
estimated that a large active underground mine posed an increased chance of a fatal lung cancer
to an individual of 2 x 10"3, primarily from breathing radon decay products, and that risks from
other types of uranium mines were somewhat lower. Releases to surface water from an average
underground mine one mile from an individual's home were estimated to increase his or her
lifetime cancer risk by 1 x 10"3, and that one additional cancer in several hundred years might
occur in nearby populations from the normal operational releases from a mine. Although the
study did not address the health effects of contaminated shallow aquifers around active or
inactive mines, it recommended that they be evaluated.
For inactive mines, the study noted that radionuclide airborne emissions were smaller than
for active mines, with the risks coming from radon emanating from unsealed mine vents,
portals, and residual waste piles. The estimates of risks from radon emissions from inactive
uranium mines were as follows:
• Individuals living for a lifetime 1 mile (1.6 km) from an inactive mine would have an
increased chance of lung cancer of about 2-3 x 10"5.
•• The amount of radon-222 released each year from all inactive uranium mine sites would
(cumulatively) cause about 0.1 lung cancers fatalities in the lifetime of the regional
population living within 50 miles (80 km) of these sites.
The study found insignificant concentrations of hazardous air emissions at inactive sites and thus
concluded that their health impacts would be insignificant as well. Although the study
acknowledged the potential for hazards from buildings that use uranium mine wastes as
construction material, it did not formally analyze the hazard. However, it did mention that
building on contaminated land could increase indoor radon concentration and, thus, increase the
risk of lung cancer in the residents (U.S. EPA 1983b). The study referenced an earlier study (out
of print) jointly conducted by EPA and the Atomic Energy Commission in 1972, that identified
about 500 buildings in several western states that exhibited anomalous gamma radiation readings
that appeared to be associated with uranium mine wastes. This is further discussed in Chapter 4
of this volume. Tables 1.2 and 1.3 present the specific lifetime cancer risk estimates due to
radioactive airborne emissions for one year of exposure and over a lifetime of exposure.
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Table 1-2. Estimated Lifetime Fatal Cancer Risks from 1 Year of Exposure to
Airborne Uranium Mine Emissions
The cancer risk from inactive uranium mine radon emissions are generally low for 1 year of exposure.
Source of Exposure
Inactive surface mines— total
Particulates and Radon-222
Radon-222 daughters
Inactive underground mines— total
Particulates and Radon-222
Radon-222 daughters
Risk to Maximum
Exposed
Individual"
4.7 x 10'7
5.5 x ID'8
4.2 x ID'7
2.8 x 1(T7
1.5 x 10'8
2.7 x ID'7
Risk to Average
Exposed
Individual11
8.9 x lO'10
6.4 x ID'11
8.3 x ID'10
1.2 x 10'9
2.0 x lO'11
1.2 x ID'9
Collective Risk to
Regional Population
1.3 x lO'5
9.1 x ID'7
1.2 x ID'5
4.5 x 10'5
7.4 x 10'7
4.4 x ID'5
a An individual living within 1 mile (1.6 km) downwind from the mine.
b The average individual in the regional population within a 50-mile (80-km) radius of the model mine.
Source: U.S. EPA 1983b, Table 6.11.
Table 1-3. Estimated Lifetime Fatal Cancer Risks from Lifetime Exposure to
Airborne Uranium Mine Emissions
The risk to the average person from uranium mine emissions is low. While the risk to the maximally
exposed individual is significantly larger, it is still within the Superfund Iff4 - KT6 risk range.
Source of Exposure
Inactive surface mines— total
Particulates and Radon-222
Radon-222 daughters
Inactive underground mines— total
Particulates and Radon-222
Radon-222 daughters
Maximum Exposed
Individual"
3.4 x ID'5
3.9 x lO'6
3.0 x lO'5
2.0 x 1Q-5
1.1 x lO'6
1.9 x lO'5
Average Exposed
Individual11
6.3 x 1Q-8
4.5 x lO'9
5.9 x lO'8
8.6 x ID'8
1.4 x lO'9
8.5 x lO'8
a An individual living 1 mile (1.6 km) downwind from the mine.
b The average individual in the regional population within a 50-mile (80-km)
radius of the model mine.
Source: U.S. EPA 1983b, Table 6.12.
1.1.3 Applicability of 1983 Risk Estimates
According to Table 6.17 of the 1983 EPA report (U.S. EPA 1983b), radon decay products
account for 88 percent or more of the fatal cancer risk due to emissions of radioactive particles
from inactive surface and underground mines. Risk estimates given for radon decay product
releases from these two types of mines in Tables 6.11 and 6.12 of the report are consistent with
the methodology used by EPA prior to 1988. At that time, 4.6 x 1CT4 cancers were projected per
1-4
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working-level month (WLM)1 of exposure. An analysis of results from the recent BEIR VI
report (NAS 1999b, U.S. EPA 2003a) on risks from exposure to radon suggests that the risk
factor should be 5.38 x 10^ per WLM.
Table 1.4 reproduces the working-level estimates of the model inactive surface mines and model
inactive underground mines which are provided in Table 6.3 of the 1983 EPA report. The values
in Table 1.5 are based on the working-level estimates in Table 1.4. Table 1.5 presents
recalculated risks from 1-year, 30-year, and lifetime exposures to radon decay product emissions
using the higher, current risk factor. The table does not account for exposures for the portion of
time spent outdoors, and for lifetime exposures it assumes an average life span of 75.4 years,
which is slightly longer than the 71-year life span used in the 1983 EPA report. The formulas
used to derive the results in Table 1.5 are as follows:
R! = Lifetime risk for 1-year exposure at 1 WL = 51.56 WLM/WL-y x 5.38 x 10'4 WLM"1 x i year;
• The risk for 30-year exposure at 1 WL = RI x 30 years= 0.83
• The risk for lifetime exposure at 1 WL = RI x 75.4 years = 2.09
Thus, the 1-year exposure risk estimate from radon decay products for the maximally exposed
individual at an inactive surface mine using the 1.8 x 10"5 WL estimate from the model mine in
Table 1.4 would be:
R! = 1.8 x ID'5 WL * 51.56 WLM/WL-y x 5.38 x 10'4 WLM'1 x 1 year= 4.99 * 10'7 = 5.0 x 1(T7
Risks using this updated estimate and presented in Table 1.5 are about 17 percent higher than in
the 1983 report, reflecting the increased risk per working level. One limitation relating to this
conclusion is that no adjustment was made in the calculations for differences in the distribution
of activity-weighted particle size for indoor and outdoor radon exposures.
Table 1-4. Annual Exposure from Radon Decay Product Emissions from Model
Uranium Mines
Source of Exposure
Inactive surface mine
Inactive underground mine
Average Radon Daughter Concentration
(Working Levels)*
Maximum Exposed Individual a
1.8 x 1(T5
1.1 x 1(T5
Average Exposed Individual b
3.5 x KT8
5.1 x KT8
* A Working Level is defined in footnote 1 of this chapter.
a An individual living 1 mile (1.6 km) downwind from the mine.
b The average individual in the regional population within a 50-mile (80-km) radius of the
model mine.
Source: U.S. EPA 1983b, Table 6.3.
1 The working level (WL) is defined as any combination of short-lived radon decay products (through
polonium 214) per liter of air that will result in the potential emission of 1.3 x 105 MeV of alpha energy. A person
exposed to one WL for 170 hours is said to have acquired an exposure of one working-level month (WLM) (Shapiro
1990). This 170-hour value is based on the typical number of hours underground miners worked in 1 month.
1-5
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Table 1-5. Estimated Individual Lifetime Fatal Cancer Risks for
Various Exposures to Radon Decay Products
With the modification for the current risk methodology, the lifetime fatal cancer risk from
radon decay products is still within or below the Superfund KT4 - KF6 risk range.
(See the discussion for additional background of the risk estimates.)
Source of Exposure
Inactive surface mine
Inactive underground mine
Exposure Duration
1 year
30 years
75.4 years (lifetime)
1 year
30 years
75.4 years (lifetime)
Lifetime Risk of Fatal Cancer
Maximum Exposed
Individual a
5.0 x 10'7
1.5 x 1Q-5
3.8 x 10'5
3.1 x 10'7
9.2 x 10'5
2.3 x 1Q-5
Average Exposed
Individual b
9.7 x 10'10
2.9 x 1Q-8
7.3 x l(T8
1.4 x 10'9
4.3 x lO'8
1.1 x ID'7
a An individual living 1 mile (1.6 km) downwind from the mine.
b The average individual in the regional population within a 50-mile (80-km) radius of the model mine.
Source: U.S. EPA 1983b, Table 6.17.
1.2 1989 EPA Study in Support of NESHAPs
In 1989, EPA conducted risk assessments for active underground uranium mines and surface
uranium mines (U.S. EPA 1989a), in support of the National Emission Standards for Hazardous
Air Pollutants (NESHAPs) for Radionuclides (U.S. EPA 1989b, c). While some of the
information in this investigation was based upon U.S. EPA 1983 (a, b, c), the study also included
some new field work and analysis. The study found that of all the radionuclides emitted, radon
decay products posed the greatest cancer risk. The maximum exposures from underground mines
would create lifetime individual fatal cancer risks of greater than 1 x !CT4,with a maximum of 4 x
1CT3. The maximum individual risk of fatal cancer from radon decay products at surface uranium
mines was estimated to be 5 x !CT5;this risk estimate, too, would be slightly higher, given the
current methodology. The 1989 study found that only a limited number of people lived within
several hundred feet of the mines and would have been exposed to the maximum levels; most of
the nearest residents lived several miles from the mines.
1.3 Uranium Mines on the National Priorities List
Although several uranium mill tailings sites are on the Superfund National Priorities List (NPL),
only two uranium mines are on the list: Midnite Mine, near Wellpinit, Washington, and the
Fremont National Forest—White King/Lucky Lass Mines, Oregon. Both sites have progressed
far enough in the Superfund process to have had a cleanup remedy selected in a Record of
Decision (U.S. EPA 2001b, U.S. EPA 2006c). Figures 1.1 and 1.2 are aerial images of Midnite
Mine and the White King/Lucky Lass Mine sites, respectively.
1-6
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Figure 1-1. Aerial Image of Midnite Mine, Washington State
Midnite Mine is on the NPL. The site has uranium and other heavy metal contamination
in the disturbed area and two pit lakes.
Source: Photo courtesy of EPA Region 10 Superfund Program.
No one is currently living at the White King/Lucky Lass site, nor is a future resident anticipated,
given that the site is on U.S. Forest Service property and is not near population centers.
However, the risk assessment did assume a future resident as a scenario. In addition, the
receptors evaluated included a site worker (e.g., timber or U.S. Forest Service employees) and a
recreational user. The following areas were used as exposure points (U.S. EPA 2001b):
• The protore stockpile at the White King Mine
• The overburden stockpile at the White King and Lucky Lass mines
• Off-pile areas at the White King and Lucky Lass mines
The primary chemicals of concern at the White King/Lucky Lass site were arsenic in soil and
shallow groundwater, uranium-234/238 in stockpile groundwaters, radium-226/228 in soil and
shallow bedrock wells, and radon in water. Of note, and in spite of several high radon flux rates,
inhalation of radon in ambient air was not an issue, since radon concentrations from the
stockpiles were equivalent to background concentrations.
1-7
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Figure 1-2. Aerial Image of White King and Lucky Lass Mines, Oregon
The Lucky King Mine pit lake is approximately 5,000 feet (1,500 meters)
northwest (left) of the White King Mine pit lake.
Table 1.6 summarizes the risks at the mine sites for the human receptors. With the approach
used in the Record of Decision, the exposure assessment indicated an extremely high risk to
future potential residents and child recreational users. The high risks were primarily due to
ingestion of arsenic in soils and shallow groundwater and external radiation from radium. In the
ecological assessment, no adverse effects were seen from the radionuclides. However, some
potential adverse ecological effects were identified due to arsenic, selenium, antimony, lead, and
mercury in surface and subsurface soils at the White King Mine. At Lucky Lass, only slightly
elevated risks (the noncarcinogen chemical hazard index ranging from 1 to 3) were predicted for
the vagrant shrew and terrestrial plants exposed to arsenic and silver in surface soil. In contrast,
Midnite Mine has a greater potential for future use, but the cancer risks were predicted to equal 8
-3.
10" for a resident of the affected area and 2*10" for recreational visitors.
1-8
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Table 1-6. Potential Cancer Risks from the White King/Lucky Lass and
Midnite Mine Sites
The combination of arsenic and radium produces very high cancer risks to potential on-site residents.
Receptor
White King Mine
current adult worker
Future recreational
user (child) at the
White King Mine
Potential future
resident (adult) at the
White King Mine
Potential future
resident (child) at the
White King Mine
Potential future
resident at the Lucky
Lass Mine
Potential future
resident at the Midnite
Mine Area
Total Cancer Risk
6 x 10'5
4 x ID'4
3 x ID'1
2 x 10'1
Hazard Index values for
noncarcinogenic effects to
current and potential future
child recreational users were
4 and 11, respectively, and
higher for potential future
residents from ingestion of
arsenic and manganese in
shallow bedrock
groundwater and ingestion of
arsenic in soil.
1 x ID'3
1
Pathway
Ingestion of arsenic in soil
and exposure to external
radiation from radium-
226/228 in the top 6 inches
of soil.
Arsenic in soil, exposure to
external radiation from
radium-226/228 in soil and
ingestion of arsenic in
Augur Creek and White
King groundwater.
Ingestion of arsenic in soil
and exposure to external
radiation from radium-
226/228 in the top 6 feet of
soil, ingestion of arsenic in
shallow bedrock
groundwater, inhalation of
radon in shallow bedrock
groundwater, and exposure
to arsenic in White King
pond surface water and
sediment.
Notes
Current exposure estimates for
soil are based on 0-6 inches;
future exposure estimates for
soil are based on 0 - 6 feet.
Deep bedrock water contains
high levels of naturally
occurring arsenic, radon, and
minerals that would preclude
its use as drinking water.
Note: A Hazard Index value below 1 indicates no adverse health effects are expected as a result of exposure.
Source: U.S. EPA 200Ib.
1.4 DOE Report on Costs of U.S. Uranium Mine Environmental Restoration
A report commissioned by the U.S. Department of Energy (DOE) found that a number of
uranium mines are undergoing or have completed remediation (U.S. DOE/EIA 2000).
According to the report, 21 mines, primarily in Wyoming and Texas, were selected for analysis
for one or more of the following reasons: (1) substantial output of uranium concentrates,
(2) major impact on the environment, and (3) significant costs required for remediation. While
the report does not specify whether these sites are undergoing risk assessments, it does specify
whether a particular site has an exposure pathway of surface water, groundwater, or windblown
particulates. The information lists groundwater as an exposure pathway for many of the mines,
while the surface water and windblown particulate pathways are not as prevalent.
1-9
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2.0 GEOGRAPHIC ANALYSIS ON THE LOCATION OF
URANIUM MINES
With the exception of some phosphate mine areas in central and northern Florida, people are
most likely to be exposed to uranium mining-related TENORM in the western United States.
This chapter provides a geographic analysis of the spatial locations of western mines in
proximity to human populations, cultural and political features and boundaries, and
environmental features. The use of geographical information system (GIS) software provides a
systematic means to understand the potential impacts and scenarios by which humans and the
environment may be impacted by uranium mines.
Figure 2.1 was generated from uranium mining-related records from the U.S. Bureau of Mines—
now U.S. Geological Survey (USGS)—Mineral Availability System/Mineral Industry Location
System (MAS/MILS) database from the EPA BASINS data (U.S. EPA 2001c). While about half
of the 8,234 locations are documented as producing ore, the remaining records may identify
mines or simply locations with uranium. Of the 8,234 records, 4,141 are categorized as
"producer" or "past producer," and these terms are being used as proxies for known mines.
Another 63 records are classified as mills or processing plants, and once these are removed, the
4,078 records that are left are assumed to be former mines. Of the 4,078 mines, about 3,000 are
in Colorado, Utah, Arizona, and New Mexico. Similar information comes from the Department
of Energy's (DOE) Energy Information Administration database (Smith 2002), which has 3,502
records for Colorado, Utah, Arizona and New Mexico. Within this set, 2,952 mines had at least
some ore production (Table 2.1), similar in number to the MAS/MILS data.
Table 2-1. Mine Sizes for Four-Corners States
Of~3,500 uranium mines in Colorado, Utah, Arizona, and New Mexico,
2,952 mines had at least some ore production.
Ore Production (Tons)
<100
100-1,000
1,000-100,000
>1,000,000
Data withheld as confidential business information.
Total
Number of Mines
1,192
615
952
5
188
2,952
Source: Smith 2002.
The definition of a mine leads to problems with determining how many mines really exist. Even
a single data set may have different interpretations for what could be considered a mine.
Records may indicate multiple mine portals for an underground mine, for example. EPA has
compiled a database of uranium locations from different sources totaling about 15,000 records,
from which an attempt has been made to remove redundant records (U.S. EPA 2006b). The EPA
database thus lists several thousand more mines than any other data set. Table 2.2 compares the
number of records by state for the USGS MAS/MILS database (U.S. EPA 200 Ic) and
unpublished USGS data sets by Finch (1998). The BASINS MAS/MILS database typically lists
more mines than the Finch data set, although Finch has noted more mines in Texas and South
2-1
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Dakota. The EPA ULD Compilation (U.S. EPA 2006b), as sorted for this analysis,1 contains
nearly 11,000 records, and typically has more uranium locations per state than the other data sets.
Figure 2-1. Mines and Other Locations with Uranium in the Western U.S.
Hundreds of active and abandoned uranium mines are scattered
over wide areas of the western United States.
Source: MAS/MILS Database.
1 For this comparison, the EPA ULD Compilation was sorted to delete the Mineral Resource Data System
(MRDS) data, because many of the records were identified as simply drill holes, or mineral locations and also
included many eastern locations not relevant to this study. In addition, location names that were variations on
unknown or unnamed in the MINE NAME field in the ULD were removed so that the remaining records were more
likely to be actual mining sites. For example, records with MINE NAME fields with entries such as "UNKNOWN,"
"UNKNOWN NAME," "UNNAMED PROSPECT," and "UNNAMED URANIUM OCCURRENCE" were
deleted.
2-2
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Table 2-2. Comparison of Data Compiled from Uranium Mine Records
Different data sets have different estimates of the number of uranium mines.
State
Arizona
California
Colorado
Idaho
Missouri
Montana
Nevada
New Mexico
North Dakota
Oklahoma
Oregon
South Dakota
Texas
Utah
Washington
Wyoming
Totals
BASINS MAS/MILS
All Records
466
243
2,286
234
2
195
363
756
23
2
100
197
69
1,542
68
1,616
8,162
Producer or Past Producer
146
23
1,631
34
0
47
24
337
16
0
15
130
69
911
13
682
4,078
Finch
403
59
1,262
6
0
31
20
330
13
8
6
203
90
1,120
20
625
4,196
EPA ULD
Compilation
1,104
268
2,268
216
2
482
396
2,247
109
0
56
307
136
2,047
98
1,172
10,908
Sources: U.S. EPA 2006b, U.S. EPA 2001c, and Finch 1998.
2.1 Errors in Mine Locations
The mine record data used for most of the geospatial analyses, have two distinct error types. In
addition to the definition of "mine" that was discussed above, there are errors of omission and
commission (i.e., erroneous locations in the database, as well as actual mines not represented).
However, accuracy of the data was checked in the EPA ULD compilation (U.S. EPA 2006b),
and the mines were typically found to be within several hundred meters of mines identified on
U.S. Geological Survey maps. The primary endpoint of the analyses described in this document
is in terms of the radiation dose to an individual, not the collective dose to a population group.
For this reason, errors in the total number of mines will not have a significant effect on the
overall conclusions. There are also location precision errors (i.e., a listed mine not in its actual
location as shown on USGS maps, for example). The latter are not likely to affect the analyses
in this document because of the focus on risks to individuals, not populations.
2.2 Number of People Potentially Exposed to Uranium Mine Wastes
The 1983 EPA study found that, for releases to air and surface waters, the cancer risks were less
than 10"4 and 10"6 for people living 1 mile or farther from active and inactive mines, respectively.
Based on this information, we have assumed that the populations primarily at risk live within 1
mile (1.6 km) of uranium mines and, thus, have estimated the number of people within 1 mile of
a uranium mine. We have also estimated the number of people who live nearby (within 5 miles
20
-3
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[8 km]) to help identify a likely population that may engage in recreational or other visitation
activities in areas with unreclaimed uranium mines.
To estimate the number of people who live in proximity to uranium locations, we queried the
4,078 records in the MAS/MILS mine database in Arc View 8.2, Arc View 9.1, and Spatial
Analyst (collectively, Arc View), using population data from the 2000 census (ESRI 2001).
About 800,000 people are estimated to live within 5 miles of a uranium mine, and about 55,000
(or about 10 to 15 people per mine on average) are estimated to live within 1 mile of such a
mine. About 10,000,000 people are estimated to be within 50 miles (80 km) of a recorded mine,
with 502 of 4,078 mines located within 50 miles (80 km) of cities whose population is greater
than 30,000. A search indicates that 33 of the recorded mines are within 1 mile (1.6 km) of a
U.S. Bureau of Census "place" in the Arc View database, most of which are in Colorado; 141 of
the mines are within 5 miles of a place (Table 2.3). In comparison, an analysis of the 10,908
"mine" locations from the ULD found that the population within 1 mile (1.6 km) and 5 miles
(8.0 km) of a uranium location was 227,692 and 3,993,642, respectively.
The low number of people living within 1 mile (1.6 km) of a mine can be attributed to the fact
that 7,076 of the MAS/MILS 8,234 records (86 percent) are located on federal land, while about
90% of the mines with known production are on federal land (Table 2.4). In the ULD data set,
8,124 of the 10,908 locations (74.5%) of the locations were on federal land (Figure 2.2 is a map
of the ULD locations and federal lands). A query of the 7,076 mine records using Arc View
revealed that 6,127 mines could be attributed to a specific federal land management agency, with
most on U.S. Department of the Interior lands or Forest Service lands (Table 2.4). With the
majority of the mines on federal land, people who use these sites for recreation would most
likely be subjected to the greatest potential for exposure to uranium mine wastes. An exception
to this would be the uranium mines on Tribal lands, where the Tribal members would receive the
greatest exposure potential. Five percent (221) of the 4,078 mine records in the MAS/MILS
database are on Bureau of Indian Affairs land, while eight percent (898) of the 10,908 records of
the EPA ULD used in this analysis are on Bureau of Indian Affairs land.
Of the 69 mines in the MAS/MILS data identified in Texas, none are on federal lands. Over one
half of the past-producer mines in Wyoming (456 of 682) are on federal lands. Of the 1,631
mines in the past-producer Colorado data set, 1,572 are on federal lands.
2.3 Mines by Watershed
One method used to view the potential for impact by mining on a region and to identify the most
likely areas to be affected is on a watershed basis using geographic information system
technology (Ferderer 1996). In Figure 2.3, uranium mines have been grouped in watersheds
identified by 8-digit hydrologic unit codes (HUCs). Several watersheds have more than 100
uranium mines while a number of others have more than 50 mines. As might be expected from
the discussion above, the highest watershed mine density is in Colorado, Utah, and Wyoming. In
the watersheds with only a few mines, the mines typically produced uranium as a by-product of
other mining, such as copper. One example is the Lefthand Creek mining area along the Front
Range in Colorado where gold and silver were the primary metals mined, but also mined were
tungsten, copper, fluorspar and uranium (U.S. EPA 2003b). Watersheds are also a unit
considered in mine remediation (U.S. EPA 2003b, Buxton et al. 1997).
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Table 2-3. Estimated Number of People within 1 Mile (1.6 km) and 5 Miles (8 km) of a
Recorded Mine
The 4,078 mine records in the BASINS MAS/MILS database and 10,908 records
from the EPA Uranium Location Database Compilation (U.S. EPA 2006b) were
queried for the number of people near the uranium locations. Colorado accounts
for most of the population living near current and past uranium mines.
State
Arizona
California
Colorado
Idaho
Montana
Nevada
New Mexico
North Dakota
Oregon
South Dakota
Texas
Utah
Washington
Wyoming
Totals
People within 1 Mile
From 4,078 Records Using
Producer or Past Producer
and 2000 Census Data
1,045
1,068
33,191
494
891
188
6,013
1,114
370
2,889
591
1,387
162
5,196
54,599
People within 1 Mile
From 10,908 Records of
EPA ULD and 2000
Census Data
21,727
34,867
67,319
5,399
5,954
17,369
46,736
1,262
1,134
2,956
871
7,169
5,144
9,785
227,692
People within 5 Miles
From 4,078 Records Using
Producer or Past Producer
and 2000 Census Data
12,160
59,437
518,357
5,803
8,233
11,332
84,869
2,159
6,162
5,954
11,700
22,376
3,472
61,701
813,715
People within 5 Miles
From 10,908 Records
of EPA ULD and 2000
Census Data
438,581
758,545
1,188,827
89,486
89,573
577,189
512,102
3,518
30,894
8,538
32,640
106,015
79,200
78,534
3,993,642
Figure 2.4 illustrates one region of high-density uranium locations in drainages in southwest
Colorado and eastern Utah. Figure 2.4 contains surface and underground mines, in addition
to mines whose types are listed as "unknown" in the MAS/MILS database. This region
typically has horizontal rock layers that have been incised by streams exposing the uranium-
bearing layers, such as the Chinle Formation. In this figure, flat-lying areas appear generally
featureless, whereas areas incised by streams show relief and appear to be v-shaped. Many
of the mine locations are adjacent to streambeds where the mining has taken advantage of
exposed uranium layers. The slopes along the canyon walls could enhance movement of
radioactive materials to streambeds via mass-movement processes. Since radium and
uranium may largely precipitate out of solution or adhere to particles and come to rest in
sediments, benthic organisms may be the most potentially affected. However, large-
magnitude events (e.g., flooding) could resuspend the material and move it around the
streambeds, with higher concentrations likely developing in slack-water deposits where the
water flow slows.
