PUBLIC HEALTH ASSESSMENT
FORT WAINWRIGHT
FORT WAINWRIGHT, FAIRBANKS NORTH STAR BOROUGH, ALASKA
Permafrost and Contaminant Transport at Fort Wainwright
Permafrost
Permafrost is defined as ground (rock or soil) whose temperature stays below 0º C for 2 or more years. Its occurrence is influenced by terrain factors (relief, or vertical variation between the high and low elevations of an area; slope aspect, or the steepness and direction of the slope face; vegetation; snow cover; moisture content; and soil and rock type) and the presence of surface water bodies (Sloan and van Everdingen 1988; USGS 1999).
Permafrost is defined solely on the basis of its temperature; it is not necessarily a frozen-solid block. It can be dry (containing no water or ice) or contain water–frozen or unfrozen, or both. Segregated ice in permafrost can exist as lenses, layers, or ice wedges. Typically, any unfrozen water that remains in a permafrost region has a lower-than-normal freezing point (e.g., because it contains dissolved minerals) and increased viscosity. Thin films of water exist in many soil-water-ice systems, adsorbed on mineral surfaces. The adsorbed water is mobile and responds to both electrical and thermal gradients (Sloan and van Everdingen 1988).
The major components of permafrost are the:
Permafrost at Fort Wainwright
Fairbanks is in the region of discontinuous permafrost. The depth to the permafrost table varies between 0.5 to 20 meters below the ground surface, while the base is commonly 10 to 50 meters below the ground surface (Ferrians 1965; Williams 1970).
"Discontinuous permafrost" refers to a region where some areas are underlain by permafrost and neighboring areas are not perennially frozen. The unfrozen zones can be isolated or interconnected (USGS 1999). In northern Alaska, essentially all areas unaffected by human activity are underlain by permafrost. Moving southward, the percentage of unfrozen acreage increases, so that areas near the southern and southeastern coastal regions of Alaska contain essentially no permafrost (USGS 1999).
Changes from frozen to unfrozen conditions can occur abruptly across an area and might not be noticeable from the ground surface (Lawson et al. 1998).
Groundwater in Permafrost
There are three major types of aquifers in permafrost regions: 1) supra-permafrost aquifers, lying above the permafrost so that the permafrost is their lower boundary; 2) intra-permafrost aquifers, found in unfrozen zones (taliks) within the permafrost; and 3) sub-permafrost aquifers, lying below the permafrost so that the permafrost acts as a somewhat impermeable upper boundary. Each aquifer type may be found in unconsolidated deposits or bedrock (Sloan and van Everdingen 1988).
Supra-permafrost aquifers can be a useful summer-time water supply, although they can have a high concentration of natural organics (humic acid). They tend to be unreliable winter water supplies due to freezing or increasing mineralization (caused by the concentration of the minerals in the unfrozen portion of the water) (Sloan and van Everdingen 1988).
Intra-permafrost aquifers do not experience seasonal freezing. Their size remains relatively constant and is influenced primarily by long-term temperature variations. Their water quality can