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Figure 2-2. Uranium Locations from EPA Database and Federal Lands
About three-fourths of the uranium locations in the EPA Uranium Location Database
are on Federal Lands. Thus, the most likely exposure or risk scenario for many of
the uranium mine locations is the recreational scenario, such as hiking,
camping, use of all-terrain vehicles or other short-term activity.
Federal Land Management Agencies
With Significant Uranium Locations
Bureau of Indian Affairs
National Park Service
Forest Service
Bureau of Land Management
EPA ULD Locations (10.908 Records)
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Table 2-4. Number of Mines on Federal Lands in Selected States
Most of the uranium mines on federal lands can be attributed to a specific federal land management agency;
the U.S. Department of the Interior and U.S. Forest Service are the two primary land management agencies.
Federal Land Management
Agency
Department of Interior
Bureau of Land Management
Fish and Wildlife Service
Bureau of Indian Affairs
National Park Service
Bureau of Reclamation
Department of Defense
Forest Service (USD A)
Unknown
Total
From 8,234 Records in
BASINS MAS/MILS
Database
4,241
7
446a
121b
3
12
1,297
949
7,076
From 4,078 Records Using
Producer or Past Producer
and 2000 Census Data
2,405
0
223
43
1
6
515
500
3,693
a Primarily on Navajo lands.in Arizona, New Mexico, and Utah, in that order
b Primarily in Utah and California, with California primarily having unnamed prospects.
Figure 2-3. Western Uranium Mine Density by 8 Digit Hydrologic Unit Code
The greatest number of mines (745) in the MAS/MILS data is found in the Upper Dolores Watershed,
located primarily in southwest Colorado with a small area in Utah. Other watersheds with more than 300 uranium
mines are the Lower Dolores (Colorado and Utah) and San Miguel (Colorado) Watersheds.
Legend
^B HUCS> 100 mines
^B HUCS 51 -100 mines
HUCS11 -50 mines
HUCS 6- 10 mines
2-7
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Figure 2-4. Uranium Locations in Southwest Colorado and Southeast Utah
This region typically has horizontal rock layers that have been incised by streams exposing the uranium bearing
layers, such as the Chinle Formation. Flat-lying areas appear generally featureless, whereas areas incised by
streams show relief and appear to be v-shaped. Many of the mine locations are adjacent to streambeds where the
mining has taken advantage of exposed uranium layers. Mines from the MAS/MILS data are superimposed on
digital elevation data.
API
of/
jroximate Location
\.rea in Main Image
3S
%— j^-»- — i
Legend
hdii^r f^C) r>f^r
n
A
6
1
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3.0 CANCER RISKS FROM ON-SITE EXPOSURE
This chapter examines the potential scenarios, exposure pathways, and risks of cancer to humans
that may be posed by exposure to TENORM from abandoned uranium mine wastes.
3.1 Potential Scenarios and Exposure Pathways for the General Public
Given our knowledge of uranium mine TENORM wastes,l there are several possible exposure
scenarios for humans to the various hazards posed by these materials: on-site recreation, homes
with contaminated building materials, on-site residents, and near-by residents.
3.1.1 On-Site Recreation
Since most uranium locations are on federal lands, the primary exposure scenarios to TENORM
wastes at uranium mines would involve recreational use of the site, in which the abandoned mine
is visited occasionally by hikers, campers, or driven through by all-terrain vehicles (ATVs).
Recreational use by children may occur if a site is located near houses, as, for example, on Tribal
lands in Arizona and New Mexico. A typical recreational scenario might take place at the White
King and Lucky Lass mines in Oregon, which are on national Forest Service land and can be
accessed only by hikers. A less common but more troubling recreational case involved the pit
lake at the Yazzie-312 surface mine in Cameron, Arizona, which was approximately 300 feet
(-100 meters) across and referred to by local citizens as the "swimming hole" (see Figure 3.1).
The site, just off a highway, attracted swimmers because the area lacks natural lakes or streams,
other than during periods of the year when the rainfall is heavy. The pit has since been filled and
the area reclaimed. Users would likely visit unreclaimed uranium mines for short periods of
time, such as two weeks, which is the common maximum time for which the National Park
Service issues backcountry permits. Occupational workers, such as government employees or
contractors performing site investigations, could also spend similar periods of time at these
locations. The primary exposure pathways would be external exposure and drinking
contaminated water from an adjacent spring or stream. Pathways of secondary importance
include inhalation of dust, exposure to radon, ingestion of dust on dried or prepared foods, and
inadvertent ingestion of soil.
3.1.2 Building Materials
A second scenario that has been known to occur, but whose frequency is unknown, is the use of
uranium mine waste materials for building construction. Although most of the uranium locations
are in areas where recreation is the most likely scenario, some uranium locations are near roads,
including unimproved dirt roads, or near rural communities where waste material could be
accessed. These materials could be transported from a nearby site and used in the construction
of houses, when other building materials are difficult or too expensive for a homeowner to
obtain. A discussion of risks from uranium mine wastes in building materials is presented in
Chapter 4 of this report.
1 Characteristics and origins of wastes mentioned in this study are more fully described in Chapter 3 of
Volume I of this report (U.S. EPA 2006a).
3-1
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3.1.3 On-Site Residents
A third scenario involves on-site residents. Given such factors as the nature of uranium mine
waste materials, the isolation of many of the sites, the lack of potable water in many cases, and
the lack of infrastructure, this scenario may have a low probability, except for some Tribal
populations. The risks for such a scenario would be at the highest end of the risk spectrum and
would provide an upper bound for risks. The White King Mine analysis of risks found that a
future resident at the White King Superfund site would have an extremely high risk of
developing cancer (see Table 1.6). Subpart B of Title 40 of the Code of Federal Regulations,
Part 192 (40 CFR 192), which establishes cleanup standards for uranium processing sites, uses a
radium surface soil standard of 5 pCi/g (185 Bq/kg) above background, or below, as the cleanup
level, with the emphasis on preventing elevated radon levels. This radium cleanup level has
been used as a relevant and appropriate requirement to establish cleanup criteria at some
Superfund sites. The radon flux standards in 40 CFR 192 assume sand-like uranium mill tailings
and limit the radon flux rate to 20 pCi m'Y1. Uranium mine overburden, or protore, has elevated
radon flux rates in a similar range as uranium mill tailings, although the average flux rates may
be lower as described by SC&A (1989) and U.S. EPA (2006a, Chapter 3).
Figure 3.1. Uranium Mine Pit Lake
Pit lake ofYazzi-312 surface mine in Cameron, Arizona, referred to by local citizens
as the "swimming hole. " Suspended sediment transformed the pit water
to a milky white color. The pit lake has been reclaimed.
Photograph by Loren Setlow (U.S. EPA)
3-2
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3.1.4 Nearby Residents
The last scenario involves people living next to a uranium mine, which has been found to occur
in the Navajo Nation. People may live within a short distance of overburden piles and be
exposed to uranium from windblown particulates (inhalation of dusts), contaminated water, and
external radiation.
The 1983 EPA Report to Congress studied this scenario as part of an investigation of risks to the
hypothetically maximally exposed individual located 1 mile (1.6 km) from the center of average
and large active and inactive mine sites (EPA 1983a, b, c). The 1983 EPA Report to Congress
examined ten pathways. The study looked at risks from pathways including inhalation of radon
decay products, external exposure, eating food grown in the area, fish consumption, and drinking
milk and eating meat from cattle that had grazed in the area and consumed contaminated water.
The study concluded that most of the pathways did not pose great risks.
The study found that radon posed the greatest risk in all scenarios, with large active underground
mines emanating the highest concentrations. The maximally exposed individual's risks from
radioactive airborne emissions from inactive surface and underground uranium mines were
modeled and estimated to be 3.4 x 1CT5 and 2.0 x 1CT5, respectively. These risk estimates
assumed exposure for 71 years to inactive mine effluents. Similar results were calculated in the
1989 NESHAPs (National Emission Standards for Hazardous Air Pollutants) study (U.S. EPA
1989c). As discussed in Chapter 1, the estimated risk per working level has increased, so these
risks would be slightly higher than those identified in the 1983 report and in the 1989 study (U.S.
EPA 1989a). The updated risk estimates for inhalation of radon decay products from the 1983
study are listed in Table 1.5. However, since this scenario was considered in the 1983 Report to
Congress and in the 1989 NESHAPs study, it is not considered further in the present analysis.
3.2 Methodology Used in This Analysis
This report focuses on risks that uranium mine TENORM wastes could pose for those people
who visit inactive uranium mine sites. This analysis complements the 1983 EPA study, which
looked primarily at off-site exposures from uranium mines, although it acknowledged the
potential on-site health hazards. A key purpose of this approach is to help prioritize the types of
uranium mine site wastes and exposures that pose the greatest risk. While some of the analysis
examines residential exposure on a site, the focus is more on non-residential uses for the reasons
discussed in this section.
Given the limited available data, multiple site characteristics, and the multimedia exposure
pathways, multiple approaches were taken to evaluate the risks at these sites. These include
reviewing existing data discussed earlier, using geographically-based queries of uranium mine
and population data, the Superfund Soil Screening Guidance (SSG) approach for chemicals and
radionuclides whenever applicable (U.S. EPA 1996a and 2000b), risk calculations produced for
the radionuclides in drinking water regulation (U.S. EPA 2000c), and the use of RESRAD
BUILD 3.21 (Yu et al. 1994) for examining building materials. This approach uses applicable
peer-reviewed methodologies. The equations in the Soil Screening Guidance: User's Guide
(U.S. EPA 1996a), Soil Screening Guidance for Radionuclides (U.S. EPA 2000b), and
Supplemental Guidance for Developing Soil Screening Levels for Superfund Sites (U.S. EPA
3-3
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2002) were used, because they are appropriate for looking at generic sites when only limited site-
specific data are available. Since the intent of this analysis is meant to be scoping in nature and
the information on the sites is limited, the SSG approach is appropriate for identifying the
situations that may be of concern. Since this approach is for screening purposes where the intent
is to ensure that potential problems are identified, the SSG methodology tends to lead to
conservative risk estimates, or risks that are more likely to be overestimated. The risk estimates
become more accurate with more site-specific data. Please note that all quantified risks included
in this report refer to lifetime cancer risk.
An approach used at Superfund and RCRA (Resource Conservation and Recovery Act) sites is to
identify preliminary remediation goals (PRGs) that are intended as initial guidelines, and not
necessarily as final cleanup levels. PRGs are risk-based concentrations (assuming a target
lifetime risk of 1 x 10"6), derived from standardized equations similar to those found in the Soil
Screening Guidance for Radionuclides (EPA 2000b). An Internet-based PRG calculator and
tables of default values for radionuclides can be found at http://epa-prgs.ornl.gov/radionuclides.
A major difference between the SSG methodology and the PRG approach is that the SSG
methodology allows examination of an individual pathway, while the PRG uses an all-pathway
approach. Since part of the intent of this analysis was to investigate individual pathways, the
SSG approach was used. In addition, the PRG approach does not have a recreational scenario,
which is a primary scenario identified for these mines. Although this approach was not used in
this report to evaluate risks, for illustrative purposes the preliminary remediation goals for
several scenarios are presented in Table 3.1.
Using the conservative SSG for radionuclides methodology, we have made some estimates of
lifetime cancer risk for different exposure time periods and different concentrations for natural
uranium, Ra-226, and Th-232. Natural uranium is assumed to include U-234, U-235, and U-238,
in natural isotopic abundances. U-238 is in secular equilibrium with its short-lived progeny, U-
234 is in secular equilibrium with Th-230, while U-235, Ra-226, and Th-232 are in secular
equilibrium with their entire decay chains. The slope factors for natural uranium are expressed
in terms of pCi of U-238.2 Arsenic was evaluated using a similar approach, but using the general
SSG (U.S. EPA 1996a and 1996b) methodology.
For example, the inhalation slope factor (lifetime risk of cancer morbidity per pCi inhaled) for Ra-226
includes the contribution of all of its short- and long-lived progeny. This approach was employed because exposure
to airborne radium particles at a mine site would most likely include most of its progeny in equilibrium. This
approach slightly overestimates the risks in the case of Ra-226, because the progeny may not be in full equilibrium
since some of the Rn-222 may have diffused away. The uranium slope factors do not include Ra-226 and its
progeny, because separate SSLs are developed for Ra-226.
3-4
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Table 3.1. Selected Radionuclide Toxicity and Preliminary Remediation Goals for
Superfund for Comparison with the SSG Pathway-Specific Approach
Element and Isotope
Radium 226 + D
Thorium 232
Uranium 238 + D
Preliminary Remediation Goals (PRGs)
(for concentrations above background)
Residential Soil
(pCi/g)
0.012
3.1
0.74
Agricultural Soil
(pCi/g)
0.0006
0.0094
0.0015
Outdoor Worker
Soil (pCi/g)
0.026
1.9
1.8
Soil to
Groundwater
DAF = 20
(pCi/g)
0.32
6.1
0.12
D means that decay products are included
DAF is Dilution/Attenuation Factor
Table Source: August 4, 2004 Radionuclide Toxicity and Preliminary Remediation
Goals for Superfund, at http://epa-prgs.ornl.gov/radionuclides/download.shtml.
3.3 Recreational Scenario Risk Calculations
3.3.1 Risk from External Exposure to Radium, Thorium, and Uranium
The SSG methodology assumes a linear relationship between a person's incremental cancer risk
from exposure to radium (Ra-226), thorium (Th-232), and natural uranium (U-238 + U-235).
The incremental lifetime cancer risk level of 10"6 is usually the baseline level of risk that is
acceptable, and 5 x 10"4 is typically at the high end of the range of acceptability. Thus the Soil
Screening Levels (SSLs) are evaluated for this range.
Soil Screening Level (SSL) =
TR
•where:
SFE * EF/365 * ED * ACF * [ETO + (ETPGSF)]
TR = Target lifetime cancer risk (unitless)
SFE = Slope factor for external exposure to soil contaminated
= 8.49 x 1Q-6 for Ra-226
EF = Exposure frequency (days/year)
ED = Exposure duration (years);
results in risk per total number of days on site
For residential exposure, ED is used to represent the
exposure over a number of years, frequently 30 years.
ACF = Area correction factor for smaller sites
= 0.9 if area < 1,000m2
ETO = Estimated fraction of time outdoors on site
ETI = Estimated time indoors
GSF = Gamma-shielding factor
10-6-5
io-4)
variable (1
1.23 x 10'5 for Th-232
2.14 x IO"7 for U-natural
variable
1
1
1
0
0
3 Includes short- and long-lived decay products, as discussed in preceding section. Slope factors for
radionuclides for all exposure pathways are based on U.S. EPA's Health Effects Assessment Summary Tables
(HEAST) (http://www.epa.gov/radiation/heast/index.html). The slope factor calculations can be found in Appendix
II Calculation of Slope Factors for NORM Decay Series.
3-5
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Because of the nature of the recreational scenario, some of the typical assumptions have been
changed. In the above equation we assume that the person spends the entire day at the site, with
no indoor time—that is, the individual spends all day on the waste material and sleeps in a tent or
other light structure that provides no appreciable shielding. Since no time is spent indoors, the
indoor part of the equation with the gamma shielding does not come into play. For a Superfund
target risk of 1 x 10"6 for 14 days of exposure and the assumptions stated above, the Ra-226 soil
screening level would be-3.1 pCi/g (-114 Bq/kg), but for one day of exposure at a 1 x 10"6
target risk, the Ra-226 soil screening level would be -43 pCi/g (-1,590 Bq/kg). Table 3.2 and
Figure 3.2 illustrate the relationship between radium concentration and risk for different times of
exposure, Table 3.3 and Figure 3.3 present the corresponding data for exposure to thorium, and
Table 3.4 and Figure 3.4 provide similar information for uranium. The relationship is linear, so
reducing the estimated time on site by one half (from 100 percent of the time on site to 50
percent) would increase the radium screening level by a factor of two for the same target risk. In
addition, if a typical residential exposure duration of 30 years is used, then the values in Table
3.2 and other tables of soil screening levels used in this chapter would need to be divided by 30;
however, the assumptions used here (i.e., entire day on the waste material) would not be
appropriate for a typical residential scenario. The risk estimated for a recreational exposure could
also be used for occupational workers (government workers or contractors for example) who
spent time at the site for their jobs.
Table 3.2. Soil Screening Levels for External Exposure to Ra-226
Table 3.2 lists the data used to generate Figure 3.2.
Exposure
Frequency
(days)
1
14
30
52
140
350
Target Lifetime Cancer Risk
5 xKT1
1 xKT1
5 x 1(T5
1 x 1(T5
SxKT6
1 x 1(T6
Concentration of Ra-226 (pCi/g)
21,485
1,535
716
413
153
61.4
4,297
307
143
83
30.7
12.3
2,149
153
72
41.3
15.3
6.14
430
30.7
14.3
8.3
3.07
1.23
215
15.3
7.2
4.13
1.53
0.614
43.0
3.07
1.43
0.83
0.307
0.123
3-6
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Figure 3.2. External Exposure - Relationship between Exposure Frequency,
Radium Concentration, and Target Lifetime Cancer Risk
Figure 3.2 is derived from Table 3.2. The x-axis is the activity concentration of radium
in the uranium mine waste material, and the y-axis is the incremental lifetime cancer risk as
a result of exposure from the radium in the waste material for different time periods.
For example, exposure to 12.3 pd/g (454 Bq/kg) of radium, in secular equilibrium
with its progeny, for 350 days, would result in a lifetime cancer risk of 10~4.
1.E-03
1.E-04
1.E-05
1.E-06
-1 day exposure
-14 day exposure
-30 day exposure
-52 day exposure
-140 day exposure
-350 day exposure
1.E-01 1.E+00 1.E+01 1.E+02 1.E+03
Ra-226 Concentration (pCi/g)
1 .E+04
1 .E+05
Table 3.3. Soil Screening Levels for External Exposure to Th-232
Table 3.3 lists the data used to generate Figure 3.3
Exposure
Frequency
(days)
1
14
30
52
140
350
Target LifetimeCancer Risk
5 x l(T4
1 x 1(T4
5 x 1(T5
1 x 1(T5
5 x 1(T6
1 x KT6
Concentration of Th-232 (pCi/g)
14,849
1,061
495
286
106
42.4
2,970
212
99
57
21.2
8.5
1,485
106
49.5
28.6
10.6
4.24
297
21.2
9.9
5.71
2.12
0.85
148
10.6
4.95
2.86
1.06
0.424
29.7
2.12
0.99
0.571
0.212
0.085
3-7
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Figure 3.3. External Exposure - Relationship between Exposure Frequency,
Thorium Concentration, and Target Lifetime Cancer Risk
Figure 3.3 is derived from Table 3.3. The x-axis is the activity concentration of thorium
in the uranium mine waste material, and the y-axis is the incremental lifetime cancer risk
as a result of external exposure to the thorium in the waste material for different
time periods. For example, exposure to 8.5 pd/g (314 Bq/kg) ofTh-232, in secular
equilibrium with its progeny, for 350 days, would result in a cancer risk of Iff4.
1 .E-03
1 .E-04
1 .E-05
1 .E-06
-1 day exposure
-14 day exposure
-30 day exposure
-52 day exposure
-140 day exposure
-350 day exposure
1.E-02 1.E-01 1.E+00 1.E+01 1.E+02 1.E+03
Th-232 Concentration (pCi/g)
1.E+04
1 .E+05
Table 3.4. Soil Screening Levels for External Exposure to Natural Uranium
Table 3.4 lists the data used to generate Figure 3.4
Exposure
Frequency
(days)
1
14
30
52
140
350
Target Lifetime Cancer Risk
5 x 1(T4
1 x 1(T4
5 x 1(T5
1 x 1(T5
5 x 1(T6
1 xlQ-6
Concentration of Natural Uranium (pCi/g U-238)
852,189
60,871
28,406
16,388
6,087
2,435
170,438
12,174
5,681
3,278
1,217
487
85,219
6,087
2,841
1,639
609
243
17,044
1,217
568
328
122
48.7
8,522
609
284
164
60.9
24.3
1,704
122
56.8
32.8
12.2
4.87
3-8
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Figure 3.4. External Exposure - Relationship between Exposure Frequency,
Uranium Concentration, and Target Lifetime Cancer Risk
Figure 3.4 is derived from Table 3.4. The x-axis is the activity concentration of U-238
in the uranium mine waste material, and the y-axis is the incremental lifetime cancer risk as a result
of exposure to uranium in the waste material for different time periods. For example,
350 days of exposure on site to 487 pd/g (18,020 Bq/kg) ofU-238, in secular equilibrium
with its progeny, as well as U-235 in the ratio of natural abundance (see discussion of
uranium progenies earlier in this chapter) would result in a lifetime cancer risk ofl 0~4.
1 .E-03
1 .E-04
1 .E-05
1 .E-06
-1 day exposure
-14 day exposure
-30 day exposure
-52 day exposure
-140 day exposure
-350 day exposure
1.E+00 1.E+01 1.E+02 1.E+03 1.E+04
U-238 Concentration (pCi/g)
1 .E+05
1 .E+06
3.3.2 Risk from Soil Ingestion
While the direct ingestion of soil is possible at a site, it is not likely to be a major exposure
pathway for adults. The following equation uses an age-adjusted soil ingestion factor to account
for the fact that children have a higher intake of soil than adults (U.S. EPA 2000b).
SSL
TR
SFs * IRs * 1 x 10 3 * EF * ED
where:
TR
SFS
IRS
1 x
EF
ED
l(T3 =
Target lifetime cancer risk (unitless)
Soil ingestion slope factor (pCi)"1
Ra-226 = 3.39x l(T9
Th-232 = 3.33 x 1(T9
U-natural = 6.48 x l(r10
Soil ingestion rate (120 mg/day)
Conversion factor (g/mg)
Exposure frequency (variable)
Exposure duration (1 year)
Sample calculation for radium, assuming a target lifetime risk of 1 x 10" and exposure for
14 days:
SSL = 1 x 1(T6 4- (3.39 x 10'9 * 120 * 1 x l(T3* 14 * 1) = 176 pCi/g (-6,500 Bq/kg)
3-9
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Table 3.5. Soil Screening Levels for Ingestion of Ra-226 in Soil
Exposure
Frequency
(days)
1
14
30
52
140
350
Target Lifetime Cancer Risk
5 x 1(T4
1 x 10 4
5 x 10 5
1 x 10 5
SxlQ-6
1 x 10 6
Concentration of Ra-226 (pCi/g)
1.23E+06
8.78E+04
4.10E+04
2.36E+04
8.78E+03
3.51E+03
2.46E+05
1.76E+04
8.19E+03
4.73E+03
1.76E+03
7.02E+02
1.23E+05
8.78E+03
4.10E+03
2.36E+03
8.78E+02
3.51E+02
2.46E+04
1.76E+03
8.19E+02
4.73E+02
1.76E+02
7.02E+01
1.23E+04
8.78E+02
4.10E+02
2.36E+02
8.78E+01
3.51E+01
2.46E+03
1.76E+02
8.19E+01
4.73E+01
1.76E+01
7.02E+00
Figure 3.5. Relationship between Exposure Frequency, Radium Concentration,
and Target Lifetime Cancer Risk from Soil Ingestion
Figure 3.5 is derived from Table 3.5. The x-axis is the activity concentration of Ra-226
in the uranium mine waste material, and the y-axis is the incremental lifetime cancer risk as a
result ofingestion of radium in the waste material for different exposure times.
1.E-03
1.E-04
1.E-05
1.E-06
-1 day exposure
-14 day exposure
-30 day exposure
-52 day exposure
-140 day exposure
-350 day exposure
1.E+00 1.E+01 1.E+02 1.E+03 1.E+04 1.E+05
Ra-226 Concentration (pCi/g)
1.E+06
1.E+07
Table 3.6. Soil Screening Levels for Ingestion of Th-232 in Soil
Exposure
Frequency
(days)
1
14
30
52
140
350
Target Lifetime Cancer Risk
5 xlO^1
1 xlO^1
5 x 10 5
1 x 10 5
5 x 10 6
1 x 10 6
Concentration of Th-232 (pCi/g)
1.25E+06
8.94E+04
4.17E+04
2.41E+04
8.94E+03
3.58E+03
2.50E+05
1.79E+04
8.34E+03
4.81E+03
1.79E+03
7.15E+02
1.25E+05
8.94E+03
4.17E+03
2.41E+03
8.94E+02
3.58E+02
2.50E+04
1.79E+03
8.34E+02
4.81E+02
1.79E+02
7.15E+01
1.25E+04
8.94E+02
4.17E+02
2.41E+02
8.94E+01
3.58E+01
2.50E+03
1.79E+02
8.34E+01
4.81E+01
1.79E+01
7.15E+00
3-10
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Figure 3.6. Relationship between Exposure Frequency, Thorium Concentration,
and Target Lifetime Cancer Risk for Soil Ingestion
Figure 3.6 is derived from Table 3.6. The x-axis is the activity concentration of thorium
in the uranium mine waste material, and they-axis is the incremental lifetime cancer risk as a
result ofingestion of thorium in the waste material for different exposure times.
1.E-03
1.E-04
1.E-05
1.E-06
-1 day exposure
-14 day exposure
-30 day exposure
-52 day exposure
-140 day exposure
-350 day exposure
1.E+00 1.E+01 1.E+02 1.E+03 1.E+04 1.E+05
"m-232 Concentration (pCi/g)
1.E+06
1.E+07
Table 3.7. Soil Screening Levels for Ingestion of Natural Uranium in Soil
Exposure
Frequency
(days)
1
14
30
52
140
350
Target Cancer Risk
5 xKT1
1 x 10 4
5 x 10 5
1 x 10 5
5 x 10 6
1 x 10 6
Concentration of Natural Uranium (pCi/g U-238)
6.43E+06
4.59E+05
2.14E+05
1.24E+05
4.59E+04
1.84E+04
1.29E+06
9.18E+04
4.29E+04
2.47E+04
9.18E+03
3.67E+03
6.43E+05
4.59E+04
2.14E+04
1.24E+04
4.59E+03
1.84E+03
1.29E+05
9.18E+03
4.29E+03
2.47E+03
9.18E+02
3.67E+02
6.43E+04
4.59E+03
2.14E+03
1.24E+03
4.59E+02
1.84E+02
1.29E+04
9.18E+02
4.29E+02
2.47E+02
9.18E+01
3.67E+01
3-11
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Figure 3.7. Relationship between Exposure Frequency, Uranium Concentration,
and Target Lifetime Cancer Risk from Soil Ingestion
Figure 3.7 is derived from Table 3.7. The x-axis is the activity concentration ofU-238
in the uranium mine waste material, and they-axis is the incremental cancer risk as
a result ofingestion of uranium in the waste material for different exposure times.
1.E-03
1.E-04
1.E-05
1.E-06
-1 day exposure
-14 day exposure
-30 day exposure
-52 day exposure
-140 day exposure
-350 day exposure
1.E+01 1.E+02 1.E+03 1.E+04 1.E+05
U-238 Concentration (pCi/g)
1.E+06
1.E+07
3.3.3 Risk from Inhalation of Radium, Thorium, and Uranium in Fugitive Dust
Windblown dust provides a pathway for radioactive materials to reach humans through
inhalation. The equation for inhalation from the Superfund SSG (EPA 2000b) is:
SSL =
•where:
TR
SFi * IRi * (1/PEF) * 1 x io3 * EF * ED * [ETO + (ETI * DFI)]
TR
SF,
= Target lifetime cancer risk (unitless)
= Inhalation Slope Factor (pCi"1)
= Inhalation Rate (m3/day)
IRi
PEF = Particulate Emission Factor (nrYkg)
1 x 103 = Conversion factor (g/kg)
EF = Exposure frequency (days/ year)
ED = Exposure duration (year)
ETO = Exposure time fraction, outdoor (unitless)
ETI = Exposure time fraction, indoor (unitless)
DFI = Dilution factor for indoor inhalation (unitless)
2.55 x id'8 Ra-226
1.92 x l(T7 Th-232
6.14 x id'8 U-natural
20
1.32 x IO9
350
1
1
0
NA
Using these parameters, the 350-day SSL for Ra-226 is 7,395 pCi/g (2.74 x IO5 Bq/kg), 985
pCi/g (3.64 x IO4 Bq/kg) for Th-232, and 3,070pCi/g (1.14 x IO5 Bq/kg) for natural uranium.
This applies to exposed individuals in the vicinity of the mine.
3-12
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3.3.4 Risk from Use of All-Terrain Vehicles (ATVs)
The recreational use of ATVs and dirt bikes in the western United States is very common. These
vehicles allow easy access to very remote areas, so the safety provided by a mine's remote
location is often negated. The soil screening levels for inhalation of dust resuspended during the
operation of ATVs are estimated from empirical data on emission of dust from unpaved roads.
A scoping scenario for this pathway was developed, as described below.
It is assumed that a rider of an ATV or other off-road recreational vehicle riders would
participate in the sport about 60 times a year (once a week plus additional days on vacations or
holidays). It is further assumed that an abandoned mine site would lie on his route, and that he
would cross the site twice on each ride, going and returning over the same route. The area of the
site is 463.5 hectares (ha), the average of the total disturbed areas of the 21 mines listed in
DOE/EIA 2000b, Appendix C. This is a bounding condition as the estimated size of an
abandoned mine is expected to be much less, on the order of two hectares (U.S. EPA 2006a).
The area is assumed to be circular, and the route to be along the diameter of the circle. The
riders have inhalation rates of 1.2 m3/h, the average rate for light activity. The vehicles travel at
an average speed of 40 mph. The airborne concentration of respirable dust, 5 mg/m3, is based on
the average of three measured dust concentrations at a height of 2 m taken at the side of a road
composed of dirt and crushed slag, during the passage of medium-duty vehicles (3-4 tons)
traveling at a speed of 15 mph (Cowherd et al. 1979). The dust had a mass-median diameter of
10-11 jim, and thus corresponds to the approximate range of respirable particles. As it happens,
this concentration is also equal to the OSHA protective exposure limit (PEL) for nuisance dust
set forth in 29 CFR 1910.1000, and thus constitutes a reasonable upper bound to the average dust
loadings that could be comfortably tolerated by the rider. The SSLs are calculated using the
preceding equation for inhalation of contaminated dust. The parameters that were changed for
the ATV scenario are presented below.
The daily inhalation rate of the rider while exposed to the dust on the mine site is calculated as
follows:
I 1 T" i
R. -J_N
71
•where:
IR: = inhalation rate during = 0.0906
exposure(m3/d)
Ih = inhalation rate for light = 1.2
activity(m3/h)
As = Area of site (m2) = 4.635 x 106
v = speed of vehicle(40 mi/h) = 64,374 m/h
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The emission factor is simply the reciprocal of the dust loading, and is evaluated as follows:
PEF = emission factor
= 1/x = 0.2 nrVmg = 2 x 105 nrVkg
3
X = concentration of respirable dust = 5 mg/m
EF = Exposure frequency = 60 d/y
ETO = Exposure time fraction, outdoor = 1
Based on these values, the SSLs calculated for this scenario are:
Soil Screening Level
Radionuclide pCi/g Bq/kg
Ra-226 1,445 5.35E+04
Th-232 192 7.12E+03
U-natural 600 2.22E+04
3.4 Other Recreational Use Scenarios
Other recreational use scenarios were considered as part of the present analysis. These
include swimming, boating, fishing, and hunting, along with the consumption of on-site fish
and game. These scenarios are either unlikely to occur, or would be an insignificant
component of the risk, as reviewed in an EPA study (1983b). This study addressed related
scenarios for nearby residents [within 1 mile (1.6 km)] of the mines, including cattle grazing
and crop ingestion, as discussed below.
Although the pit lake at the Yazzie-312 Mine was used for swimming by local residents, the
lake was drained and filled in as part of the remediation of the mine site after 40 years of
abandonment. The number of other abandoned uranium mines with pit lakes is unknown.
However, swimming, through water immersion and ingestion pathways, contributes little
total dose (< 10 mrem or < 10"1 mSv) or risk. Estimates of risk from swimming are provided
in Appendix 1. Fishing is not considered in this analysis. Pit lakes, being artificial and not
connected to any natural bodies of water, are assumed to be devoid offish or expected to
contain minimal fish populations.
The majority of mine sites found in the uranium location database are typically in an arid
environment that does not readily support plant life unless irrigated. In such arid environments,
the overburden or protore piles are not expected to be able to provide much forage for animals,
especially if they are covered with a desert varnish. In addition, the size of the abandoned mine
sites would typically be relatively small and thus provide little forage for game animals.
Consequently, any game taken on a mine site would be expected to have obtained most of its
forage elsewhere. The meat from such game is thus not expected to be significantly
contaminated with TENORM from a mine site.
3.5 Metals in Uranium Mines
Metals and other minerals of commercial value frequently occur in the same ore deposits with
uranium (See Volume I, Chapters I and II, U.S.EPA 2006a) and, in some cases, it is economical
to mine them together. The most common commodities associated with uranium in the BASINS
3-14
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MAS/MILS (Mineral Availability System/Mineral Industry Location System) database are
phosphate, vanadium, gold, and copper. U.S. EPA (1999) provided an extensive review of
TENORM contamination, including uranium, associated with copper mines in Arizona.
However, numerous other commodities are associated with uranium, including antimony,
molybdenum, fluorine, rare earths, thorium, lead, mica, tantalum, and beryllium. For example,
in Colorado, 83 of 2,304 records had gold associated with uranium, and 10 had silver as a
secondary commodity; 38 records listing vanadium as a primary commodity also listed uranium
as a secondary commodity; and vanadium was listed as the primary or secondary commodity
with uranium in over 2,000 of the records. While multiple metals are associated with uranium
mines, limited information is available to determine the concentrations of the metals at the
different sites.
The 1983 EPA report to Congress indicated that at uranium mines, no adverse effects were
expected from nonradiological constituents because of the low airborne concentrations, with the
exception of fugitive dusts from operating mines (U.S. EPA 1983 a, b, and c). Nevertheless,
mining in general in the West has been known to generate problems with heavy metal
contamination in sediments and water, and some mines are Superfund sites (U.S. EPA 200Id).
3.5.1 Risk from Exposure to Arsenic
Arsenic, a carcinogen, is a metal of special concern. This naturally occurring metal may be a
common contaminant in uranium mine wastes. The presence of arsenic in extremely high
amounts in soils, as well as in the water, posed a significant risk at the White King/Lucky Lass
uranium mines. In the study (Portage Environmental 2005) of the Riley Pass Uranium Mines in
Harding County, South Dakota, arsenic was considered to be "the primary risk driver." The
primary exposure scenario at that site also involved recreational users of the site. The following
equation is used to estimate the lifetime cancer risk from ingestion of arsenic:
Arsenic SSL = TR * AT * 365
SF0 * ID'6 * EF * IF
where:
TR = Target lifetime cancer risk Variable
AT = Averaging time (years) 70
SF0 = Slope factor for arsenic (mg/kg-d)"1 1.5
EF = Exposure frequency (days/year) Variable
365 = Conversion factor (days/y)
1(T6 = Conversion factor (kg/mg)
IFSoii/adj is the ingestion factor (age-adjusted), in units of mg y kg"1 d"1. Because the recreational
use of the mine site is assumed to be episodic—it would occur for a limited period of time during
a given year—the limiting exposure would be to a child.4 Employing the data for a child, 0-6
4 The risks to a child were calculated for this chemical carcinogen because the expression for the ingestion
factor is age dependent. This is unlike the calculation of risks from radionuclides, where the reference slope factors
calculated by EPA are age adjusted.
3-15
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years old, the ingestion factor is given by the following expression, modified from the expression
for the residential scenario in U.S. EPA 1996b:
IF soil/a
il/age 1-6
BW
age 1-6
where:
IR
Soii/agei-6
soil ingestion rate of child (mg/d) 200
body weight of child (kg) 15
-v-5
The results are presented in Table 3.8. For a target lifetime risk of 5 x 10" and an exposure of
7 days/year, the arsenic soil screening level for children would be -8,250 mg/kg. The lowest
SSL is 3 mg/kg for the 350-day exposure at Ix 10"6 target risk. For perspective, for the White
King/Lucky Lass Superfund site, arsenic concentrations in surface soil were 769 mg/kg and
12 mg/kg, respectively, while background arsenic soil concentrations in the area were ~4 mg/kg.
The Riley Pass Uranium Mines arsenic average concentrations were over 500 mg/g (Portage
Environmental 2005). Although an occasional visitor to these sites does not incur much risk
from arsenic, it could pose a problem for those who frequent the sites.
Table 3.8. Target Lifetime Cancer Risk for Ingestion of Arsenic by Children Up to
6 Years Old
Exposure
Frequency
(days/year)
1
5
7
14
350
Target Lifetime Cancer Risk
5 x 10 5
1 xKT1
5 x 10 5
1 x 10 5
5 x 10 6
1 x 10 6
Soil Screening Level for Arsenic (mg/kg)
638,750
115,387
82,419
41,210
1,648
115,387
23,077
16,484
8,242
330
57,694
11,539
8,242
4,121
165
11,539
2,308
1,648
824
33
5,769
1,154
824
412
16
1,154
231
165
82
3
3.5.2 Risk from Drinking Mine-Contaminated Water
In addition to their potential to pose health risks on the site, uranium mines and their wastes can
affect surface or groundwater. For example, the pond in the mining pit could be contaminated
with radionuclides or metals, which would make the pond an exposure pathway. In addition, the
overburden (or protore) waste materials could leach into the ground and move into the
groundwater below. Material could also be physically transported from the waste piles by runoff
or wind (see discussion and data on theYazzie-312 Mine in Volume I, Chapter 3, U.S. EPA
2006a). In another scenario, the mine workings could intersect and contaminate groundwater.
There are multiple scenarios in which people could drink water contaminated from unreclaimed
uranium mining operations. For a recreational user of the site, the exposure may be short-term
from a spring, stream, or pond. Others could have lifetime exposure due to proximity to a
uranium mine. Also, someone who does not live on contaminated property could be exposed to
radionuclides from communal wells, which occurs on the Navajo Reservation in the Four
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Corners area (U.S. EPA and USAGE 2000). The radionuclides in groundwater can be due to
contamination from mining activities or from high natural background sources, including the
uranium ore body exploited by the mining operation. However, many large uranium mining
operations have reported problems of groundwater contamination (U.S. DOE/EIA 2000b,
Appendix C).
EPA's 1983 Report to Congress studied concentrations in, and risks from, waters discharged
from active mines to surface waters. The authors estimated that an insignificant health risk
accrues to populations from waterborne radionuclides due to water discharges from an average
existing active mine (U.S. EPA 1983b). However, the report acknowledged that some
abandoned underground mines were probably discharging contaminated waters into streams and
shallow aquifers, and the data were insufficient to determine the health risks from drinking the
water. Furthermore, due to a lack of data, the authors could not determine the health hazard to
individuals who drink from contaminated surface or underground sources. However, Volume I
of this study (U.S. EPA 2006a) reports on concentrations of radionuclides in ponds and streams
associated with open pit uranium mines, and case studies where shallow groundwater and surface
springs or streams were contaminated by uranium mine discharges.
EPA has established maximum contaminant levels (MCLs) for several radionuclides in
community water supplies that serve more than 25 customers (Table 3.9). These MCLs can be
used to help establish soil cleanup levels at a site. The SSG approach is used to conservatively
identify a soil level that would prevent a site contaminant from attaining the MCL in
groundwater. The drinking water MCL for uranium is based primarily on kidney toxicity, rather
than radiological effects.
Table 3.9. Radionuclide Maximum Contaminant Levels for Public Water Supplies
EPA has established drinking-water maximum contaminant levels for
several radionuclides. Although these values are for public water supplies,
the Superfundprogram has applied them to site cleanups.
Radionuclide
Uranium
Man-made beta/photon emitters
Alpha emitters (excluding radon and uranium)
Combined radium-226 and radium-228
Maximum Contaminant Level
30ug/L
4 mrem/y (0.04 mSv/y) to whole body or any organ
15 pCi/L (555 Bq/m3)
5 pCi/L (185 Bq/m3)
Source: Modified from EPA 2000c.
While the number of people who drink water contaminated by uranium mining activities is
unknown, it is possible to calculate an individual lifetime risk for various concentrations of
radionuclides. The numbers in Table 3.10 are based on the risk calculations presented in the
technical support document for the radionuclides in drinking-water regulation (U.S. EPA 2000d).
3-17
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Table 3.10. Radionuclide Mortality and Morbidity Risk Coefficients'
While the number of people who drink water contaminated by uranium mining
activities is unknown, it is possible to calculate an individual lifetime risk for
various concentrations ofradionuclides.
Radionuclide
Radium3
Th-232b
Ra-228b
Th-228b
Ra-224b
Uranium0
Gross alphad
Mortality Risk Coefficient
per pCi Consumed
5.66 x 1(T10
6.92 x 10'11
7.40 x 1(T10
6.73 x ID'11
1.01 x 10"10
4.4 x KT11
1.14 x 10~10
Morbidity Risk Coefficient
per pCi Consumed
8.03 x 1Q-10
1.01 x 10~10
1.04 x 1Q-9
1.07 x 1Q-10
1.67 x 10~10
6.81 x 1Q-11
1.83 x 10"10
a Average weighted by relative prevalence of Ra-226 and Ra-228
b Principal members of Th-232 decay chain
0 Arithmetic average for natural uranium isotopes: U-234, U-235, U-238
d Average weighted by relative prevalence of Ra-224 and Ra-226
Source: U.S. EPA 2000d.
The equation used to calculate the risks from these radionuclides is:
Risk = Concentration (pCi/L) * Risk coefficient * Water consumed (L/day) * Exposure
frequency (days/year) * Number of years
Figures 3.8-3.10 depict the risks from radium, gross alpha, and uranium for (1) 70 years of
exposure, 365 days a year, drinking 2 liters of water a day from the contaminated source,
representing lifelong consumption; and (2) 10 years of exposure, 14 days a year, drinking 2 liters
a day, representing recreational consumption. For the first situation, long-term exposure
produces risks of up to 1 x 10~3 for some of the higher concentrations. However, for the long-
term recreational user consuming contaminated water, the lifetime risk remains less than 6 x
10'6.
Morbidity risk is the risk of getting cancer, and mortality risk is the risk of dying from cancer.
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Figure 3.8. Cancer Risks from Lifetime and Recreational Exposures to Radium in
Drinking Water: 70 Years, 365 Days/Year & 10 Years, 14 Days/Year Exposure
Long-term exposure to radionuclide-contaminated water produces risks
of up to 1 x 1(T3 for some of the higher concentrations. However, the risk
from long-term recreational consumption is less than 6x1 (T6.
IE-OS --
1.E-07
» Mortality Risk, 70 y
• Morbidity Risk, 70 y
A Mortality Risk, 10 y
• Morbidity Risk, 10 y
1 10
Radium, pCi/L (MCL = 5 pCi/L)
100
Figure 3.9. Cancer Risks from Lifetime and Recreational Exposures to Gross Alpha in
Drinking Water: 70 Years, 365 Days/Year & 10 Years, 14 Days/Year Exposure
Long-term exposure to radionuclide-contaminated water produces risks of up
to 1 x 1(T3 for some of the higher concentrations. However, the risk from
long-term recreational consumption is less than 6 x 1 (T6.
1.E-08
-Mortality Risk,
-Morbidity Risk
-Mortality Risk,
-Morbidity Risk
70 y
1 10
Gross Alpha, pCi/L (MCL = 15 pCi/L)
3-19
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Figure 3.10. Cancer Risks from Lifetime and Recreational Exposures
to Uranium in Drinking Water:
70 Years, 365 Days/Year and 10 Years 14 Days/Year Exposure
Long-term exposure to radionuclide-contaminated water produces risks of up
to 1 x 1(T3 for some of the higher concentrations. However, the risk from
long-term recreational consumption is less than 6x1 (T6.
1.E-03 -,
1.E-04 --
J.E-05 --L
1.E-08
1.E-09 4
• Mortality Risk, 70 y
• Morbidity Risk, 70 y
A Mortality Risk, 10 y
• Morbidity Risk, 10 y
1 10
Uranium, pCi/L (MCL = 30 ug/L a 21 pCi/L)
100
Table 3.11 estimates the potential lifetime cancer risk from radionuclides in the shallow Yazzie-
312 Mine pit water (Panacea 2002), at concentrations measured before the pit was remediated.
At these levels, long-term consumption of drinking water containing the radionuclides would be
a significant health risk, but shorter-term exposures would not.
Table 3.11. Lifetime Risks Estimated from Drinking Unremediated
Yazzie-312 Mine Pit Water
While long-term consumption of drinking pit water from the Yazzie-312 Mine
posed a significant health risk, shorter-term exposures would not.
Contaminant
Total Radium
Total Uranium
Gross alpha3
Total Risk
Average
Concentration
2.3 pCi/L
173 pCi/L
84pCi/L
-
Exposure Duration
70 Years, 365 Days/Year
Mortality
7 x 1(T5
4 x 1(T4
5 x 1(T4
9.7 x 1(T4
Morbidity
9 x 1(T5
6 x 10"1
8x KT*
1.5 x KT4
10 Years, 14 Days/Year
Mortality
4 x 1(T7
2 x 1(T6
3 x 1(T6
5.4 x IQ-6
Morbidity
5 x 1(T7
3 x KT6
4x IQ-6
7.5 x IQ-6
a Without uranium and radon
Note: Other periods of exposure may be of interest, such as a 30-year period, often used in
Superfund calculations. Since the relationship between concentration and risk is linear, a ratio
can be used to calculate risks at different time periods. To estimate the risk for 30 years of
exposure, divide the 70-year risk number by 2.33 (70 y/30 y). Arsenic was measured in the pit at
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an average concentration of 55 |ig/L, just over the MCL in effect in 2005 and five times higher
than the 10 |ig/L MCL that became effective in 2006. In calculating the risk from arsenic in the
water, the following equation and defaults from Superfund Risk Assessment Guidance (U.S.EPA
1989b) were used:
Target lifetime cancer risk, TR = (SF0 * C * IRW * EF * ED) / (BW * AT * 365 days/year)
where:
SFO = Slope factor for arsenic (mg/kg-d)"1 1.5
C = Pit water arsenic concentration (mg/L) 0.055
IRW = Daily water ingestion rate (L/day) 2
EF = Exposure frequency (days/y) 350
ED = Exposure duration (years) 30
BW = Body weight (kg) 70
AT = Averaging time (years) 70
Using the default values listed above, we estimate the risk for drinking arsenic to be about 1 in
1,000, assuming 30 years of exposure (1.5 * 0.055 * 2 * 350 * 30 /[70 * 70 * 365] = 1,732.5 /
1,788,500 = 9.7 x 10~4 ~ = 1 x 10~3). For an exposure of 14 days/year for 10 years, the risk
estimate is 1.3 x io~5 or ~ = 1 x io~5. Thus, the pit water at the Yazzie-312 Mine could have
posed a high risk from both radionuclides and arsenic, if the water were consumed over long
periods of time.
The 1983 EPA report to Congress also reported Wyoming and New Mexico field studies of
trace elements and radionuclides from inactive mining areas at off-site locations (U.S. EPA
1983c). In both cases, precipitation is seasonal and adjacent streams are dry much of the year.
The general observations were that concentrations of Ra-226 and U-238 from spoils piles and
in stream channels decreased rapidly with distance from the mines. However, the migration of
trace metals did not show as distinct a trend. The transport processes were believed to be wind
erosion and sheet erosion from cloudbursts, and they appeared to move mine spoils material up
to 2,000 feet (-600 m) in 10 years. Preliminary data from recent sampling by Burghardt
(2003) at several uranium mines have identified decreasing uranium and arsenic
concentrations from the toe of the pile to background levels within several hundred meters.
3.6 Migration of Uranium Waste into Groundwater
Chemical and physical processes can enhance or retard the movement of the contaminants into
and through an aquifer. Infiltration of water into soil is an example of a physical process, while
partitioning of the contaminant between the soil and water is an important chemical process
(which gives rise to the soil-water distribution coefficient, Kd). On the Colorado Plateau, where
many uranium mines are located, the dry climate limits the available water for transporting the
radionuclides and for drinking. Much of the precipitation is lost to evapotranspiration, thus
limiting the infiltration, although high intensity precipitation events may contribute to increased
infiltration at times. In large parts of the Colorado Plateau, the only usable water available in
quantity is from groundwater (U.S. EPA 1983b), particularly in relatively deep confined
aquifers, but near-surface aquifers are present in some areas. The impact of small surface
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uranium mines on most of the groundwater in this region is expected to be minimal. As an
example described in more detail below, drilling and sample analysis of a groundwater aquifer
located under theYazzie-312 pit lake found no direct communication or correlation of water
chemistry with the overlying lake (Panacea 2002). However, underground mines that intersect
an aquifer could contaminate the aquifer, as could large surface mines with deep pits. Also, in
areas with greater precipitation or near-surface unconfmed aquifers, including higher elevations
in the Colorado Plateau, contaminated water may more easily reach the groundwater, where it
could be transported and pose significant cancer risks to people who obtain their drinking water
from the aquifer.
3.7 Mobility of Uranium and Radium through Groundwater
EPA's Soil Screening Guidance for Radionuclides is one method that can be used to
conservatively estimate the potential for a radionuclide to move into groundwater and to develop
a general understanding of the resulting health risks (U.S. EPA 1996a, 2000b). This approach,
which is modified as site-specific conditions are understood, relies on the use of distribution
coefficients. This generalized approach is useful for this scoping analysis, since many
potentially different site conditions and parameters would need to be considered otherwise.
Indeed, for an individual site it is important to gather site-specific information before decisions
are made for the particular site. A goal in establishing a soil contaminant concentration is to
avoid future contamination of groundwater above the maximum concentration level (MCL)
established for the contaminant in potable water. This general approach is also applicable to
metals, but the focus here is on key radionuclides.
In calculating the SSL, in pCi/g, for groundwater the equation is:6
Ct=Cw*(l x 10-3)*(Kd + 0w/pb)
•where:
Ct = Total concentration in soil (pCi/g) 1.5
Cw = Target concentration in leachate (pCi/L) element-specific,
—20 pCi/L for uranium
1 x 1CT3 = Conversion factor (kg/g)
Kd = Soil-water partition coefficient (mL/g) Element-specific
9W = Water-filled porosity (unitless) 0.3
pb = Dry soil bulk density (kg/L) 1.5
CW; the target concentration in the leachate, is derived by multiplying the MCL by a dilution
factor of 207, the soil-water partition coefficient is specific to the contaminant of concern, and
default values are used for the unitless water-filled porosity, and the dry soil bulk density (U.S.
EPA 2000b, Equation 6).
6 There are additional variations on this equation, including a mass-limit version that includes infiltration.
More detail on this and alternative ground-water transport models are discussed in the EPA Soil Screening Guidance
Technical Background Document (U.S. EPA 1996b).
7 Default value from U.S. EPA 1996b, Part 2.
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The SSL generally corresponds to a risk of 1 x 1CT6, and the actual cleanup goal is modified from
there; however, for groundwater it is based on achieving the MCL. Tables 3.13 and 3.15 provide
the soil screening levels for uranium and radium, respectively, assuming varying soil-water
partition coefficients with the target concentration as the MCL. Thus, Cw= 600 |ig/L of uranium
for an MCL of 30 |ig/L [or -20 pCi/L using the uranium specific conversion 0.67 pCi/|ig (U.S.
EPA2000d)].
In using this equation, it is important to note the following simplifying assumptions applied in
the Soil Screening Guidance methodology. The assumption that soil contamination extends from
the surface to the water table adds a conservative element to the equations, since this condition
would be uncommon in the Colorado Plateau, where the depth to water can be tens of meters or
more, precipitation is limited, and the aquifer is typically confined. However, in other areas
where water is near the surface, this list of simplifying assumptions may not be as conservative.
Simplifying Assumptions for the Migration of Radionuclides to Groundwater
• The source is infinite (i.e., steady-state concentrations will be maintained in
groundwater).
• Contaminants are uniformly distributed throughout the zone of contamination.
• Soil contamination extends from the surface to the water table (i.e., adsorption sites
are filled in the unsaturated zone beneath the area of contamination).
• There is no chemical or biological degradation in the unsaturated zone.
• Equations do not account for radioactive decay.
• Equilibrium soil/water partitioning is instantaneous and linear in the contaminated soil.
• The receptor well is at the edge of the source (i.e., there is no dilution from recharge
downgradient of the site) and is screened within the plume.
• The aquifer is unconsolidated and unconfmed (surficial).
• Aquifer properties are homogeneous and isotropic.
• Chelating or complexing agents are not present.
• No facilitated transport (e.g., colloidal transport) of inorganic contaminants occurs in the
aquifer.
Source: U.S. EPA 2000b.
3.7.1 Uranium
Depending on the environmental conditions, uranium can be mobile enough to leach into and
move through groundwater, especially in the oxidizing conditions at low pH levels that are
present in acid mine drainage. Uranium tends to be relatively immobile under reducing
conditions. Table 3.12 illustrates the range of uranium mobility as a function of pH, and Table
3.13 indicates the soil screening level above background needed to achieve the MCL of 30 |ig/L.
A higher partition coefficient (Kd) means that the movement of uranium would be slower relative
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to the movement of water. In the White King monitoring wells, the ore pile area pH was
between 4.2 and 6.9, the mine spoil area pH was between 5.6 and 7.0, and the pH in unaffected
areas in the vicinity was between 6.3 and 7.7 (USFS 1991).
Although no Kds were calculated at the White King site, no downgradient uranium was detected,
even though pore water samples in the protore stockpile were over 27,000 pCi/L (106 Bq/m3).
The overburden stockpile activities were less than 18 pCi/L (670 Bq/m3), with a concentration of
only 75 pCi/L (2775 Bq/m3) immediately under the protore stockpile (Weston 1997). Thus, the
uranium appears to be immobile, with a high Kd, at this site. Radium, in the form of radium
sulfate, apparently had not migrated at all. In the 1983 EPA report to Congress, soil profiles
obtained at a uranium mine in Wyoming also showed no downward migration of radionuclides
(U.S. EPA 1983c).
Table 3.12. Look-up Table for Estimated Range of Kj Values for Uranium Based on pH
Kd (mL/g)
Minimum
Maximum
pH Levels
3
<1
32
4
0.4
5,000
5
25
160,000
6
100
1,000,000
7
63
630,000
8
0.4
250,000
9
<1
7,900
10
<1
5
Source: U.S. EPA 1999
Table 3.13. Soil Screening Values for Uranium as a Function of Ka
Uranium (MCL = 30 pg/L ~ 20 pd/L ~ 760 Bq/m3)
Target
Concentration
30 ug/L8
Assumed Partition Coefficient
(Ka) (L/kg)
1
10
25
50
100
Soil Screening Values (pCi/g Above Background Levels) Resulting in Groundwater
Target Concentration Using the Groundwater Soil Screening Approach
0.5
4
10
21
41
In contrast to the White King/Lucky Lass site, at Midnite Mine the groundwater indicator map
from preliminary investigation work (U.S. EPA 2003c) plots concentration exceedances for
shallow and deep wells. Uranium and other metals have been detected in several of the
downgradient alluvial wells and in a couple of shallow bedrock aquifer wells adjacent to a pit
and a stockpile.
The Yazzie-312 Mine has no near-surface water table because of the dry Arizona climate. There
is a confined aquifer at 105 feet (32 m) below ground surface in the southern part of the site in a
sand-and-gravel unit, with a static water level of 27 feet (8.2 m) below ground surface. This unit
was thought to be part of a former alluvial channel, since no water was found in another well
north of the mine. Since only 2.6 pCi/L (96 Bq/m3) uranium was in the well water while 173
pCi/L (6,400 Bq/m3) was in the pit water, the interpretation is that the pit water is not
contributing to the radionuclide concentration in the aquifer. On the other hand, Longsworth
! Conversion factor for naturally occurring uranium from ug/L to pCi/L (U.S. EPA 2000d): 0.67 pCi/ug
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(1994) measured shallow groundwater in the vicinity of mines in the Monument Valley area of
Arizona and Utah with significant levels of uranium, radium, and radon (up to 14,000 pCi/L U-
238, 110 pCi/L Ra-226, and 250,000 pCi/L of Rn-222). The impact on groundwater from surface
and near-surface uranium mines would appear to be highly dependent on local geological and
hydrological conditions.
3.7.2 Uranium Plume Migration
In a review of uranium plumes in groundwater from natural analogues, in-situ leaching
operations, and uranium mill tailings sites, Colon et al. (2001) identified a "clear and reasonably
consistent picture of [uranium] plume behavior" in which plumes appear to reach a steady-state
condition: the plumes rarely exceed 1.25 miles (2 km) in length and exhibit natural attenuation
under different circumstances, with the low-pH in-situ leaching process contributing to the
greatest plume distances. Of the natural analogues, the maximum axial9 plume length was 1
mile (1.6 km) from the Oklo uranium deposit that acted as a natural reactor ~ 2 billion years ago.
If this attenuation were to hold true at uranium mines, the distance of influence on uranium
transport from an abandoned uranium mine (in the absence of added acids) in the groundwater
could be less than 1.25 miles (2 km). Fracture networks, facilitated (colloid) transport, or other
site-specific characteristics may act to limit this attenuation.
3.7.3 Radium
Information on radium soil-water distribution coefficients is less common, but radium Kd values
that span a large range are found in the literature. U.S. EPA (2004) cautions the reader that
many of the high values are suspect, because they may be the result of co-precipitation of radium
with other ionic species, rather than absorption of radium itself. One EPA study indicates that
very little radium is available for transport, and strong acids were necessary to extract the radium
(DeLaune et al. 1996). Tachi et al. (2001) calculated Kds of 102-104 mL/g for bentonite clays
with a dependence on pH. U.S. EPA (2004) mentions one study of four sandy soils from Utah
with a range of radium Kd values from 214 to 354 ml/g for pH that varies between 7.6 and 8.0.
EPA (2004) confirms the paucity of Kd data, stating: "Development of Kd look-up tables for
radium is not possible given the minimal number of adsorption studies." U.S. EPA (ibid.) then
goes on to suggest the use of the Kd table for strontium presented by U.S. EPA (1999, Vol. 2) as
general guidance for radium. This table is reproduced as Table 3.14. Table 3.15 provides SSLs
for radium as a function of Kd for a range of KdS from 1 to 500.
9 Along the center line of the contamination where the greatest concentration would be expected.
3-25
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Table 3.14. Relationship Between pH Levels and Strontium Mobility as a Surrogate for
Radium
Look-up table for estimated range ofK^valuesfor strontium as a surrogate for radium based
on clay content and pH. [Tabulated values pertain to systems consisting of natural soils (as
opposed to pure mineral phases), low ionic strength (< 0.1M), low humic material
concentrations (<5 mg/l), no organic chelates (such as EDTA), and oxidizing conditions.]
Kd (ml/g)
Minimum
Maximum
Soil Clay Content (wt.%)
<4%
pH
<5
1
40
5-8
2
60
8-10
o
3
120
4 - 20%
pH
<5
10
150
5-8
15
200
8-10
20
300
20 - 60%
pH
<5
100
1,500
5-8
200
1,600
8-10
300
1,700
Table 3.15. Soil Screening Values for Radium as a Function of
Radium (MCL = 5 pd/L [185 Bq/m3])
Target
Concentration
5pCi/L
Assumed Partition
Coefficient
1 10
25
50
100
500
Soil Screening Levels
Concentration Values (pCi/g) Above Background Resulting in
Target Groundwater Concentration
0.12 1.0
2.5
5
10
50
3.7.4 Potential for Groundwater Infiltration and Contamination
From Figure 3.11 below, the general annual precipitation range for the Colorado Plateau area is
5-15 inches (13-38 cm). This area also has high evapotranspiration rates. The 1983 EPA Report
to Congress (U.S. EPA 1983 a, b, and c) estimated that about 97 percent of the precipitation was
lost to evapotranspiration. Evaporation tables indicate that the general area experiences greater
than 75 inches (190 cm) of evapotranspiration annually. Thus, very little precipitation infiltrates.
The Maxey-Eakin empirical method for estimating recharge in the southwest (Maxey and Eakin,
1949) assumes recharge would be zero if precipitation was less than 8 inches (20.3 cm/y), and
only 3% if precipitation was between 8-12 inches (20.3-30.4 cm/y). Flint et al. (2002) modified
this for areas of shallow soil, so that the minimum precipitation threshold for recharge to occur
was 10 cm/y.
Thus, for 15 inches/y (38.1 cm/y) of precipitation, or the maximum of the range of annual
precipitation in the Colorado Plateau, the average recharge would be -0.5 inches/y (1.1 cm/y).
If this average value is assumed to be a simple velocity estimate to an aquifer and assuming no
retardation, it would take hundreds of years or longer to reach an aquifer at depth. Doubling the
velocity (i.e., infiltration rate) would reduce the travel time by one-half. Thus, abandoned
uranium mines in the proximity of shallow aquifers may contaminate the aquifer within tens of
years, but this process would take longer for the deeper mines. This simple analysis suggests
that the abandoned uranium mines that don't intersect aquifers pose a greater immediate risk
from surface pathways and use than from the groundwater pathway.
3-26
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Site-specific characteristics such as precipitation, depth to water, soil characteristics (e.g.,
permeability or pH), or presence or absence of fractures, would dictate the actual infiltration,
potential recharge and potential to contaminate an aquifer, and the time frame over which such
contamination could occur. Once the radionuclide enters an aquifer, its transport would be
dependent on several site-specific factors—including the aquifer's permeability, water velocity,
and chemistry (e.g., pH)—that affect retardation. Although much of the discussion in this section
has focused on radionuclides, similar concepts apply if metals are also present at a site.
Figure 3.11. Average Precipitation (inches/year) for the Western United States
The Colorado Plateau, where many of the uranium mines are or were located, is a
region characterized, in general, by low precipitation and high evapotranspiration.
Legend
[ \ Colorado Plateau Boundary
Avg. Precipitation
Inches/Yr
| 0-5
5-10
Map Derived From National Atlas Data
3-27
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3.8 Consideration of Multiple Exposure Pathways
The fundamental criterion for applying the SSLs to a single exposure pathway is that C; <
SSLi; k; that is, the concentration of pollutant i, Q, is less than the SSL for pathway k, SSKijk.
This implies that, for multiple exposure pathways, the SSL should be reduced to account for
additive contributions to the pollutant intake from these additional pathways such that:
+... Ci/SSLu_ < 1.0.
Dividing both sides by the concentration term C; and inverting the equation gives:
Q < I/ [1/SSLU+ l/SSLi;2+ l/SSLi;3+ ... 1/SSLU].
The term on the right side may be viewed as a multi -pathway SSL. Tables 3.16, 3.17, and 3.18
show the application of this methodology to the external exposure, soil ingestion, and inhalation
of fugitive dust pathways for the on-site exposure scenario discussed earlier in this chapter. The
SSLs for external exposure and soil ingestion are listed in Tables 3.2 - 3.7. The calculation of
SSLs for the inhalation of fugitive dust is discussed in the text. The risk from recreational use of
off-road vehicles is not included, because the riders of these vehicles will not, in general, be the
same individuals exposed in the other on-site scenarios. Likewise, the consumption of drinking
water from a well would affect residents on or off the site many years in the future, after the
activity has percolated into the groundwater. These would not be the same individuals exposed
to the radioactivity in the surface soil due to recreational use of the site at the present time.
However, for a particular site the risk from drinking surface or near-surface water could be added
to risks from the other pathways. However, risk estimates conducted for this chapter indicate
that the risks in the recreational scenario from external exposure are much greater than from
drinking water contaminated with radionuclides.
A comparison of the multi-pathway SSLs for Ra-226 listed in Table 3.16 with the SSLs for
external exposure shown in Table 3.2 shows a difference of about 1.75%; thus, the external
exposure pathway for this nuclide and its progeny is dominant, and the other pathways make
minor contributions to the total risk. A similar comparison for Th-232, using the SSLs in
Tables 3.3 and 3.17 shows an even smaller difference— about 1 .2%— indicating that the external
exposure pathway is dominant for this nuclide and its progeny. This is not the case for natural
uranium; although external exposure constitutes over 86% of the risk, soil ingestion makes a
significant contribution. The inhalation of fugitive dust makes a minor contribution.
Figures 3.12 through 3.14 portray the same data in graphical form.
3-28
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Table 3.16. Multi-pathway Soil Screening Levels for Ra-226
Exposure
Frequency
(days/year)
1
14
30
52
140
350
Target Lifetime Cancer Risk
SxKT*
1 X 1Q-4
5 x 1Q-5
1 X 1Q-5
5 x 1Q-6
1 X 1Q-6
Concentration of Ra-226 (pCi/g)
21,116
1,508
704
406
151
60.3
4,223
302
141
81.2
30.2
12.1
2,112
151
70.4
40.6
15.1
6.03
422
30.2
14.1
8.12
3.02
1.21
211
15.1
7.04
4.06
1.51
0.603
42.2
3.02
1.41
0.812
0.302
0.121
Table 3.17. Multi-pathway Soil Screening Levels for Th-232
Exposure
Frequency
(days/year)
1
14
30
52
140
350
Target Lifetime Cancer Risk
SxKT1
1 x 1(T4
5 x 1(T5
1 x 1(T5
5 x 1(T6
1 x 1(T6
Concentration of Th-232 (pCi/g)
14,674
1,048
489
282
105
41.9
2,935
210
97.8
56.4
21.0
8.38
1,467
105
48.9
28.2
10.5
4.19
293
21
9.78
5.64
2.10
0.838
146.7
10.5
4.89
2.82
1.05
0.419
29.3
2.10
0.978
0.564
0.210
0.0838
Table 3.18. Multi-pathway Soil Screening Levels for Natural Uranium
Exposure
Frequency
(days/year)
1
14
30
52
140
350
Target Lifetime Cancer Risk
5 x 1(T4
1 x 1(T4
5 x 1(T5
1 x 1(T5
5 x 1(T6
1 xlQ-6
Concentration of Natural Uranium (pCi/g U-238)
751,392
53,671
25,046
14,450
5,367
2,147
150,278
10,734
5,009
2,890
1,073
429
75,139
5,367
2,505
1,445
537
215
15,028
1,073
501
289
107
42.9
7,514
537
250
144
53.7
21.5
1,503
107
50.1
28.9
10.7
4.29
3-29
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Figure 3.12. Multi-pathway Soil Screening Levels for Ra-226
1.E-03
1.E-04
1.E-05
1.E-06
-1 day exposure
-14 day exposure
-30 day exposure
-52 day exposure
-140 day exposure
-350 day exposure
1.E-01 1.E+00 1.E+01 1.E+02 1.E+03 1.E+04 1.E+05
Ra-226 Concentration (pCi/g)
Figure 3.13. Multi-pathway Soil Screening Levels for Th-232
1.E-03
1.E-04
1.E-05
1.E-06
-1 day exposure
-14 day exposure
-30 day exposure
-52 day exposure
-140 day exposure
-350 day exposure
1.E-02 1.E-01 1.E+00 1.E+01 1.E+02 1.E+03 1.E+04 1.E+05
Th-232 Concentration (pCi/g)
3-30
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Figure 3.14. Multi-pathway Soil Screening Levels for U-238
1.E-03
1.E-04
1.E-05
1.E-06
-1 day exposure
-14 day exposure
-30 day exposure
-52 day exposure
-140 day exposure
-350 day exposure
1.E+00 1.E+01 1.E+02 1.E+03 1.E+04
U-238 Concentration (pCi/g)
1.E+05
1.E+06
3-31
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4.0 RISK FROM URANIUM MINING WASTE IN BUILDING
MATERIALS
In general, building materials contain low levels of radioactivity. For example, the range of
natural uranium concentrations may average as low as about 0.5 ppm (0.34 pCi/g or 13 Bq/kg)
total uranium activity in sandstone building materials to as high as 5 ppm (3.4 pCi/g or
130 Bq/kg) in granitic building materials. Concrete and brick buildings are estimated to
contribute an average of about 10 mrem (0.1 mSv) annual effective dose equivalent (NCRP
1987) to the average person's background exposure to radiation. However, exceptions can occur
to this generalization, especially in buildings constructed with materials containing uranium
TENORM mine wastes. In the Grand Junction, Colorado area, thousands of homes and
properties were constructed using uranium mill tailings (U.S. EPA 1983a, b, c) in the past as a
source of construction sand, gravel, and clays. However, a number of homes have also been
built with materials that have been attributed to "uranium ore" that are not considered to be mill
tailings. In a 1972 EPA and Atomic Energy Commission (AEC) survey intended to locate
building materials contaminated with mill tailings, 537 buildings were identified where uranium
ore may have been the source of gamma-ray exposure anomalies (U.S. EPA 1983b):
We do not know to what extent the wastes from uranium mines have been removed
from mining sites and used in local and nearby communities. However, while
surveying in 1972 for locations with higher-than-normalgamma radiation in the
Western States to locate uranium mill tailings used in local communities, EPA
and AEC identified more than 500 locations where "uranium ore " was believed
to be the source of the elevated gamma radiation. The specific type of ore (mill-
grade, sub-ore, low-grade waste rock) was not determined as this was beyond the
scope of the survey. At some locations, however, surveyors attempted to
characterize the ore by using such terms as "ore spillage, " "ore specimens, "
"low-grade crushed ore, " or "mine waste dump material. " Some locations were
identified as sites of former ore-buying stations [U.S. EPA 1973].
Since it is unlikely that valuable mill-grade ore would have been widely available for off-site
use, we suspect that uranium mine waste (perhaps protore) may be the source of the elevated
gamma radiation levels at many of the locations where large quantities of ore material are
present.
About three-fourths of the 537 buildings were in Colorado and Utah, with the rest distributed
among several other states. Figure 4.1 identifies the localities from the 1972 survey that had at
least one building thought to have used "uranium ore" construction materials. Many of these
same localities also had additional anomalies attributed to either a radioactive source or natural
radioactivity. The original report that discusses the survey is unavailable, so it is not possible at
this time to determine the basis used for the attribution of the cause. Of the 53 localities with at
least one anomaly attributed to uranium ore, 20 are on or within approximately 25 miles (40 km)
of Bureau of Indian Affairs (BIA) Reservations. Without knowing the design of the study, it is
not possible to determine the statistical significance of the survey. Nevertheless, the survey does
indicate the potential problem of contaminated buildings in uranium mining areas, especially on
and around Tribal lands. EPA has provided support to Tribal authorities since that time to
identify buildings on Tribal lands constructed with uranium mine wastes.
4-1
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Contaminated buildings are among the mine waste issues that have been publicized regarding the
Navajo Nation. Although not specifically addressed herein, anecdotal information is amenable
to the methods and models for dose and risk estimates contained in this report. For mattresses
placed directly on a contaminated slab (reported in the Los Angeles Times on November 19,
2006), a geometrical variation would be applied to the analysis that follows. (As an example
involving other exposure pathways, for children who "dug caves in piles of mill tailings and
played in the spent mines," variations in the recreational scenario of Chapter 3 would be
applied.)
A specific case of the potential problem on Tribal lands is illustrated by hogans with elevated
radioactivity found in the Monument Valley area of Utah. In April 2001, EPA razed and
removed a building that had been used as a hogan (sacred home) by a Navajo family. As shown
in the photograph in Figure 4.2, the hogan was a small, one-room round structure with a concrete
slab for a floor and stucco walls, although the building originally had a dirt floor. Figure 4.3 is a
picture of another house taken from the vantage point of uranium mine workings.
Figure 4-1. Locations of Building Gamma Anomalies Due to Uranium Ore
from 1973 EPA-AEC Study
Building Gamma Anomalies
Due to Uranium Ore
Gamma Anomaly Locations
Gamma Anomalies < 25 Miles of B1A Land
Bureau of Indian Affairs Land
0 100 200
Source: U.S. EPA.
4-2
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Short-term gamma-ray exposure rates and radon concentrations were measured prior to the
demolition of the hogan (Sowder et al. 2001). Radiation exposures were between 370 jiR/h and
600 |iR/h. This is equivalent to doses in air of 325-525 |irad/h (-3-5 |iGy/h). (Typical indoor
background dose rates are in the range of 1.2-16 jirad/h [12-160 nGy/h]). Several stones in the
hogan exhibited levels of 1,000 jiR/hour on contact. Short-term indoor radon measurements
using multiple methods averaged 50-90 pCi/L (1,850-3,300 Bq/m3) under pseudo-closed
conditions. Outdoor exposure rates as high as 75 jiR/hour at 3.3 feet (1 m) from the structure
were observed. Stones used in the exterior construction produced exposure rates of 500-
1,000 |iR/hour. Inspection of the floor after demolition revealed that uranium ore had been used
as aggregate for the concrete. Apparently, the source of the sand and stones in the building
material was a nearby uranium mine or outcrop adjacent to the mine (possibly the Skylight
Mine). Other possibilities for the material include mine-waste material debris piles alongside
roads, such as the one in Figure 4.4, which is on Navajo Nation land. Readily available
construction materials, including clay, sand, gravel, cobbles, and boulders in above-ground piles,
make them attractive for houses, stoves, chimneys, and barbecues, and for stucco, cement for log
houses, driveways, walkways, and fill dirt.
Figure 4-2. Monument Valley Navajo Hogan
Monument Valley Navajo family hogan razed due to high gamma readings. Note the talus
in back, much of which originated from Skylight Mine on top of the mesa directly above.
Photograph by Andrew Sowder (U.S. EPA)
4-3
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Figure 4-3. Navajo Home in Proximity to Uranium Mine
This picture is another example of the proximity of some homes to uranium mines.
A New Mexico mine, now reclaimed, lies in the foreground of the picture, while the house
in the background was originally constructed with mine waste but has since be
reconstructed to remove the contaminated material.
Photograph by Loren Setlow (U.S. EPA)
Figure 4-4. Uranium Mine Debris Pile
Debris pile of uranium mine wastes just off a road on Navajo Nation land.
Photograph by Andrew Sowder, (U.S. EPA)
4.1 Building Materials Analysis
Given that some homes incorporate uranium mine waste building material, the question arises as
to the radium and uranium concentrations in these materials that would result in exposure levels
of concern. To identify potential gamma and radon exposures over a range of uranium and
4-4
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radium concentrations from contaminated concrete used as building materials for the floor and
each wall, we used the RESRAD-BUILD 3.21 computer code (Yu et al. 2001).
The building we used for our modeling was based on the concrete Monument Valley Navajo
hogan. The building modeled had one room with a floor area of 16.4 x 16.4 feet or 269 ft2
(5 x 5 m or 25 m2). Each wall is assumed to be 8.2 feet (2.5 m) high, 16.4 feet (5 m) long, with
an area of 134 ft2 (12.5 m2) (Figure 4.5). Occupancy is assumed to be 70 percent for 365 days a
year (NAS 1999). Since the calculations were scoping in nature, we used the RESRAD-BUILD
default parameters. We assumed that the floors and walls were made of concrete, the radium and
uranium concentrations were equal, and the receptor was at a height of 3.28 feet (1 m).
However, RESRAD-BUILD calculates the contribution of the floor and the wall, so that the
contribution from each part can be separated. The calculations assume no contribution from the
soil beneath the concrete floor. The concrete was assumed to be 6 inches (15 cm) thick, with a
density of 2.4 g/cm3. Results are presented in doses, which are calculated by RESRAD-BUILD.
4.1.1 Results of Building Materials Analysis
From the modeling conducted using RESRAD-BUILD, we calculated doses from external
exposures to U-238 and Ra-226 in full secular equilibrium with their short-lived progenies.1
These doses are listed in Tables 4.1 and 4.2 and are presented graphically in Figures 4.6 and 4.7.
1 This is somewhat different from the way uranium was characterized in the analyses presented in
Chapter 3. In the latter case, all uranium isotopes were assumed to be present in proportion to their natural
abundance, and all long-lived progenies except Ra-226 and its decay chain were included, whereas the analysis in
this chapter addresses only U-238, the dominant isotope, and its short-lived progeny.
4-5
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Figure 4-5. Navajo Hogan Building Model
This 3-D schematic of the Navajo hogan indicates the sources of exposure modeled, with the floor as source 1, and
the walls as sources 2-5. The origin of the geometry is at the lower left-hand corner, where z represents the
vertical extent of the room and x andy represent the lateral extent of the walls.
Source
Table 4-1. Doses from 30 Years of External Exposure to U-238 in a Navajo Hogan
The dose from the floor is about equal to all of the walls combined.
Activity Concentration
(pCi/g) (Bq/kg)
1(37)
50 (1850)
150 (5550)
Dose from Floor
(mrem) (mSv)
1.88 (.02)
93.9 (.9)
282 (2.8)
Dose from One Wall
(mrem) (mSv)
0.554 (.006)
27.7 (.3)
83.1 (.8)
4-6
-------
Figure 4-6. Doses from 30 Years of External Exposure to U-238 in a Navajo Hogan
The floor in the Navajo hogan contributed the most gamma exposure.
300
250
200
£
E 150
100
50
50 100
U-238 Activity Concentration (pCi/g)
150
Doses are listed from external exposure to the floor and to a single wall to allow for an estimate
of the dose if just a concrete slab is contaminated. The calculated dose from a single wall is
between one-fourth and one-third the calculated dose from the floor. The total dose from the
entire structure may be estimated by multiplying the dose from one wall by a factor of four and
adding the result to the dose from the floor.
In order that the uranium in building materials could pose a significant risk from external
exposure, the uranium concentrations in the building materials must be quite high relative to
background concentrations. For example, for a dose of 300 mrem (3 mSv) from the uranium in
the floor over a 30-year period, the U-238 activity would need to be about 180 pCi/g
4-7
-------
(6,660 Bq/kg or about 540 ppm). However, this level could be found in uranium overburden,
and especially in protore.
Table 4-2. Doses from 30 years of External Exposure to Ra-226 in a Navajo Hogan
The dose from the floor is about equal to all of the walls combined.
Activity Concentration
(pCi/g) (Bq/Kg)
1(37)
10 (370)
20 (740)
Dose from Floor
(mrem)(mSv)
139(0.14)
1394(14)
2787 (28)
Dose from One Wall
(mrem)(mSv)
40 (.4)
401 (4)
801 (8)
Figure 4-7. Doses from 30 years of External Exposure to Ra-226 in a Navajo Hogan
3000
2500
2000
u
.§ 1500
8
o
Q
1000
500
6 8 10 12 14 16 18 20
Ra-226 Concentration (pCi/g)
Although U-238 would contribute to the overall radiation exposure, the Ra-226 in the mining
waste materials is the more hazardous of the two radionuclides. A concentration of 1 pCi/g (37
Bq/kg) of Ra-226 in the floor is estimated to result in a dose of about 140 mrem (1.4 mSv)
during 30 years of external exposure. According to the 1985 EPA report to Congress, most of
the uranium mines sampled had Ra-226 concentrations of 20 pCi/g (740 Bq/kg) or more in the
waste. If waste with this radium activity were incorporated into a concrete floor slab, it would
result in a 30-year dose of about 2.8 rem (28 mSv). Figure 4.7 illustrates the relationship
between Ra-226 concentrations and doses from external exposure calculated with RESRAD-
BUILD.
The dose rate from the floor and four walls is approximately 50 |irem/h per pCi/g of Ra-226
(1.4 x 10"4 mSv/hr per Bq/kg). If the exposure rates measured in the Monument Valley Navajo
hogan above were primarily from radium in the floor and walls, and the measurements were
-------
made in the center of the hogan, we estimate that the materials in the hogan contained up to
about 10 pCi/g of Ra-226 (370 Bq/kg).
In addition to direct radiation exposure, radon generation from radioactive decay could also
contribute to risk posed by living in buildings constructed with uranium mine waste, depending
on frequency of air exchange and other factors. As mentioned above in the Sowder et al. (2001)
study of the hogan in Monument Valley, Utah, short-term indoor radon measurements using
multiple methods averaged 50-90 pCi/L (1,850-3,300 Bq/m3) under pseudo-closed conditions.
This greatly exceeds EPA's radon action level of 4 pCi/L (U.S. EPA 2004). However, studies of
other houses constructed with uranium mine waste on Navajo Lands found many had much
lower concentrations of radon, which may have been the result of construction methods and
chimneys which allowed inside air to quickly exit the buildings (L. Setlow, U.S.EPA, personal
communication, 2007)
4.2 Risk of Exposure of On-site Residents to Uranium Mining Waste
As described in Volume 1, the overburden and protore are typically left as piles, and consist of
poorly sorted materials ranging from clay-sized fractions to boulders. Thus, it is not likely that
the material would have a building located on it unless it has been flattened by erosion, was
accessible from a higher elevation, or had been disposed off a hillside to create a terrace. In
populated areas, however, it could be possible for the material to be spread out and a home
subsequently built upon the leveled material. This scenario is included here as an upper bound
on the potential risks from uranium mines, but it is not a focus of this scoping analysis because
there are already guidelines for the amount of radium that is acceptable for Superfund
remediation actions (U.S. EPA 1997a) and in the standards at 40 CFR 192 promulgated by EPA
under the Uranium Mill Tailings Radiation Control Act (UMTRCA).2
The results of a study in Florida (U.S. EPA 1979) developed a relationship between Ra-226 in
soil and indoor working levels (WL). The 1983 EPA report to Congress (U.S. EPA 1983b)
references this document and assumed a similar relationship for a home built on uranium mine
waste material. These data indicate that a concentration of 1 pCi/g (37 Bq/Kg) of Ra-226 in soil
produces an indoor concentration of 1 pCi/L (0.03 Bq/L) of Rn-222, which is equal to 0.004 WL,
assuming an equilibrium factor of 0.4 (UNSCEAR 2000). Thus, a concentration of 5 pCi/g
(185 Bq/kg) of Ra-226 in the soil would produce an indoor radon concentration that is above the
current recommended action level of 4 pCi/L (148 Bq/m3).
The lifetime risk from the indoor radon decay products using current risk estimates is included in
Table 4.3, along with the original estimate from 1983. Since the 1983 report was published,
numerous studies have concluded that indoor radon concentrations are influenced by a
EPA regulations at 40 CFR 192 include limitations for radium and radon at UMTRCA sites: The
disposal areas must be designed to limit releases of radon-222 from uranium byproduct materials to the atmosphere
so as not to exceed an average release rate of 20 pCi/m2/s. This requirement, however, applies only to a portion of a
disposal site that contains a concentration of radium-226 that, as a result of uranium byproduct material, exceeds the
background level by more than 5 pCi/g (185 Bq/Kg) averaged over the first 15 cm below the surface, or more than
15 pCi/g (555 Bq/Kg), averaged over 15 cm thick layers more than 15 cm below the surface.
4-9
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combination of factors, including foundation slab integrity and permeability, indoor and soil
pressure differentials, and the soil radium concentration. Thus, it is difficult to predict the indoor
radon concentration based on soil parameters. However, modeling can provide a general
indication of the radium/radon relationship.
Table 4-3. Estimated Lifetime Risk of Fatal Lung Cancer from Living on
Contaminated Land
This table assumes an average individual is inside the home 75 percent of the time for the 1983 estimate, and
70 percent occupancy for the 2006 estimate. Because the estimate of risk per working level has increased from
that used in 1983, and it is greater than the decrease in occupancy, the estimated cancer risk is higher in 2006.
Rasium-226 in Soil
(pCi/g) (Bq/Kg)
5 (185)
10 (370)
20 (740)
30(1110)
Indoor Working Levels
(WL)
0.02
0.04
0.08
0.12
Lifetime Risk of Fatal Lung Cancer
1983
0.025
0.050
0.100
0.150
2006*
0.029
0.059
0.117
0.176
* The 2006 risk estimate is calculated using the equation presented in Chapter 1 of this volume, under the
Applicability of 1983 Risk Estimates section.
Source: U.S. EPA 1983b.
Additional modeling was conducted using the RESRAD computer code, which embodies a one-
dimensional multi-pathway model for residual radioactivity at sites (Yu et al. 2001). This code
was chosen because of its applicability, widespread use, testing and review, and ease of use.
Most of the RESRAD default values were chosen for this scoping analysis. For the Colorado
Plateau on-site resident scenario, we initially assumed that consumption of groundwater was not
an exposure pathway. We assumed a ventilation rate of 0.5/h, that the foundation was at the
surface with no basement, and that 70 percent of the time was spent indoors and 30 percent
outdoors. With these assumptions, the model predicted indoor radon and external exposure to
direct penetrating radiation to be the major source of radiation exposure, with the indoor radon
exposure higher than the external exposure. Most of the risk from living on contaminated
materials is from the decay of indoor radon. When the ventilation rate is reduced to 0.25/h, the
working levels increase (-0.031 WL for 5 pCi/g (185 Bq/Kg) radium). When we repeated the
analysis with the drinking-water pathway included, using a value of-82 feet (25 meters) for
depth to the aquifer and conservative parameters, such as an evaporation coefficient of 0.5, and
KdS of 10 mL/g for uranium and radium, the indoor radon and external exposure pathways are
still dominant. Pending any consideration of the food chain, which is of most potential
importance for subsistence ranching and hunting, the risks from uranium are dwarfed by the risks
posed by radium and radon.
Uranium mine wastes have the potential to create very high risks to an on-site resident, as
indicated by this analysis and the analysis of the White King/Lucky Lass mine site. Ra-226 is
the primary contributor to risk from the external exposure and indoor radon inhalation pathways.
While the indoor radon concentrations and corresponding working levels resulting from a given
concentration of Ra-226 depend on multiple factors, it is possible to estimate approximate
relationships among these quantities.
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5.0 POTENTIAL ECOLOGICAL IMPACTS FROM URANIUM MINES
This document has focused on the potential risks to humans from exposures to unreclaimed
uranium mining materials. The potential effects on relevant ecosystems have not been
addressed, because they are beyond the scope of this report. Although not analyzed here,
ecosystem effects are briefly mentioned because of the potential importance of the topic in the
consideration of unreclaimed uranium mines. Although the Superfund characterization process
includes radionuclides in the ecological risk assessment and for some individual species, the lack
of an accepted standard methodology for demonstrating protection of ecosystems from radiation
makes the identification of potential effects due to uranium mining difficult. There is, however,
a general framework for ecological risk assessment. As defined in the 1992 Framework for
Ecological Risk Assessment (U.S. EPA 1992), an ecological risk assessment (ERA) is a process
for evaluating the likelihood that adverse ecological effects may occur, or are occurring, as a
result of exposure to one or more stressors.l This framework was applied in the Superfund
guidance, Ecological Risk Assessment Guidance for Superfund: Process for Designing and
Conducting Ecological Risk Assessments, Interim Final (U.S. EPA 1997b).
Ecological risk assessment addresses two major elements, characterization of effects and
characterization of exposure, which provide the focus for three primary phases of activities:
problem formulation, analysis, and risk characterization (U.S. EPA 1998). In these three phases,
the risk assessment process provides a way to develop, organize and present scientific
information so that it is relevant to environmental decisions. Issues to consider are spatial and
temporal, along with assessment endpoints, and whether it is the terrestrial or aquatic
environments that are of concern (U.S. EPA 2000a). When conducted for a particular area such
as a watershed, the ecological risk assessment process can be used to identify vulnerable and
valued resources, prioritize data collection activity, and link human activities with their potential
effects. However, a risk does not exist unless: (1) the stressor has the ability to cause one or
more adverse effects, and (2) it co-occurs with or contacts an ecological component long enough
and at a sufficient intensity to elicit the identified adverse effect (U.S. EPA 1997b). As
discussed in this chapter, it is very possible that the stressors to the surrounding ecosystem may
not be the radioactive materials, but rather the other hazardous constituents that may be
associated with uranium mine sites.
Efforts are underway to extend the ecological risk assessment approach to radiation. In recent
work, Jones et al. (2003) state that, "potentially susceptible receptors [to radiation] include
vertebrates and terrestrial plants." EPA has no radiation dose standards for the protection of
flora and fauna, but the Department of Energy (DOE) (Jones et al. 2003) has suggested levels of
exposure for the protection for the following: natural populations of aquatic biota (1 rad d-1 or
10 mGy d-1), terrestrial plants (1 rad d-1 or 10 mGy d-1) and animals (0.1 rad d-1 or
1 mGy d-1).2 The question remains whether these levels are indeed protective.
1 A "stressor" is any chemical, physical, or biological entity that can induce adverse effects on individuals,
populations, communities, or ecosystems.
2 1 gray = 100 rad; thus 1 mGy = 0.001 Gy = 0.1 rad or 100 mrad.
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DOE has recently issued a technical standard on applying these levels in the document^ Graded
Approach for Evaluating Radiation Doses to Aquatic and Terrestrial Biota (U.S. DOE 2002).
The graded screening approach uses three tiers, becoming progressively more rigorous and
detailed: a scoping assessment, a screening ERA, and a more detailed ERA that uses site-
specific information (Jones et al. 2003). As the tiers become more site-specific, the assumptions
become less conservative. In the screening phase, this process uses biota concentration guides
(BCGs) for water and sediment for evaluating aquatic systems, and water and soil for evaluating
a terrestrial system. These BCGs are set "so that doses received by real biota exposed to such
concentrations are not expected ever to exceed the biota dose limits" (Higley et al. 2003). The
BCGs for aquatic and terrestrial systems are reproduced in Tables 5.1 and 5.2. The radiation
levels found at some of the uranium mines where sub-ore and ore-grade materials have been left
on site could exceed the levels identified by DOE, especially for radium.
Table 5-1. Biota Concentration Guides (BCGs) for Water and Sediment for Evaluation
of an Aquatic System
Nuclide
226Ra
228Ra
232Th
233U
234U
235U
238U
BCG for Water
Bq/m3
2x 102
2 x 102
1 x 104
7 x 103
7x 103
8x 103
8 x 103
pCVL
5.4 x 10°
5.4 x 10°
2.7 x 102
1.9 x 102
1.9 x 102
2.2 x 102
2.2 x 102
Organism
Responsible for
Limiting Dose in
Water
Riparian3 Animal
Riparian Animal
Aquatic Animal
Aquatic Animal
Aquatic Animal
Aquatic Animal
Aquatic Animal
BCG for Sediment
Bq/kg
4x 103
3 x 103
5 x 104
2 x 105
2x 105
1 x 105
9 x 104
pCi/g
1.1 x 102
8.1 x 101
1.4 x 103
5.4 x 103
5.4 x 103
2.7 x 103
2.4 x 103
Organism
Responsible for
Limiting Dose in
Water
Riparian Animal
Riparian Animal
Riparian Animal
Riparian Animal
Riparian Animal
Riparian Animal
Riparian Animal
a A "Riparian Animal" is an animal that lives on a riverbank and hence spends time on land and in water, e.g., a
muskrat.
Source: Reproduced from Higley et al. 2003.
Table 5-2. Biota Concentration Guides for Water and Soil for Evaluation of a
Terrestrial System
Nuclide
226Ra
228Ra
232Th
233U
234U
235U
238U
BCG for Water
Bq/m3
3 x 105
3 x 105
2x 106
1 x 107
1 x 107
2x 107
2 x 107
pCi/L
8.1 xlO3
8.1 x 103
5.4 x 104
2.7 x 105
2.7 x 105
5.4 x 105
5.4 x 105
Organism Responsible
for Limiting Dose in
Water
Terrestrial Animal
Terrestrial Animal
Terrestrial Animal
Terrestrial Animal
Terrestrial Animal
Terrestrial Animal
Terrestrial Animal
BCG for Sediment
Bq/kg
2xl03
2 x 103
6x 104
2x 105
2 x 105
1 x 105
6 x 104
Bq/m3
5.4 XlO1
5.4 x 101
1.6 x 103
5.4 x 103
5.4 x 103
2.7 x 103
1.6 x 103
Organism
Responsible for
Limiting Dose in
Water
Terrestrial Animal
Terrestrial Animal
Terrestrial Animal
Terrestrial Animal
Terrestrial Animal
Terrestrial Animal
Terrestrial Animal
Note: 1 pCi/L = 37 Bq/m3, 1 pCi/g = 37 Bq/kg
Source: Reproduced from Higley et al. 2003.
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5.1 Other Metals
There could be multiple stressors from uranium mining, especially in watersheds where a high
density of uranium mines could have a larger effect than a single mine. The metals associated
with uranium may cause adverse ecological effects, depending on the concentration and
bioavailability. Arsenic, a human carcinogen, is one and it was discussed in Chapter 3. Other
common associations include copper, phosphate, molybdenum, and vanadium. Lead and
selenium are additional metals noted in some Arizona mines in the EPA Abandoned Mine Lands
portion of the CERCLIS3 database. See Table 5.3 for mineral ores with which uranium (and
radium) may be associated. Vanadium and uranium are commonly mined together on the
Colorado Plateau (U.S. EPA 2006a).
Most of the mines located in the sedimentary sandstone deposits of the southwestern United
States are not in pyritic formations, and the resulting runoff waters or pit lakes are generally
neutral to alkaline in character (pH of 7 or higher). Low precipitation rates and the resultant lack
of water may further reduce the potential for generation of acid mine or rock drainage (AMD or
ARD) from waste rock, for example, in both the Colorado Plateau and the Shirley Basin of
Wyoming (U.S. EPA 2006a). For mines elsewhere, AMD/ARD can be a problem. Midnite
Mine in Washington State is an example of a large uranium mine in which AMD did occur.
While AMD/ARD can enhance contaminant mobility by promoting leaching from exposed
wastes and mine structures, releases can also occur under neutral pH conditions (U.S. EPA
2000a).
The effects of the metals can be assessed within the Superfund methodology. An example of this
was mentioned as part of the discussion of the White King/Lucky Lass Superfund site. In that
ecological risk assessment, no adverse ecological effects were seen from the radioactive
materials, but the associated metals did pose a potential ecological threat to a local shrew
species. Other mining sites have created environmental problems, and some are on the National
Priorities List for cleanup. Midnite Mine, for example, underwent a preliminary ecological risk
assessment (URS 2003), and a number of metals were examined, including copper, lead, arsenic,
selenium, uranium, vanadium, zinc, molybdenum, and chromium. Uranium-235, uranium-238
and thorium-232 were also evaluated. According to the final ecological risk assessment, there
were situations where both the radioactivity and the metals exceeded guidelines (Lockheed
Martin 2005). The record of decision concludes that, "Contaminants in surface water, ground
water, surface materials, and air represent a threat to human and ecological receptors" (U.S. EPA
2006c).
Although not analyzed here, there may be environmental effects, in addition to potential human
health effects, from unreclaimed uranium mines. While many of the mines are remote and may
not be visited by humans, the flora and fauna would be exposed for much longer periods of time,
and thus could be affected by unreclaimed mines. Issues to be considered for an ecological risk
3 The Comprehensive Environmental Response, Compensation and Liability Information System
(CERCLIS) Database contains general information on sites across the nation and U.S. territories including location,
contaminants, and cleanup actions taken. The database can be downloaded from the web at
http://www.epa.gov/superfund/programs/aml/amlsite/nonnpl.htm.
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assessment of unreclaimed mines could include the identification of stressors for the different
types of uranium mines, affected species at different sites, the potential exposures, and the
endpoints for determining effects.
Although radiological and chemical toxicity should be treated as concerns, the closure of mine
shafts that have long remained unreclaimed must also be considered carefully. In parts of the
country where open mine shafts have long been part of the landscape, animal species—most
notably bats—may rely on those mines shafts as critical habitat. Endangered bat species have
been documented nesting in unreclaimed mines. If a survey by a biologist determines the
presence of bat species in an abandoned mine, adequate closure may be accomplished by means
of a "bat gate," a metal grate that prevents humans from entering but allows the free passage of
bats (Burghardt 2003).
Table 5-3. Mineral Commodities with Uranium Associations
Several mineral ores often, though not always, have TENORM-associated wastes resulting
from co-occurrence of uranium and radium.
Aluminum (bauxite)
Coal (and coal ash)
Copper
Fluorospar (fluorite)
Gypsum
Molybdenum
Niobium
Phosphate (phosphorus)
Potassium (potash)
Precious metals (gold, silver)
Rare earths: yttrium, lanthanum, monazite, bastanite, etc.
Tin
Titanium (leucoxene, ilmenite, rutile)
Tungsten
Vanadium
Zircon
Source: U.S. EPA 2003d.
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6.0 UNCERTAINTIES
The major uncertainty in this analysis is the actual exposure that people will experience.
Because many abandoned uranium mines are on federal lands, the most likely exposure scenario
is recreational use, but the true nature and extent of the recreational use is unknown. Exceptions
to this are Native Americans who live around the uranium mines and personnel who may work
around the sites. In addition, the number of people exposed would depend on the number of
mines that have been reclaimed. Some mines have been reclaimed, but the total number is
unknown. In the 1983 EPA study (U.S. EPA 1983b), the authors noted that many of the mines
from the 1950s and 1960s had not been reclaimed at that time. Survey work done by Otten for
EPA (1998) found that in many uranium-producing states, perhaps half of the mines or more had
been reclaimed. No other survey has been conducted since that time. In the 1970s, surveys
identified hundreds of potential buildings constructed from what was believed to have been
uranium mining-related material. However, little is known about the extent of building
contamination or the level of contamination in the building materials, or whether they remain or
are occupied.
Another uncertainty is the true effect uranium mines have on the ground water and the
subsequent use of the water. In many parts of the Southwest, where many of the mines are
located, the primary sources of drinking water are deep-lying aquifers, so shallow open-pit or
underground mines may not contaminate the water because of the limited infiltration.
Furthermore, since uranium mines are in mineralized areas, it can be difficult to differentiate
between a groundwater problem caused by a uranium mine and naturally occurring uranium. In
other instances, in areas with surface water flow, such as the Ross-Adams Mine in Alaska, or
Orphan Mine in Arizona (see U.S. EPA 2006a), a local source of drinking water may be
contaminated by water flowing through uranium mine waste or the mine itself, and serve as a
possible ingestion pathway for radiation exposure.
The other major uncertainly involves the concentrations of contaminants. The primary
radiological contaminant of concern is Ra-226, which would contribute the greatest risk—from
external exposure—to the occasional recreationalist. Uranium may also be a contaminant of
concern, especially if it can migrate to a drinking-water source where its chemical toxicity
becomes the health hazard. There is information that can be used to bound the potential
exposures to both of these radionuclides, but the concentrations vary within a site and between
the true overburden and amount of protore at a specific mining location. Arsenic, a carcinogen,
has been shown to be associated with uranium mine wastes and can reach high levels at mine
sites, but arsenic concentrations can be highly variable. At some sites, the risk from arsenic may
dominate the radiological risk, and other metals may also contribute some uncertain level of
hazard. Since this analysis was done on a generic, scoping basis, site-specific analyses would
remove much of the uncertainties encountered here.
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7.0 CONCLUSIONS
7.1 Summary
The majority of uranium production in the U.S. has come from several hundred underground and
open-pit mines out of the thousands of mines and exploration workings known to exist. Some of
these have been closed and remediated, at least two have been placed on the National Priorities
List (Superfund) for cleanup, and others have been in standby mode where the owners have been
waiting for the price of uranium to increase, as it has in 2006. The focus of this scoping report,
however, has been on an investigation of potential risks from the thousands of relatively small
mines and exploration sites that were abandoned over the years. With this report we have tried
to identify the most likely exposure scenario for the abandoned mines, develop a first order
estimate of cancer risks using some conservative assumptions, and identify if there are potential
ecological effects that may develop around these mines.
Of the thousands of uranium mines in the continental United States, most are concentrated in
Colorado, Utah, Wyoming, and New Mexico, and to a lesser extent, Arizona and Texas. For the
small number of uranium mines in other regions, uranium is typically a byproduct of other
mineral production. Many of the Four Corners States' mines are concentrated in a small number
of watersheds. Though some Superfund removal actions have taken place within the Colorado
Plateau, the two uranium mines on the National Priorities List are outside of the major uranium-
producing states.
Most abandoned uranium mines are likely to have elevated radium and uranium concentrations,
and possibly elevated levels of other contaminants such as arsenic. An analysis of the location of
uranium mine records indicates that many are on federal lands, so a primary exposure scenario
pertains to short-term recreational activities, including short-term occupation. Another scenario
of concern is the use of mine waste material as building materials for those situations where the
mines are not remote and material can be transported by nearby populations. In the recreation
scenario, short-term exposure to radium, uranium, and arsenic appears to create only minimal
additional cancer risk. This additional risk is dominated by external gamma exposure associated
with radium in the waste material. The radioactivity in sub-ore grade uranium mine waste can be
very high, so longer-term exposures from repeated visits to a high radium/high gamma site could
begin to create a higher risk, even to a recreational user. The highest end of the risk spectrum is
the scenario in which abandoned mine areas are used as home sites, which could pose a
significant cancer risk to any long-term inhabitant. Long-term inhabitants who live near the
mine sites might also use uranium mine waste material in building materials, and they would
face additional risk from those radioactive building materials. It appears that those living on
western Tribal lands appear to be most at risk as potential residents on or near abandoned
uranium mine sites, or from the frequent visiting or passing through contaminated sites and
wastes.
In general, the risks from these sites are primarily from occasional exposures and are likely to be
minimal, even with conservative assumptions. The risk resulting from frequent use of a site,
however, approaches a resident's exposure. Due to the predominant recreation scenario, the risk
analysis examined risks in terms of days of exposure instead of the typical annual exposure,
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although 350-day scenarios have been included to represent the exposure level for one year.
From the estimates of the risk provided in this document, it is possible to quickly determine a
first order estimate of the risks from a site, given the predominant contaminants, with the caveat
that specific site conditions and site use would need to be factored in for a more realistic risk
estimate.
Many of the abandoned uranium mines occur in areas with low precipitation and deep
groundwater so that risk to ground-water drinking water sources is often low for at least the
short-term (tens of years). However, some abandoned uranium mines occur in areas with higher
precipitation. Abandoned uranium mines that are the most likely to affect groundwater are
those that intersect groundwater (e.g., underground mines or deep surface mines) or are above
shallow aquifers. Both radium and uranium have had MCLs established for them in drinking
water supplies, but uranium is the most likely candidate to contaminate groundwater, since
radium-226 is typically more immobile. In the case of uranium, the MCL is based on the
limiting effect of chemical toxicity, not the radiological properties.
Ecological effects were not a focus of this report, but they were considered. Radionuclide and
other heavy metal concentrations could be high enough to affect flora and fauna around
abandoned mines, especially in watersheds with a high mine density. Indeed, it may be the flora
and fauna that are affected much more than human health, and it may be the non-radioactive
metals that produce the more significant ecological effects. This may be especially true where
uranium is a secondary commodity, such as in the Lefthand Creek watershed in Colorado. At
the same time, however, species may have grown accustomed to the presence of mine shafts that
remain unreclaimed, and may, in fact, rely on them for habitat.
7.2 Potential Considerations for Site Prioritization
Ideally, all abandoned uranium mine sites would be remediated; however, given budget
restraints, it is recognized that the most likely sites to be remediated are those that pose the
greatest threat to human health and the environment. There are a number of items that could be
considered when trying to prioritize the mines to be remediated. For example, in the cases where
the radionuclides are likely to reach the groundwater, surface water, or springs, uranium may be
the limiting radionuclide, because it is typically more mobile than radium. Radium may most
often be the limiting factor in other cases because of the risk from external exposure. Less
information is known about thorium values and the importance of thorium relative to radium. In
some cases, the non-radiological metals may be the most hazardous of the mine waste
constituents.
7.2.1 Depth to Groundwater and Annual Precipitation
EPA considers groundwater a resource for which it is easier to prevent pollution than to treat
pollution after the fact. Those uranium mines that are located in areas with shallow (<50-60 feet
or <~20 meters) groundwater resources have the potential to contaminate underlying aquifers
within decades. Coupled with moderate amounts of precipitation (>~20 inches or >~50 cm),
radioactive and metal contaminants at uranium mines could create a groundwater problem if not
addressed. Large mines and underground mines that intersect aquifers have caused groundwater
contamination. A scoping study such as this can identify some potential issues in this area, but it
7-2
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cannot account for all the variations in site characteristics, so the geology and hydrology of a
particular site would have to be examined (e.g., pH) when making remediation decisions.
7.2.2 Frequency of Use
The main tenets of protection from radioactive materials are time, distance and shielding. At one
end of the spectrum, if one were to live on the mine waste materials or be exposed to mine
wastes as part of construction materials, the risk of cancer from doing so could be relatively high.
The scoping analysis in this report indicates that people who spend only small amounts of time at
these abandoned mines probably have low risk. This low-risk consequence changes if one of
these sites gets frequent use, creating a continuum of risk which we have tried to illustrate
through the use of exposure calculations based on days per year of exposure.
In addition to current uses (e.g., recreation), the potential for future population growth and use
could also be considered. The West and Southwest have experienced population growth in
recent years, and second homes have also recently become popular in areas that were formerly
primarily recreational. Anecdotal information suggests that home developments may be
encroaching on areas of abandoned mines or mine wastes. In these cases, nearby populations
may increase the potential use of these properties, with a concurrent increase in potential
exposures.
The frequency of use may be related to their distance from roads. In other words, how remote
are the mines? With the mines located on federal property, access may depend on fire roads or
roads previously used during the mine's operation which are likely in disrepair so that access
would be by foot, all terrain vehicles or possibly four-wheel drive vehicles. Some mines,
however, may be located along well developed roads with easy access which may lead to more
frequent visits or visits of longer duration.
7.2.3 Presence and Concentrations of Contaminants in Soils, Water, and Sediments
A major driver for the overall risk is the presence of contaminants. In the case of abandoned
uranium mines, the contaminants would be both radioactive and stable metals. Radium,
uranium, and possibly thorium could pose risks from external gamma exposures, but arsenic and
other heavy metals (e.g., vanadium, selenium, copper, molybdenum) could pose a risk as well,
especially to flora and fauna if there are enough waste materials. Some of the waste material
quantities may be so minimal in area or volume that they do not pose a problem.
7.2.4 Density of Mines
One observation from this analysis is that the uranium mines are often along drainages where
there can be a high density of mines or mine portals and associated wastes (see Figure 2.4 for
example). While one mine may not pose a problem, a number of mines close together may
increase the potential for adverse health or ecological effects, which may be seen at some
distance from an individual mine site.
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7.2.5 Level of A cceptable Risk
Lastly, the level of acceptable risk will also be important to determining how to prioritize the
mines. The level of cancer risk typically used by EPA in the Superfund program is the risk range
of 1 in 10,000 (10"4) to 1 in 1,000,000 (10"6) and the level of acceptable risk for non-carcinogens
(i.e., some metals) is a hazard ranking less than 1. Sites which get frequent visitation may
approach the upper end of the cancer risk range, while other sites would be at the lower end of
the risk range. Residential exposure to uranium mine wastes, if it were to occur, would most
likely be at the high end of risk range or even above.
The scoping analysis presented in this report indicates that at least some of the abandoned
uranium mines have the potential to pose health and ecological hazards from both radioactive
and non-radioactive materials. Data indicate that the concentrations of contaminants can be high
enough to create adverse health effects if people were to spend substantial time on the sites.
Non-radiological contaminants may be the most significant hazard, especially for flora and
fauna. Since many of the sites are on federal lands, the largest exposure would be from
recreational visits, or occupational use by a government employee or contractor, where the
relatively short period of exposures would minimize the impact of high concentrations of
contaminants. For the occasional visitor to abandoned mines, the mine wastes typically do not
produce a significant radiation risk. However, individuals who visit a site frequently or for long
periods of time can incur substantial risks. Residential exposure through on-site exposure or
through the use of contaminated building material is not likely in most cases, except for some
Tribal members, such as in the Navajo Nation, or other nearby residents. Where it does occur,
the risks from these situations could be quite high.
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8.0 BIBLIOGRAPHY
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Mines—Site Characterization (September 2002)." Preliminary results presented at U.S.
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S.P. March, D.F. Woodward, B.A. Kimball, S.E. Finger, L.S. Ischinger, J.C. Fordham, M.S.
Power, C.M. Bunck, and J.W. Jones. 1997. A Science-based, Watershed Strategy to Support
Effective Remediation of Abandoned Mine Lands: Fourth International Conference on Acid
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Colon, C.F. Jove, P.V. Brady, M.D. Siegel, andE.R. Lindgren. 2001. Historical Case
Analysis of Uranium Plume Attenuation. NUREG/CR-6705 SAND2000-2554. Washington,
DC: U.S. Nuclear Regulatory Commission, Office of Nuclear Regulatory Research, February
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Cowherd, C., R. Bohn, and T. Cuscino. 1979. "Iron And Steel Plant Open Source Fugitive
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CRCPD (Conference of Radiation Control Program Directors). 2004. Part N, Regulation and
Licensing of Technologically Enhanced Naturally Occurring Radioactive Material (TENORM).
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DeLaune, R., J. Pardue, W. Patrick, Jr., and C. Lindau. 1996. Mobility and Transport of Radium
in Sediment and Waste Pits. Final project report submitted to the South and Southwest
Hazardous Substance Research Center, and conducted under EPA Grant Number R8255 13.
Baton Rouge, Louisiana: Louisiana Water Resources Research Institute, March 1996.
ESRI (Environmental Systems Research Institute, Inc.) 2001. ESRIData & Maps 2000: An
ESRI White Paper. Redlands, California: ESRI, August 2001. http://www.esri.com
Ferderer, D.A. 1996. National Overview of Abandoned Mine Land Sites Utilizing the Minerals
Availability System and Geographic Information System (GIS) Technology. Open-File Report
96-549. Reston, Virginia: U.S. Department of the Interior, U.S. Geological Survey, 1996.
Finch, W. 1998. Unpublished compilation of uranium production data from U.S. Geological
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8-2
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A Continuing Legacy for the Navajo Nation." Washington, DC: U.S. Environmental Protection
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-------
U.S. EPA (U.S. Environmental Protection Agency). 2006c. Midnite Mine Superfund Site
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Midnite Mine." November 2003.
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Contamination Associated with Mineral and Resource Extraction Industries. Washington DC:
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Response, OSWER 93565.4-24 Washington, DC: U.S. EPA, December 2002.
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TENORM—
Evaluating Occurrence and Risks, an SAB Advisory; A Science Advisory Board Advisory on
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Integrating point and non-points Sources - 2001.
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are: EPA-823-C-01-006, BASINS Version 3.0 CDROM, EPA Region 6; EPA-823-C-01-008,
BASINS Version 3.0 CDROM, EPA Region 8; EPA-823-C-01-009, BASINS Version 3.0
CDROM, EPA Region 9; and EPA-823-C-01-010, BASINS Version 3.0 CDROM, EPA Region
8-4
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10. (The mineral location data is no longer available on-line from the BASINS site. The EPA
metadata for this database is still available on-line at
http://www.epa.gov/waterscience/basins/metadata/mines.htm. The GIS layer in BASINS is
found at http://www.epa.gov/waterscience/basins/metadata/hydunits.htm.)
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Radionuclides: User's Guide. EPA/540-R-00-007. Washington, DC: U.S. EPA, October
2000.
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Water Regulations: Radionuclides. Final Rule." 40 CFR Parts 9, 141, and 142. Federal
Register. Vol. 65, No. 236, December 7, 2000.
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Water and Drinking Water, Office of Indoor Air and Radiation, and U.S. Geological Survey,
March 2000.
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Partition Coefficient, Kd, Values. EPA/402-R-99-004B. Washington, DC: U.S. EPA, August
1999.
U.S. EPA (U.S. Environmental Protection Agency). 1998. Guidelines for Ecological Risk
Assessment. EPA/630-R095-002F. 01 April 1998. Washington, DC: U.S. EPA, April 1998.
U.S. EPA (U.S. Environmental Protection Agency). 1997a. Establishment of Cleanup Levels
for CERCLA Sites with Radioactive Contamination. OSWERNo. 9200.4-18. Washington,
DC: U.S. EPA, August 1997.
U.S. EPA (U.S. Environmental Protection Agency). 1997b. Ecological Risk Assessment
Guidance for Superfund: Process for Designing and Conducting Ecological Risk
Assessments. Interim Final Document. EPA/540-R-97-006, OSWER 9285.7-25, PB97-96321
1. Washington, DC: U.S. EPA, Office of Solid Waste and Emergency Response, June 1997.
U.S. EPA (U.S. Environmental Protection Agency). 1996a. Soil Screening Guidance:
User's Guide. EPA/540-R-96-018. Washington, DC: U.S. EPA, Office of Solid Waste and
Emergency Response, July 1996.
8-5
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U.S. EPA (U.S. Environmental Protection Agency). 1996b. Soil Screening Guidance:
Technical Background Document. EPA/540-R-95-128. Washington, DC: U.S. EPA, Office
of Solid Waste and Emergency Response, May 1996.
U.S. EPA (U.S. Environmental Protection Agency). 1995. Extraction and Beneficiation of
Ores and Minerals: Uranium. Vol. 5. Technical Resource Document. EPA/530-R-94-032.
Washington, DC: U.S. EPA, Office of Solid Waste, January 1995.
U.S. EPA (U.S. Environmental Protection Agency, Science Advisory Board). 1994. An SAB
Report: Review of Diffuse NORM Draft Scoping Document. EPA-S AB-RAC-94-013.
Washington, DC: U.S. EPA, May 1994.
U.S. EPA (U.S. Environmental Protection Agency). 1992. Framework for Ecological Risk
Assessment. EPA/630/R-92/001. Washington, DC: U.S. EPA, February 1992.
U.S. EPA (U.S. Environmental Protection Agency). 1989a. Risk Assessments,
Environmental Impact Statement, NESHAPSfor Radionuclides. Vol. 2. Background
Information Document. EPA/520/1-89-006-1. Washington, DC: U.S. EPA, September
1989.
U.S. EPA (U.S. Environmental Protection Agency). 1989b. "Human Health Evaluation
Manual." In Risk Assessment Guidance for Super/and. Vol.1. EPA/540/1-89/002.
Washington, DC: U.S. EPA, December 1989.
U.S. EPA (U.S. Environmental Protection Agency). 1989c. "National Emission Standards
for Hazardous Air Pollutants; Radionuclides; Final Rule and Notice of Reconsideration." 40
CFRPart 61 Federal Register, Vol. 54, No. 240, December 15, 1989.
U.S. EPA (U.S. Environmental Protection Agency). 1985. Report to Congress on Wastes
from the Extraction and Beneficiation of Metallic Ores, Phosphate Rock, Asbestos,
Overburden from Uranium Mining, and Oil Shale. EPA 530/S W-85-003. Washington, DC:
U.S. EPA, December 1985.
U.S. EPA (U.S. Environmental Protection Agency). 1983a. Report to Congress on the
Potential Health and Environmental Hazards of Uranium Mine Wastes. Vol.1. EPA 520/1-
83-007. Washington, DC: U.S. EPA, June 1983.
U.S. EPA (U.S. Environmental Protection Agency). 1983b. Report to Congress on the
Potential Health and Environmental Hazards of Uranium Mine Wastes. Vol.2. EPA 520/1-
83-007. Washington, DC: U.S. EPA, June 1983.
U.S. EPA (U.S. Environmental Protection Agency). 1983c. Report to Congress on the
Potential Health and Environmental Hazards of Uranium Mine Wastes. Vol. 3. EPA 520/1-
83-007. Washington, DC: U.S. EPA, June 1983.
U.S. EPA (U.S. Environmental Protection Agency). 1979. Indoor Radiation Exposure Due
to Radium-226 in Florida Phosphate Lands. EPA-520/4-78-013. Washington, DC: U.S.
EPA, September 1979.
8-6
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U.S. EPA (U.S. Environmental Protection Agency). 1973. Summary Report of Radiation
Surveys Performed in Selected Communities. Washington, DC: U.S. EPA, Office of
Radiation Programs, 1973.
U.S. EPA (U.S. Environmental Protection Agency), n/d. "Radionuclide Table: Radionuclide
Carcinogenicity - Slope Factors." http://www.epa.gov/radiation/heast/docs/heast2 table 4-
d2_0401.pdf
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Engineers). 2000. Abandoned Uranium Mines Project Arizona, New Mexico, Utah—Navajo
Lands Project Atlas, 1994-2000. San Francisco and Los Angeles, California: U.S. EPA and
USACE, December 2000.
USFS (U.S. Forest Service). 1991. Draft EIS RI/FSfor the Cleanup & Rehabilitation of the
White King and Lucky Lass Uranium Mines, Fremont National Forest, Lakeview, Oregon.
R6-FRE-00 1-91. Klamath Falls, Oregon: U.S. Department of the Interior, USFS, August
1991.
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Oklahoma City, Oklahoma: Kerr-McGee Corporation, August 1997.
Yu, C., AJ. Zielen, J.-J. Cheng, DJ. LePoire, E. Gnanapragasam, S. Kamboj, J. Arnish, A.
Wallo III, W. A. Williams, and H. Peterson. 2001. User's Manual for RESRAD. Version 6.
Argonne, Illinois: Argonne National Laboratory, Environmental Assessment Division, July
2001.
Yu, C., DJ. LePoire, J.-J. Cheng, E. Gnanapragasam, S. Kamboj, J. Arnish, B.M. Biwer. A.J.
Zielen, W.A.Williams, A. Wallo III, and H. Peterson. 2003. User's Manual for RESRAD-
BUILD Version 3, ANL/EAD/03-1, Environmental Assessment Division, Argonne National
Laboratory, Argonne, Illinois, June 2003. http://web.ead.anl.gov/resrad/documents/ANL-
EAD-03-l.pdf.
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Appendix I. Swimming Risk
This appendix provides supplemental information on the swimming risks discussed in Chapter 3.
Swimming risks were assumed to come from two sources; (1) ingestion and (2) immersion. In
the drinking water discussion of Chapter 3 of the main report, we identified a recreational
exposure scenario of 10 years of exposure, 14 days a year, and drinking 2 liters of water a day.
For this recreational scenario, the lifetime cancer risk from drinking water contaminated with a
range of uranium concentrations was in the 10"5 to 10"6 risk range. However, the analysis in
Chapter 3 did identify that long-term use of pit-lake water could pose cancer risks. The potential
hazard from the pit lakes may be greater from metals, such as arsenic, than from radionuclides.
Since ingestion risks from ranges of radionuclide concentrations were discussed in Chapter 3,
they are not discussed further here.
To calculate the immersion risks from exposure to radionuclides, we first calculated a dose using
the formula modified from Whelan et al. (2006), and then applied a dose to risk coefficient from
Tables 7.3 and 7.6 of Federal Guidance Report No. 13 (U.S. EPA 1999) to develop age-averaged
site-specific cancer mortality and morbidity risk estimates. The formula for the immersion dose
is as follows:
swimming external dose = Cw * EDFS * T exposure
where Cw is the radionuclide concentration in the water in pCi/L, the EDFS is the External Dose
Factor for Swimming in rem/hr per pCi/L, and the time of exposure is length of time a swimmer
would be in the water in hours. To calculate the EDFS for the uranium and thorium decay series,
we used the DCAL program (Eckerman et al. 2006), a comprehensive software system for the
calculation of tissue dose and subsequent health risk from intakes of radionuclides or exposure to
radionuclides present in environmental media. The results are listed below in Tables All and
AI.2 for the U-238 and Th-232 decay series. Note that Ra-226 is included in the U-238 dose and
risk calculations. The totals would apply if secular equilibrium were assumed; this is an unlikely
case, because of the tendency for the radionuclides to settle into the sediment, as well as being
dissolved in the water column. Table AI.3 shows the dose equivalent and risks per pCi/L for
both decay series combined as a function of time spent immersed in the water.
In summary, the cancer risks from immersion due to swimming are very small per pCi/L from
the U-238 and Th-232 decay series. Even if secular equilibrium were assumed and all the
radionuclides in either series were present, the cancer risks from immersion while swimming are
negligible for the recreational scenario, even at concentrations of 10s to 100s of pCi/L. The
cancer risks from ingesting water while swimming are also anticipated to be low, based on the
drinking water discussion in Chapter 3 of the main report.
AI-1
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Table ALL Uranium-238 Dose Equivalent Rate and Risk per pCi/L per hour (EDFS):
Water Immersion
Nuclide
U-238
Th-234
Pa-234m
Pa-234
U-234
Th-230
Ra-226
Rn-222
Po-218
At-218
Pb-214
Bi-214
Po-214
Pb-210
Bi-210
Po-210
Total
Dose Equiv. Rate
[(rem/hr)per (pCi/L)]
7.8E-14
8.8E-12
2.7E-11
2.5E-09
1.9E-13
4.5E-13
8.3E-12
5.1E-13
1.2E-14
3.0E-12
3.2E-10
2.1E-09
1.1E-13
1.4E-12
4.0E-12
1.1E-14
5.0E-09
Mortality Risk1
4.5E-17
5.0E-15
1.5E-14
1.4E-12
1.1E-16
2.6E-16
4.8E-15
3.0E-16
7.0E-18
1.7E-15
1.8E-13
1.2E-12
6.3E-17
8.0E-16
2.3E-15
6.5E-18
2.9E-12
Morbidity
Risk
6.6E-17
7.4E-15
2.2E-14
2.1E-12
1.6E-16
3.8E-16
7.1E-15
4.3E-16
l.OE-17
2.5E-15
2.7E-13
1.8E-12
9.3E-17
1.2E-15
3.4E-15
9.5E-18
4.2E-12
Table AI.2. Thorium-232 Dose Equivalent Rate per pCi/L per hour(EDFS): Water
Immersion
Nuclide
Th-232
Ra-228
Ac-228
Th-228
Ra-224
Rn-220
Po-216
Pb-212
Bi-212
Po-212
Tl-208
Total
Dose Equiv. Rate
[(rem/hr) per (pCi/L)]
2.20E-13
O.OOE+00
1.29E-09
2.41E-12
1.25E-11
5.00E-13
2.24E-14
1.82E-10
2.54E-10
O.OOE+00
4.86E-09
6.61E-09
Mortality
Risk
1.26E-16
O.OOE+00
7.44E-13
1.39E-15
7.21E-15
2.87E-16
1.29E-17
1.05E-13
1.46E-13
O.OOE+00
2.80E-12
3.80E-12
Morbidity
Risk
1.86E-16
O.OOE+00
1.10E-12
2.04E-15
1.06E-14
4.23E-16
1.89E-17
1.54E-13
2.15E-13
O.OOE+00
4.11E-12
5.59E-12
Mortality risk is 5.575 E-4 per rem; Morbidity risk is 8.46 E-4 per rem.
AI-2
-------
Table AI.3. Total Dose Equivalent and Risk per pCi/L per hour (EDFS): Water
Immersion
Time Spent
Swimming
(hours)
1
2
3
4
5
10
15
20
25
30
40
50
75
100
Dose Equivalent
(rem per pCi/L)
1.2E-08
2.3E-08
3.5E-08
4.6E-08
5.8E-08
1.2E-07
1.7E-07
2.3E-07
2.9E-07
3.5E-07
4.6E-07
5.8E-07
8.7E-07
1.2E-06
Mortality Risk
(per pCi/L)
6.5E-12
1.3E-11
1.9E-11
2.6E-11
3.2E-11
6.5E-11
9.7E-11
1.3E-10
1.6E-10
1.9E-10
2.6E-10
3.2E-10
4.9E-10
6.5E-10
Morbidity
Risk
(per pCi/L)
9.8E-12
2.0E-11
2.9E-11
3.9E-11
4.9E-11
9.8E-11
1.5E-10
2.0E-10
2.5E-10
2.9E-10
3.9E-10
4.9E-10
7.4E-10
9.8E-10
AI-3
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Appendix I References
Eckerman, K., F., R.W. Leggett, M. Cristy, C.B. Nelson, J.C. Ryman, A.L. Sjoreen, R.C. Ward,
2006. User's Guide to the DCAL System ORNL/TM-2001/190, Oak Ridge National
Laboratories. August 2006.
U.S. EPA (U.S. Environmental Protection Agency). 1999. Cancer Risk Coefficients for
Environmental Exposure to Radionuclides, Federal Guidance Report No. 13, U.S. EPA Office of
Air and Radiation, Washington, DC, EPA 402-R-99-001, September 1999.
Whelan, G., C. Sivaraman, W.D. Millard, MJ. Simpson, G.M. Gelston, J.K. Young, M.A. Pelton,
T.P. Khangaonkar, Z. Yang, T.R. Downing, D.L. Strenge, B.L. Hoopes, C. Lee, and L.E.
Hachmeister, 2006. Rapid Risk Assessment FY05 Annual Summary Report. Laboratory Directed
Research and Development, PNNL-15697, March 2006. Prepared for the U.S. Department of
Energy under Contract DE-AC05-76RL01830.
AI-4
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Appendix II. Calculation of Slope Factors for Naturally Occurring
Radionuclides
In developing the target soil screening levels found in Chapter 3 of this report, the slope factors
for external exposure to, ingestion of, and inhalation of soil at an infinite depth must be
considered. This appendix calculates the slope factors for the naturally occurring radionuclides
under consideration. The Radionuclide Table, Radionuclide Carcinogenicity, formerly the
Health Effects Assessment Summary Tables or HEAST Tables
(http://www.epa.gov/radiation/heast/index.html), lists slope factors for individual radionuclides
or for decay chains consisting of a parent nuclide and its short-lived progeny (i.e., radioactive
daughter products with half-lives of less than 6 months). As explained below, naturally
occurring radionuclides are often associated with their long-lived decay products. The slope
factors for three naturally occurring radioactive decay series—natural uranium, Ra-226, and
Th-232—used in the present report include the contributions from these decay products. This
appendix explains the methodology used to calculate these combined slope factors.
The following guidance is excerpted from U.S. EPA 1996a:
Selected radionuclides and radioactive decay chain products are designated with
the suffix "+D" (e.g., U-238+D, Ra-226+D, Cs-137+D) to indicate that cancer
risk estimates for these radionuclides include the contributions from their short-
lived decay products, assuming equal activity concentrations (i.e., secular
equilibrium) with the principal or parent nuclide in the environment.
Note that there may be circumstances, such as long disposal times or
technologically enhanced concentrations of naturally occurring radionuclides,
that may necessitate the combination of the risks of a parent radionuclide and its
decay products over several contiguous subchains. For example, Ra-226 soil
analyses at a site might show that all radium decay products are present in
secular equilibrium down to stable Pb-206. In this case, Ra-226 risk calculations
should be based on the ingestion, inhalation and external exposure slope factors
for the Ra-226+D subchain, plus the ingestion, inhalation and external exposure
factors for the Pb-210+D subchain.
Radium-226 slope factors for the external exposure, soil ingestion, and inhalation pathways used
in this analysis were calculated according to the guidance cited above. The same logic was
applied to Th-232, whose progeny includes Ra-228, which has a half-life of 5.75 y, and Th-228,
with a half-life of 1.91 y. Since the naturally-occurring thorium at the uranium mines will be in
equilibrium with this progeny, the thorium slope factors are calculated as the sum of the slope
factors for Th-232, Ra-228+D, and Th-228+D that are listed in the HEAST tables. Natural
uranium is assumed to consist of U-234, U-235, and U-238, in ratios corresponding to natural
isotopic abundances. We first calculated a slope factor for the U-238 decay series, which we will
call U-238series, by taking the sum of the slope factors for U-238+D, U+234, and Th-230.
Radium-226 was not included, because separate soil analyses are normally performed for radium
which, due to its different chemical properties, is often not in equilibrium with uranium. In
AII-1
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similar fashion, we calculated a slope factor for the U-235 decay series (U-235series) as the sum of
the slope factors for U-235+D, Pa-231, and Ac-227+D. We then calculated slope factors for
natural uranium, by multiplying the slope factor for the U-235 decay series by the ratio of the
specific activities of U-235 to U-238 and adding this product to the slope factor for the U-238
decay series, as shown by the following expression:
Unat = U-235s
f235 + U-238s
where f235 is the ratio of the specific activities of U-235 and U-238 in natural uranium and is
shown in the following expression:
[U-235]
[U-238]
= 0.046
Because uranium concentrations in soil are commonly reported as pCi/g of U-238, the natural
uranium slope factors are normalized to a unit activity concentration of U-238. To apply these
slope factors, multiply by the actual activity of U-238, not the total activity of the uranium
isotopes.
Details of these calculations are shown in the Table AII-1 below.
Table AII-1. Calculation of Slope Factors for NORM Decay Series
Series
U-Series
Ra-Series
Th-Series
Nuclide
U-238+D
U-234
Th-230
U-235+D
Pa-23 1
Ac-227+D
Activity
Fraction
1
1
1
0.046
0.046
0.046
Total3
Ra-226+D
Pb-210+D
1
1
Total
Th-232
Ra-228+D
Th-228+D
1
1
1
Total
SF
External
(risk/y per pCi/g)
1.14E-07
2.52E-10
8.19E-10
5.43E-07
1.39E-07
1.47E-06
2.14E-07
8.49E-06
4.21E-09
8.49E-06
3.42E-10
4.53E-06
7.76E-06
1.23E-05
Ingestion
(risk/pCi)
2.10E-10
1.58E-10
2.02E-10
1.63E-10
3.74E-10
1.16E-09
6.48E-10
7.30E-10
2.66E-09
3.39E-09
2.31E-10
2.29E-09
8.09E-10
3.33E-09
Inhalation
(risk/pCi)
9.35E-09
1.14E-08
2.85E-08
1.01E-08
4.55E-08
2.09E-07
6.14E-08
1.16E-08
1.39E-08
2.55E-08
4.33E-08
5.23E-09
1.43E-07
1.92E-07
Sum, weighted by fractional activities
AII-2
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Appendix II References
U.S. EPA (U.S. Environmental Protection Agency), n/d. "Radionuclide Table:
Radionuclide Carcinogenicity - Slope Factors"
http://www.epa.gov/radiati on/heast/docs/heast2_table_4-d2_0401.pdf
U.S. EPA (U.S. Environmental Protection Agency). 1996a. Soil Screening Guidance:
User's Guide. EPA/540-R-96-018. Washington, DC: U.S. EPA, Office of Solid Waste
and Emergency Response, July 1996.
AII-3
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Appendix III. Occupational and Public Risks Associated with In-Situ
Leaching
Introduction
EPA's Science Advisory Board (SAB) recommended that EPA present information on in situ
leaching (ISL) mining operations and uranium mill operations to provide a more complete
picture of uranium production. This appendix summarizes information on environmental and
health-related aspects of ISL operations. The primary sources used for this review are,
Technologically Enhanced Naturally Occurring Radioactive Materials from Uranium Mining.
Volume 1: Mining and Reclamation Background^ U.S. EPA (2006), An Environmental
Overview of Unconventional Extraction of Uranium by Marlowe (1984) and^4 Baseline Risk-
Informed Performance Approach for In Situ Leach Uranium Extraction Licensees by Mackin
etal. (2001).
Background
In situ leaching is an extraction process that is regulated by the NRC or its Agreement States; the
waste materials and fluids are considered byproduct material (see Appendix VI of U.S. EPA
2006). However, ISL operation wells are subject to permitting under EPA's Underground
Injection Control (UIC) program (U.S. EPA 2006, Appendix VI). ISL operations, also known
within the uranium industry as "in situ recovery," or ISR, are discussed here to provide a more
complete representation of the impacts from uranium production.
ISL is used when specific conditions exist, such as the following:
• The ore is too deep to be mined economically by conventional means
• The uranium is present in multiple-layered roll fronts that may be offset by faulting
• The ore body is below the water table
• Considerable methane and hydrogen sulfide are associated with the ore
• The ore grade is low, and the ore body is too thin to mine by conventional means
• A highly permeable rock formation exists in which uranium can be economically
produced using in situ leaching
In this method of extraction, uranium ores are leached underground by the introduction of a
solvent solution, called a lixiviant, through injection wells drilled into the ore body. The process
does not require the physical extraction of ore from the ground, which makes it a much more
economical option in many cases. Lixiviants for uranium mining commonly consist of water
containing added oxygen and carbon dioxide or sodium bicarbonate, which mobilize uranium.
Other ISL facilities, especially in Eastern Europe, employ an acid-based lixiviant, though this
method is rarely, if ever, utilized in the United States. The lixiviant is injected, passes through
the ore body, and mobilizes the uranium. The uranium-bearing solution is pumped to the surface
from production wells.
The pregnant leach solution is processed to extract the uranium, usually by ion exchange or by
solvent extraction. The ion exchange process employs a resin that, once fully saturated with
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uranium, is flushed with a highly concentrated salt (e.g., sodium chloride) solution. This
reverses the exchange process and releases uranium into the solution. The uranium solution is
then sent to another process for concentration, precipitation, and drying as yellowcake. The
solvent extraction process relies on unmixable properties between the pregnant leach solution
and (uranium) solute. Normally, the solvents are organic compounds that can combine with
either cationic or anionic solutes. For example, anionic solutions include amine chains and
ammonium compounds, and cationic solutions are phosphoric acid-based. Figure AIII-1 shows a
simplified version of the ISL process.
Figure AIII-1. Illustration of ISL Process
This figure shows a simplified version of how ISL solution mining works. Lixiviant is injected into the ground
through wells on the left and far right, the fluid flows underground, dissolving uranium and carrying it in solution
until it reaches a production well in the center. The fluid carrying dissolved uranium is returned to the surface from
the production well, and piped to a production facility for refinement into yellowcake.
Injection Well
Lixiviant In
Production Well
Uranium, Lixivia:
Injection Well
Lixiviaiit In
Idealized ISL Operation
Source: Modified after AN AW A : http: //www .anawa. org. au/mining/isl-diagram .html
When the ISL process is completed, the ore body and aquifer are placed in a restoration phase, as
required by mine permits and NRC and Agreement State regulatory programs. Typically, the
aquifer must be restored to background levels where possible or practical, or to its prior
classification for water use in terms of the presence of metals, organics, pH level, and
radioactivity. Therefore, in some cases, restoring it to the pre-operation level does not
necessarily make it potable. Through the aquifer exemption process, EPA and its Delegated
States determine if an aquifer or part of an aquifer is exempt from protection as an underground
source of drinking water, because it is currently unusable as a source of drinking water and will
not serve as a source of drinking water in the future. Approval of this exemption is necessary
before a UIC permit may be issued for ISL mineral extraction wells. The aquifer exemption is
permanent, and so for some operations in some states, there is no requirement for restoration of
an aquifer, or part of an aquifer depending on the UIC permit, once it is exempted. EPA requires,
however, that non-exempted groundwater sources be protected from contaminates migrating
from the exempted portion of the aquifer.
According to Commission Order CLI-00-22, in situ leach mining (ISL) produces two categories
of waste; (1) gaseous emissions and airborne particulates resulting from drying of yellowcake,
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and (2) liquid waste associated with operations including well field processing and aquifer
restoration (Dicus et al. 1999). A variety of methods exist to address liquid waste disposal and
storage at ISL facilities, including the use of evaporation ponds, deep-well injection, land
application, and surface discharge under a National Pollution Elimination System (NPDES)
permit.
Figure AIII-2. Picture of an in situ leach field
Unlike a conventional mine, ISL operations produce minimal solid waste. This picture from the
Wyoming Association web site shows an ISL well field.
Source: http://www.wma-minelife.com/uranium/insitu/insitufr.htm
Potential Environmental and Health Issues
While the primary environmental concern from ISL operations may be related to groundwater,
Mackin et al. (2001) identify four primary risks from ISL operations in three categories:
(i) Surface environment chemical hazards
(ii) Surface environment radiological hazards
(iii) Groundwater chemical and radiological contamination hazards
The main risks to the worker are from the surface chemical and radiological hazards associated
with various types of accidents at the site. Conversely, the risks to the general public pertain to
the contamination of drinking water sources. Therefore, site-specific accidents would not affect
the public unless a large prolonged release of hazardous chemicals and/or radionuclides were
allowed to contaminate the local water supplies. In addition to hazards during ISL operation, site
rehabilitation presents environmental and health concerns. Each of these issues is discussed in
the sections to follow.
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(i) Surface Environment Chemical Hazards (Mackin et al. 2001):
Twelve chemicals are commonly used in ISL operations and could pose hazards to ISL workers,
but are unlikely to affect the general public. These chemicals, along with their intended purpose
at ISL facilities, are summarized below in Table AIII-1. Potential hazardous situations involving
each of these chemicals are discussed in the paragraphs following the table.
Table AIII-1: Typical Chemicals Found at ISL Operations
Chemical
Anhydrous Ammonia
Sulfuric Acid
Oxygen
(gaseous and liquid)
Hydrogen Peroxide
Sodium Hydroxide
Barium Chloride
Carbon Dioxide
Hydrochloric Acid
Sodium Carbonate
Sodium Chloride
Hydrogen Sulfide
Sodium Sulfide
Formula
NH3
H2SO4
O2
H202
NaOH
BaCl2
C02
HC1
Na2CO3
NaCl
H2S
Na2S
Purpose at ISL Operations
pH adjustment during uranium precipitation phase
Used to separate the uranium from the carbonate complex pumped from
below the surface
Oxidant added to lixiviant used for extraction of uranium forming UO3
Oxidant used during the precipitation phase of uranium
pH adjustment during radium removal phase
Used as a precipitant for radium during restoration and wastewater
treatment
Carbonate used to keep oxidized uranium in solution, also used for pH
adjustment of lixiviant
pH adjustment during radium precipitation phase
Carbonate used to keep oxidized uranium in solution, also used in the
regeneration/recycling resin
Used to regenerate/recycle the resin for further use in uranium
extraction
Used in groundwater restoration to decrease the solubility of various
heavy metals
Used in groundwater restoration to decrease the solubility of various
heavy metals
The main hazard posed by ammonia would be if a pipe were to break inside the processing plant.
The liquid ammonia, assumed to be under high pressure, would likely have a significant spray in
such an event and would pose a risk to the skin and eyes of any localized worker. In addition, as
the ammonia quickly evaporates, an inhalation hazard would exist that would be exacerbated by
poor ventilation. The possibility also exists for a leak in the primary holding tank or associated
piping which transfers the ammonia from outside the plant to its application site.
Similar to ammonia, a break in the pipes used to transfer sulfuric acid, sodium hydroxide, and
hydrochloric acid inside the plant would pose a hazard, as it is highly corrosive to the skin.
Sulfuric acid and sodium hydroxide would not pose a significant inhalation hazard unless the
ventilation systems in the plant were not in operation or if a worker encountered a "spray"
caused by smaller leaks in the piping system. A hydrochloric acid leak could lead to a vapor
inhalation hazard, especially in confined spaces. These chemicals are also highly reactive with
one another and so multiple localized failures, as might be the case with fire or explosions,
would cause an even greater hazard.
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Oxygen gas poses a significant hazard because of its combustible properties. Similarly,
hydrogen sulfide and sodium sulfide also pose risks because of their flammable properties, in
addition to an inhalation as well as an eye/skin irritation hazard.
Hydrogen peroxide's main risk pertains to the degradation of the chemical into hydrogen and
oxygen gas which can be caused by mechanical shock, incompatible materials, light, ignition
sources, excess heat, strong oxidants, rust, dust, or a pH greater than 4.0. Also, if the chemical is
contained within an especially rigid tank, the casual degradation of the H2O2into water and
oxygen gas would cause expansion which could rupture the holding tank. A pipe failure event
within the processing plant holds similar risks to that of ammonia and sulfuric acid.
Barium chloride is only considered a hazard if it is inhaled or ingested. Since the chemical is in
solution form at an ISL plant, this would only become significant if the worker encountered a
"spray" from a leaky piping system. Carbon dioxide from a leak can pose a risk of asphyxiation
if it occurs in a sufficiently confined space. This can be avoided if a self-contained breathing
apparatus were used when entering confined spaces where the displacement of oxygen with
carbon dioxide is possible.
Sodium chloride and sodium carbonate both are very irritating to the eyes and the skin. In
addition, sodium carbonate can pose an inhalation hazard when it is in its salt stage (dust
inhalation) or from small leaks which form a spray of the sodium chloride/carbonate solution.
Sodium carbonate also reacts readily with HC1 and H2SO4.
(ii) Radiological risks
Thickener Tank Failure
The thickener tank stores wet yellowcake slurry before it is sent to a precipitation operation and
dried into UsOg yellowcake. Thickener tank failure can pose an inhalation risk to workers if
spills are not cleaned up before the contaminants are allowed to dry. This accident scenario
would not be a significant risk to off-site residents.
The thickener tank itself does not pose any external exposure risk, as most of the uranium
progeny have been removed and the alpha component would be significantly attenuated by the
slurry. Annual external exposures have been calculated to be 120 mrem for the limiting case of a
worker standing directly next to the thickener tank for an entire 2,000 hour work year (Mackin
etal. 2001).
If the yellowcake slurry is allowed to dry after a spill incident, it would pose a significant risk of
uranium inhalation. Conservative treatments indicate that the dose to the public from a massive
spill and subsequent airborne contamination event remain below the radiation dose limits
established by 10 CFR 20 for members of the general public, however, the intake to an
unprotected worker has the potential to exceed the 5 rem annual occupational limits (Mackin
etal. 2001).
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Yellowcake Dryer Accident
As discussed above, the dried yellowcake which consists of quantities of UsOg, can pose a
significant inhalation hazard to the onsite worker when it is allowed to dry. Failure of the dryer
cake systems can stem from a number of accidents, including fire/explosion (worst case), spill
over of dryer contents due to a faulty discharge valve, failure of offgas treatment systems causing
the gases to release into the dryer area, and damage to the facility via natural disasters. It is
important to note that the failure of the yellowcake dryer systems due to natural disasters is
effectively bounded by the fire and explosion scenario. Exposures from a yellowcake dryer
accident would presumably be of similar magnitude to that of the thickener tank scenario.
(Mackinetal. 2001)
Exposure to Pregnant Lixiviant or Loaded Resin
Pregnant lixiviant and loaded uranium resin may pose a radiological hazard as an external
exposure source, and present the possibility of inhaling elevated levels of radon-222. The most
likely indoor exposure incident would occur if the pregnant lixiviant/resin were released due to a
pipe or valve failure during the ion-exchange process, at which point the solution would drain
from the ion-exchange column and the radon gas would be released to the air.
In addition to the inhalation hazard from radon, the pregnant lixiviant contains some other
radioisotopes of interest that may also cause a significant exposure. These radioisotopes are
shown in Table AIII-2, along with typical activity concentrations (Mackin et al. 2001).
Table AIII-2: Radionuclides with Typical Activity Concentrations* found in Pregnant
Lixiviant/Loaded Resin
Radionuclide
222Rn
226Ra
Natural Uranium (234U, 235U, 238U)
218p0
214Bi
214Po
Activity Concentration (pCi/L)
8.0 x 105
3.4 x 103
1.7 x 105
3.4 x 103
3.4 x 103
3.4 x 103
*Progeny assumed to be in equilibrium
Conservative treatments of a possible spill incident have been modeled to show that a maximum
annual exposure would be 27 mrem to a subject standing on a spill of infinite area and depth;
with the consideration of loaded resin, this value becomes much lower. Since such a spilling
event would likely be cleaned up expeditiously, such an exposure is not likely and is also well
within the limits established in 20 CFR 20 for the general public, as well as the site worker
(Mackinetal. 2001).
Exposures from the failure of near surface piping and subsequent runoff into containment ponds
can also pose a possible hazard to workers. It is likely that the inhalation component in this
scenario is negligible due to the dilution of the radon gas releases by ambient air; however, the
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external component would be similar to the indoor scenario previously described. See the next
section for further discussion of outdoor near-surface releases.
(iii) Groundwater Contamination Risks
Due to the nature of the ISL process (specifically the low pH and oxidation mechanisms), other
heavy metals and hazardous elements are also mobilized from the ore and can contaminate the
groundwater. These elements include the radioisotopes and progeny of uranium, thorium,
radium, and radon, as well as the non-radioactive elements such as arsenic, vanadium, zinc,
selenium, and molybdenum (for a more complete list see Table AIII-3). Because these elements
become mobilized in the target aquifer by the process of uranium extraction, it is possible for
them to migrate out of the ore body into surrounding aquifers which might feed the local water
supply. The underground propagation of this contamination into surrounding water is known as
an excursion. Horizontal excursions refer to the lateral movement of the water, while vertical
excursions indicate contamination of aquifers above and below the target ore body.
In order to detect and minimize this process, ISL facilities drill monitoring wells outside of the
main well-field at a distance sufficient to detect any excursion events, while minimizing any
erroneous indicators as a result of normal fluctuations. Horizontal excursions are more common
than vertical excursions, but do not often become problematic to the outside water supply as long
as they are detected and cleaned up within a reasonable time period. Vertical excursions are
generally a result of well casing failure (ineffective cementing of well casing), improper sealing
of abandoned exploration wells, or discontinuous or permeable natural confinement layers.
Similar to horizontal excursions, vertical excursions do not pose a significant threat unless
allowed to persist over significant periods of time—this is unlikely if geological properties of the
confinement layers are accurately characterized (to prevent downward vertical excursions), and
the well shafts are effectively cased and proper monitoring well stations have been established.
Along with well monitoring techniques, general practice at ISL facilities is to limit the injection
of lixiviant so that it is always slightly less in volume than the product solution that is pumped
out of the aquifer. This operating policy, known as "process bleed," would effectively preclude
excursions caused by overloading the aquifer, and the subsequent expansion and redistribution of
the water.
In the United States, excursions have been frequently detected by the monitoring wells located
around the well field. One of the more infamous and environmentally problematic ISL
operations was located at Irigary, Wyoming. This facility was plagued by persistent
environmental excursions which began in mid-March of 1979, and were not brought under
control until early July of that same year. The Wyoming Department of Environmental Quality
reported that these excursions were a result of the neglect of injection pressure monitoring as
well as testing the integrity of the well casings (Mudd 1998). Another significant example is the
Bruni mine in Texas, where there was a continued problem with both leachate spills and
excursions. The Texas Department of Water Resources reported that at one point during the
operational period the Bruni mine was cited for fourteen excursion incidents, while only five had
originally been reported (Mudd 1998). Despite these scenarios, no significant contamination of
local water supplies has been reported as a result of these excursions.
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In addition to the below ground excursion incidents, the groundwater can become contaminated
due to failure of the near-surface or surface piping systems which transfer the pregnant lixiviant
from the well field area to the processing facility. Typical activity concentrations for the
radionuclides present in lixiviant are given in Table AIII-2. In addition, Table AIII-3 displays
the maximum measured concentrations of non-radioactive contaminants in pregnant lixiviant
based on a survey of available licensing documents (Mackin et al. 2001). Once the pregnant
lixiviant solution is released, there are three potential outcomes for the contamination; runoff into
surface bodies of water, absorption into the soil and possible subsequent infiltrations of the
groundwater, or runoff into a surface pond designed to prevent groundwater contamination. The
first two scenarios show the possibility for contamination of drinking water sources and would
have an obvious environmental impact if not dealt with in a timely fashion. The third scenario
poses a possible radiological hazard for workers at the site and is discussed in Section ii of this
appendix.
Table AIII-3. Maximum Measured Non-radioactive Contamination in Pregnant Lixiviant
Contaminant
Arsenic
Boron
Chloride
Copper
Iron
Manganese
Molybdenum
Nickel
Selenium
Sulfate
Concentration
(mg/L)
0.3
0.2
1,800
0.04
0.02
6
62
0.09
5
1,200
Contaminant
Barium
Cadmium
Chromium
Flouride
Lead
Mercury
Nickel
Nitrate
Silver
Total dissolved solids
Concentration
(mg/L)
0.6
0.01
0.03
1
0.01
0.0001
0.09
1
0.01
5,500
(iv) Post-Operation Site Restoration and Rehabilitation
There are two main methods employed to restore the contaminated aquifer back to its
preoperational conditions. In general, the first method employed is termed "groundwater
sweep," and involves pumping out the equivalent volume of groundwater from the mined aquifer
and replacing it with fresh uncontaminated water. The volume of water pumped out of the
mined ore zone is known as the "pore volume." The pore volume can then be moved to an
evaporation pond to remove the water and then dispose of the residual wastes. An alternate
disposal of the pore volume is to inject the water into much deeper aquifers designated for waste
disposal. In this case, the increased levels of contaminant should not affect neighboring aquifers
or potential drinking water sources. This method has proven to be useful at the beginning stages
of the restoration process. However, because of the heterogeneous properties of the ore zone
aquifer, complete restoration of the mining site by this technique alone is not economical.
Furthermore, many site locations do not have the resources for the large amount of clean
groundwater that is required for an extensive groundwater sweep operation.
The second technique that can be employed is treating the contaminated pore volume via reverse
osmosis. Here, the water is pumped out of the ore zone and passed through a reverse osmosis
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membrane at high pressure. This process separates the aquifer water into a highly concentrated
liquid contaminant and a clean water volume known as the reverse osmosis (RO) permeate. The
RO permeate is then recirculated into the ore zone using alternating pumping wells to effectively
flush the heterogeneously distributed lixiviant present in the aquifer. The benefits of reverse
osmosis are that no outside source of groundwater is needed to replace the pumped pore volume,
since the volume is being treated and re-injected into the depleted ore zone. In practice, this
method can only be employed after groundwater sweeping, because the high concentrations of
contaminants during the initial stages of the restoration process tend to disrupt the RO
membranes (Davis and Curtiss 2005).
Chemicals such as hydrogen sulfide or sodium hydrosulfide may also be added to the re-injected
water during the later stages of restoration to achieve a state of "chemically reducing
conditions." The effect of these chemicals is to decrease the solubility of several contaminating
metals that are of concern, including uranium, selenium, arsenic, and molybdenum. However,
there are other contaminants, such as radium, which remain mobile under chemically reduced
conditions. Barium chloride is often used to precipitate radium out of waste water and can also
be used during aquifer restoration to mitigate the effect of radium contamination (Mackin et al.
2001).
Despite these efforts at returning the mining site to its original preoperational state, it is very
difficult to achieve complete site rehabilitation. Not all of the contamination can be removed
because lixiviant will be present in sections of the aquifer that are in areas of lower porosity.
The efforts to create a chemically reduced condition to render the heavy metals insoluble do not
apply to all contaminants of interest. Furthermore, achieving complete rehabilitation of the site
is very time consuming and costly.
Summary
In situ leaching for uranium poses several possible environmental and health-related concerns.
Through the extraction and processing of uranium ore into yellowcake, many hazardous
chemicals and radionuclides are utilized or concentrated which, coupled with certain accident
scenarios, can pose significant risk to workers at these facilities. From a radiological standpoint,
risks are mainly significant to on-site workers, and have been shown to be minimal for the public
(Mackin et al. 2001). From a hazardous chemical standpoint, the immediate concern is for on-
site workers; however, the risk to the public can become significant if a prolonged release of
hazardous material is allowed to contaminate nearby drinking water sources.
The leaching process poses the risk of contaminating neighboring aquifers which, in turn, might
affect significant water supply sources. This can happen through horizontal and vertical
excursions below the surface, or from events such as pipe failure on or near the surface. The risk
of excursions is mitigated by the inclusion of vertical and horizontal monitoring wells located
around the perimeter of the ore zone, as well as the operational practice of "process bleeding."
The wells are designed to detect excursions in a short period of time, so that corrective actions
and cleanup operations can take care of the problem before the water sources outside of the
mining site are significantly degraded.
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Finally, in situ leaching poses a problem from a restoration standpoint. Although there are
multiple techniques to restore the mined aquifer to its preoperational state, in many cases the
lixiviant can never be completely purged from the site. Attempts to bring the aquifer to a
chemically reduced state cannot account for all types of contaminants, and the entire
rehabilitation process is both expensive and time consuming.
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References:
Davis, J.A., and G.P. Curtiss 2005. Consideration of GeochemicalIssues in Groundwater
Restoration at Uranium In-Situ Leach Mining Facilities Draft Report for Comment.
NUREG/CR-6870, U.S. Geological Survey, Menlo Park, California. June 2005.
Dicus, G.J., N.J. Diaz, E. McGaffigan, Jr., and J.S. Merrifield, 1999. Commission Order: CLI-
99-22, Nuclear Regulatory Commission. Albuquerque, New Mexico.
IAEA (International Atomic Energy Agency) 2005. Guidebook on Environmental Impact
Assessment for InSitu Leach Mining Projects. IAEA-TECDOC-1428 Industrial Applications and
Chemistry Section. Vienna, Austria. May 2005.
Mackin, P.C., D. Daruwalla, J. Winterle, M. Smith, and D.A. Pickett, 2001. A Baseline Risk-
Informed Performance Approach for In Situ Leach Uranium Extraction Licensees, NUREG/CR-
6733, Nuclear Regulatory Commission, Washington, DC. September 2001.
Marlowe, J.I., 1984. An Environmental Overview of Unconventional Extraction of Uranium,
EPA 600/7-84-006, January 1984, NTIS PB84141167.
Mudd, G., 1998. An Environmental Critique of In Situ Leach Mining - The Case Against
Uranium Solution Mining. Victoria University of Technology, July 1998.
U.S. EPA (U.S. Environmental Protection Agency) 2006. Technologically Enhanced Naturally
Occurring Radioactive Materials from Uranium Mining. Volume 1: Mining and Reclamation
Background. EPA 402R-05-007, Washington, DC: U.S. EPA, 2006.
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Appendix IV. Risks Associated with Conventional Uranium Milling
Operations
Introduction
Although uranium mill tailings are considered byproduct materials under the AEA and not
TENORM, EPA's Science Advisory Board (SAB) recommended that EPA present information on
uranium mill operations, as well as in situ leaching (ISL) mining operations, to provide a more
complete picture of uranium production. This appendix summarizes information on environmental
and health aspects of uranium mill operations. The primary sources used for this review are
"Technologically Enhanced Naturally Occurring Radioactive Materials from Uranium Mining.
Volume 1: Mining and Reclamation Background" by U.S. EPA (2006), "Final Generic
Environmental Impact Statement on Uranium Milling Volume 1 and 2 " by U.S. NRC (1980), "Final
Environmental Impact Statement for Standards for the Control of Byproduct Materials from
Uranium Ore Processing (40 CFR 192) Volume 1" by U.S. EPA (1983), and "Uranium Mining and
Milling Wastes -An Introduction" by Peter Diehl of the WISE Uranium Project (2004).
Background
Uranium milling is the process of converting raw ore as it arrives from mining operations into a
product known as uranium yellowcake. The raw uranium ore and resultant yellowcake are shown in
Figure AVI-1, and a generalized schematic of a typical milling process is shown in Figure AVI-2.
The first steps in the milling process involve crushing and grinding the ore in order to obtain smaller,
uniform particle sizes throughout. Often, water is added during this stage to control dust, or lixiviant
may also be added to facilitate the extraction process. Screens separate fine particles, which continue
to the next stage in the milling process, from coarse particles, which are recirculated in the milling
circuit. Dust that is not sufficiently suppressed by the addition of water/lixiviant is generally
collected by air pollution control mechanisms, which return the fugitive particles to the milling
process.
Once the ore is ground into uniform small particles, the processed ore moves to the leaching stage.
In the most common leaching method, known as "acid leaching", uranium is removed from the
processed ore with sulfuric acid. Sodium chlorate is also added as an oxidizing agent to improve the
solubility of the uranium. An alternative approach is alkaline leaching, which is preferable when the
raw ore contains a significant portion of limestone (greater than 12%), because the acid leaching
process then requires uneconomically large amounts of acid to be effective. Alkaline leaching,
however, requires much finer grinding of the ore in comparison to acid leaching. Both methods of
leaching have similar environmental and health impacts; however, the waste produced from acid
leaching is generally more mobile and will be used as the bounding scenario in this treatment (U.S.
EPA 1983).
AIV-1
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Figure AIV-1. Raw Uranium Ore and Yellowcake Product
This figure shows the incoming raw uranium ore as it enters the uranium milling process (left), as well as the
final product of uranium yellowcake (right)
Source: http://www.eoearth.Org/upload/thumb/c/cl/Uranium ore square.jpg (left)
http://www.eia.doe.gov/kids/energy fungames/energyslang/images/vellowcakel.jpg (right)
Figure AIV-2. Generalized Uranium Mill Physical Layout
This figure shows how a uranium mill is physically set up to crush raw ore into particles amenable to chemical
treatments for extracting uranium.
Loach rig
QQQQOQ-
Source: http://www.eia.doe.gov/cneaf/nuclear/page/uran enrich fuel/uraniummill.html
AIV-2
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After the leaching stage, the pregnant lixiviant generally contains about 50-60% solids. These solids,
called "tailings," are filtered out and sent to on-site tailings piles or impoundments in the form of
sands and slimes. Once most of the solids have been removed, the filtered lixiviant is transferred to
an extraction circuit where the desired uranium is stripped from the pregnant lixiviant, followed by a
precipitation and drying process, which produces the desired yellowcake product.
Potential Environmental and Health Issues from Mill Tailings
The wastes produced during the milling process and stored in tailings impoundments are the
principal source of milling-related health and environmental hazards. Typical properties of these
mill tailings are shown in Table AVI-1. During the milling process, nearly 90% of the uranium
contained in the ore is removed, and so the primary radiological concern is the remaining progeny
associated with uranium such as thorium, radium, radon, and lead. The actual activity of these
uranium progeny can vary depending on the specific methods employed,; however, as much as 50-
86% of the original activity of the ore is retained in the mill tailings (U.S. EPA 2006). Hazardous
stable elements are also extracted from the ore and transferred to the tailings piles, including arsenic,
copper, selenium, vanadium, molybdenum, and other trace heavy metals.
Table AIV-1: Typical Properties of Uranium Mill Tailings
This table displays the chemical and radiological properties of the three classifications of uranium mill tailings
(sand, slime, and liquid). Table was adapted from U.S. NRC 1980 and found in U.S. EPA 2006
Tailings
Component
Particle
Size (um)
Chemical Composition
Radioactivity Characteristics
Sands
75 to 500
SiO2 with <1 wt% complex silicates of
Al, Fe, Mg, Ca, Na, K, Se, Mn, Ni, Mo,
Zn, U, and V; also metallic oxides
0.004 to 0.01 wt % U3(V Acid Leaching:
26-100 pCi 226Ra/g; 70 to 600 pCi 230Th/g
Slimes
45 to 75
Small amounts of SiO2, but mostly very
complex clay-like silicates of Na, Ca,
Mn, Mg, Al, and Fe; also metallic
oxides
U3O8 and 226Ra are almost twice the
concentration present in the sands
Acid leaching:* 150 to 400 pCi 226Ra/g; 70
to 600 pCi 230Th/g
Liquids
Acid leaching: pH 1.2 to 2.0; Na+,NH4 ,
SO42, Cl, and PO43; dissolved solids up
to 1 wt %
Alkaline leaching: pH 10 to 10.5; CO32
and HCO3; dissolved solids 10 wt %
Acid leaching: 0.001 to 0.01% U; 20 to
7,500 pCi 226Ra/L; 2,000 to 22,000 pCi;
230Th/L
Alkaline leaching: 200 pCi 226Ra/L;
essentially no 230Th (insoluble)
a U3O8 content is higher for acid leaching than for alkaline leaching
* Separate analyses of sands and slimes from alkaline leaching process are not available. However, total 226Ra
and 230Th contents of up to 600 pCi/g (of each) have been reported for the combined sands and slimes.
c Particle size does not apply. Up to 70 % vol. of the liquid may be recycled. Recycle potential is greater in the
alkaline process.
The five on-site environmental pathways through which these tailings impoundments pose a risk are
represented schematically in Figure AVI-3. In addition to the on-site scenarios, tailings have also
been taken off-site and used as an inexpensive building material by some local populations. Each of
these hazard pathways is listed below and the associated risks are discussed later.
(i) The release of gaseous radon-222 to the atmosphere and subsequent inhalation
(ii) Possible dust loading of contaminants from the impoundment due to natural wind
conditions
AIV-3
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(iii) The localized effect of direct external gamma radiation exposure from the tailings
impoundment
(iv) Ground seepage and subsequent contamination of local aquifers, which has the potential
to affect the water supply
(v) Dam failure due to erosion or natural disasters (flood, earthquake, etc.)
(vi) Improper use of tailings as a building material
All six of these hazard scenarios can apply to the general public and, with the exception of building
materials, to the plant workers themselves. In addition, plant workers have added risks associated
with accidents that may occur within the mill. The additional issues associated with workers are
discussed in a separate section.
Figure AIV-3: On-Site Accident and Risk Scenarios Associated with Uranium Mill Tailings
This Figure shows a visual depiction of the possible environmental and health related pathways of concern
Uranium Mill Tailings Hazards
rad on -ex halatio n
gamma-radiation
7 7 T
dust blowing
(radium, arsenic,...)
dam failure
flood
earthquake
heavy rain
groundwater
seepage
(uranium, arsenic,...) JjSSSj
Source: http://www.wise-uranium.org/uwai.html
(i) Gaseous Radon-222 Inhalation
Radon-222 is an inert radioactive gas that can readily diffuse to the surface of a tailings
impoundment where it would be released to the atmosphere. The main hazard of radon inhalation is
the damage to the lung from four of its shorter-lived decay products (Po-218, Pb-214, Bi-214, and
Po-214). Of particular concern are the two isotopes of polonium (Po-218 and Po-214), because they
produce alpha particles, which are approximately 20 times more destructive than gamma or beta
radiation. Because radon-222 has a half-life of approximately 3.8 days, it has the opportunity travel a
significant distance in the atmosphere before decaying. U.S. EPA 1983 states that the health of
populations living at a distance greater than 80 km from a tailings pile might be affected. The radon
concentration at the edge of a typical tailings pile is approximately 4 pCi/1 (WISE 2004). Using the
AIV-4
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methodology outlined in Chapter 1 of this report, a year-long exposure under these conditions would
correspond to a lifetime risk of lung cancer of l.lxlO"2.
(ii) Inhalation of Particles from Dust Loading
Dust loading occurs when wind blows over a dried portion of the tailings and dust containing
hazardous contaminants is suspended in the air. Dust loading typically becomes a hazard in the post-
operational phase of a uranium mill, as the tailings pile begins to dry, and may be exacerbated by any
de-watering treatment that is performed to minimize ground seepage [see section (iv)]. The hazards
associated with dust loading are dependent on the weather conditions and the amount of dried
material that is available for suspension. It has been estimated that a person would have to inhale 2
grams of uranium mill tailings in a year to reach the annual dose limit for the general public (100
mrem). Assuming a continuous exposure and a breathing rate of 0.9 m3/hr, this would correspond to
a dust loading of 0.24 mg/m3 (WISE 2004).
(iii) Direct Gamma Exposure
Uranium mill tailings pose an external exposure hazard from radioactivity that is present in the
waste. Although milling operations generally remove about 90% of the uranium from the ore, the
remaining waste can contain up to 86% of the original radioactivity which is mostly composed of
uranium decay products such as radium and thorium. Worst-case external exposures have been
estimated to be 0.41 mrem/h, if the subject were standing directly on top of the tailings; for a
continuous yearly exposure, this yields a dose of 3.6 rem.
(iv) Groundwater Contamination
Groundwater contamination is so heavily dependent on site-specific parameters, such as the chemical
characteristics of the waste products and soil, the location of neighboring aquifers, and the hydrology
and geology of the site, that any general numerical risk assessment of groundwater contamination is
of limited utility. Groundwater contamination can become a problem if liquid wastes from tailings
impoundments seep into the ground and are transferred into shallow local aquifers. Mills employing
acid leaching processes are of special concern, because this method renders the waste products more
soluble than an alkaline leach process. The radiological contaminants would likely be pulled out of
the seepage water into the immediate soil and so do not have the mobility to move offsite into
neighboring aquifers. However, water-soluble non-radiological hazards may be problematic,
including molybdenum, selenium, chlorine, sulfate, nitrate, arsenic, lead, and vanadium. An NRC
report (1980) concluded that 95% of any possible groundwater contamination would occur while the
site was in operation. Also, seepage should be expected unless the tailings pile was built on an
artificial liner or impermeable natural clay formations. Besides lining tailings impoundments,
milling waste is sometimes dewatered before disposal to reduce the risk of groundwater
contamination. Dewatering, however, causes an increase in the rate of radon gas emissions (increase
by a factor of 3.4 when comparing wet versus dry tailings) and also makes the pile more susceptible
to wind-driven dust loading. An example of dewatering occurs at the White Mesa Mill, where the
dry tailings are stored in an approved below-grade disposal cell. This disposal cell is covered with
the excavated earth to mitigate the effects of radon emission and dust loading (Hochstein 2003).
AIV-5
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(v) Tailings Pile Dam Failure
The least predictable risk associated with conventional uranium milling operations is the failure of a
tailings dam. A dam might fail because of poor design, natural erosion of the dam, or natural
disasters such as flooding, heavy snow fall, tornados, or earthquakes. In the United States, notable
dam failures include the 1977 spill in Grants, New Mexico (50,000 tons of sludge and several million
liters of contaminated water), and the 1979 spill in Church Rock, New Mexico (1000 tons of sludge
and 400 million liters of contaminated water). The second of these noted spill events, Church Rock,
is the most notorious. It heavily contaminated the Rio Puerco river and shallow aquifers located near
the river, which were used by the Navajo Nation as both an agricultural and domestic water source.
As of 2003, the Navajo are still unable to use this water (Ali 2003).
(vi) Improper Use of Mill Tailings as a Building Material
As stated in Chapter 4 of the main report, the risk of radiological exposure to the general public is not
only from the tailing piles themselves, but also the improper use of mill tailings as building materials.
The sandy properties of mill tailings and their availability in certain economically depressed areas
make their inclusion in concrete and use as a building material possible. This has occurred when
tailings piles have been abandoned without having been properly closed, or when piles of tailings
have fallen from trucks along rural highways. Though the problem has been documented on the
Navajo reservation in New Mexico and cited anecdotally, its pervasiveness remains unknown.
Tables 4.1 and 4.2 of the main report present annual dose values based on a few sample activity
concentrations within a Navajo hogan. See Chapter 4 of the main report for more in-depth discussion
and analysis of the improper use of tailings.
Summary of Modeled Risks to the Public
In a study by the Nuclear Regulatory Commission, a generalized case was modeled in which it was
assumed that a "low level" of environmental controls were in place. This report concluded that if the
mills in place during the time of the study (by 1980 there were 16 mills producing approximately
43,900 megatons of ore annually) were in full operation through the year 2000, it would result in
approximately 610 premature deaths in North America through the year 2100 and 6,000 premature
deaths through the year 3000. This model was based on a low level of environmental control, and
did not take into account mitigating factors, such as covering the tailings to reduce the atmospheric
release of the radon. The estimated 15-year committed dose to the public is shown in Table AVI-3,
at the end of the document, which also includes an estimate of the risk as a percentage of the risk
from normal background radiation exposure. For example, an individual near by a cluster of mills
would accrue a 15-year committed dose of 340 mrem to the lung (an effective dose equivalent* of 41
mrem), and would represent an increase of 38% above the normal risk from background exposure
(U.S. NRC 1980).
These risk estimates for fatal cancer have since been updated in U.S EPA 1983 and the results are
shown in Table AVI-2. This study estimated the individual risk of cancer for a 15-year exposure to
an individual at distances of 1,000-20,000 meters from the mill. The model also takes into account
whether the mill was in an operational or post-operational phase. For each phase of operation, the
individual 15-year risk is given as an average and a maximum value. The maximum value represents
* Effective dose equivalent based on the tissue weighting factors of ICRP-26
AIV-6
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the individual who is downwind of the mill, while the average value represents the average of all
wind directions (U.S. EPA 1983).
Table AIV-2: Results of the 1983 EPA Study" - Estimated 15-Year Risk of Fatal Cancer by Region
and Phase of Operation
Distance (meters)
1000
2000
3000
4000
5000
10000
20000
Total Risk (Operational Phase)
Average
1.12E-03
3.39E-04
1.76E-04
1.17E-04
8.48E-05
3.18E-05
1.40E-05
Maximum
1.97E-03
6.78E-04
3.60E-04
2.33E-04
1.74E-04
6.57E-05
2.76E-05
Total Risk (Post-Operational Phase)
Average
1.82E-03
5.51E-04
2.76E-04
1.89E-04
1.38E-04
5.09E-05
2.33E-05
Maximum
3.18E-03
1.12E-03
5.72E-04
3.82E-04
2.76E-04
1.04E-04
4.45E-05
"Risk estimates are derived U.S. EPA 1983 Tables 6-1 and 6-2
Additional Risks to Workers
Mill workers, beyond the six pathways described above, experience added risks associated with
accidents inside the milling facility. The hazards due to chemical spills inside the plant exist, but
may be minor relative to potential radiological accident scenarios.
At acid leaching mills, sulfuric acid is present. Though the acid is corrosive to the skin and eyes, the
leaching process is carried out at atmospheric pressure, and the risk of workers coming into contact
with a spray during a pipe failure is not plausible. If there were a fire coupled with the release of
sulfuric acid, then the inhalation of acid aerosols and sulfur dioxide could result in severe irritation of
the eyes, mucous membranes, and respiratory tract. In addition to sulfuric acid, ammonia is often
added to help control the pH level during the uranium precipitation phase. It is likely that this
ammonia would be under significant pressure, creating the risk of a spray, in the event of a pipe
failure, that poses a risk to the skin and eyes of any nearby worker. The ammonia would also quickly
evaporate, adding an inhalation hazard if the accident occurred in a poorly ventilated area.
The radiological hazards associated with milling work potentially involve the yellowcake product in
a dangerous respirable form. The two most notable accident scenarios are a thickener tank failure
where the yellowcake slurry is spilled to the floor and allowed to dry, or a yellowcake dryer accident.
Inhalation of the yellowcake particulates is a significant inhalation hazard, because of the presence of
UsOg in the cake. The reader is referred to Appendix III: Risks Associated with In Situ Leaching [see
section (ii) Radiological Hazards] for a more detailed description of operational accidents in the
milling facility, specifically those involving yellowcake.
In the NRC report (U.S. NRC 1980), it was calculated that the committed annual dose to a worker at
a conventional milling facility ranges from 2.0 rem to the bone up to 7.1 rem to the lung. These
annual doses would result in an effective dose equivalent of 240 mrem to the bone marrow (red) and
60 mrem to the bone surface and lung. Any exposures accrued because of accidental exposure to
yellowcake would be in addition to this. This information is summarized in Table AVI-3 found at
the end of the document.
AIV-7
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Summary
The primary hazard associated with conventional uranium milling operations is the high level of
radioactive contamination contained in the mill tailings (waste products). The decay progeny of
uranium are the most significant of these radioactive contaminants, including radium and radon-222,
which readily moves through the interstitial spaces of the tailing pile and is released to the
atmosphere. Once inhaled, radon and its decay progeny can cause significant damage to the lung via
alpha radiation. Other radiological hazards include direct gamma exposure from the tailings pile and
the inhalation of any dust resuspended by wind. These hazards are typically mitigated through the
use of a suitable cover over the tailing to reduce the radon released to the atmosphere and attenuate
direct gamma exposure. A suitable cover can also eliminate the risks associated with the suspension
of dust in the air.
Ground seepage of chemically hazardous constituents of tailings piles has been known historically to
contaminate nearby aquifers. Modern milling facilities often employ a liner beneath tailings piles to
prevent any ground seepage and subsequent groundwater contamination. The NRC concluded that
95% of the possible contamination would happen while the mill was operating, and that the threat
was mainly from toxic elements such as arsenic, not the radioactive constituents of the pile.
As with any industrial facility, safe management practices are critical to the safe operation of
uranium mills. Catastrophic accidents, such as a dam failure, have the potential to release large
quantities of tailings, resulting in the contamination of local water supplies and the residential
population. The improper use of mill tailings as a building material can also pose a severe
radiological risk to private individuals, particularly in tribal communities. Accidents occurring
within the milling facility could expose workers to chemical risks, and radiological risks from contact
with or inhalation of uranium yellowcake.
AIV-8
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Table AIV-3: Results of the 1980 NRC Model Uranium Mill Study - Committed Dose Values
Receptor
Dose Commitment2 (mrem)
Whole Body
Bone
Lung
Risk from Mill as
Percentage of Risk
Due to Background
(%)*'c
Nearbv Individual''
Annual 40 CFR 190 doses (excluding radon)
Imill
Mill cluster
o
J
4
45
51
30
36
—
—
Total Dose (including radon)
Imill
Mill Cluster
9.7
13
51
61
220
340
25
38
Average Individual6
Imill
Mill Cluster
0.061
0.66
0.50
5.8
1.6
16
0.19
1.9
Average Worker^
Annual
Career*
Background
450
2.1xl04
143
2000
9.3xl04
250
7100
3.3xl05
704
800
800
~
All doses shown are total annual 15 -year dose commitments except where noted as being those covered by 40
CFR 190 limits.
The range in risks due to uncertainties in health effects models extends from about one-half to two times the
central value. This range does not include uncertainties in other areas (e.g. source term estimates and dose
assessment models).
Risk comparisons are presented for exposure received during entire mill life; that is, 15 years of exposure during
operation of the mill, and 5 years of post-operation exposure while tailings are drying out, are considered. This
value is greater than that from annual exposures presented because tailings dust releases increase in the period
when tailings are drying.
The "nearby individual" occupies a permanent residence at a reference location about 2 km downwind of the
tailings pile.
The "average individual" exposure is determined by dividing the total population exposure in the model region by
its population total.
The "average worker" exposure is determined by averaging exposures expected at the various locations in the
typical mill.
The career dose is based on a person who has worked 47 years in the milling industry (that is, from ages 18 to 65).
AIV-9
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References
AH2003. Mining, the Environment, and Indigenous Development Conflicts. SaleemH. Ali. University
of Arizona Press, Tuscon. Copyright 2003.
Hockstein, Ron F., Warner, Rod, Wetz, Terry, V., Transportation oftheMoab Uranium Mill Tailings to
White Mesa Mill by Slurry Pipeline. International Uranium (USA) Corporation. WM '03 Conference,
February 23-27, 2003. Tuscon, Arizona.
U.S. EPA 1983-U.S. Environmental Protection Agency. Final Environmental Impact Statement for
Standards for the Control of Byproduct Materials from Uranium Ore Processing Volume I, EPA 520/1-
83-008-1 Washington, DC: U.S. EPA, 1983
U.S. EPA 2006—U.S. Environmental Protection Agency. Technologically Enhanced Naturally
Occurring Radioactive Materials from Uranium Mining. Volume 1: Mining and Reclamation
Background. EPA 402R-05-007 Washington, DC: U.S. EPA, 2006.
U.S.NRC 1980-U.S. Department of Energy. Final Generic Environmental Impact Statement on Uranium
Milling Project M-25, Volumes I and III NUREG-0706 Office of Nuclear Material - Safety and
Safeguards - U.S. Nuclear Regulatory Commission
WISE 2004. World Information Service on Energy - Uranium Project (WISE). 2004. "Uranium Mill
Tailings Deposits - USA" (August 15, 2005)
AIV-10
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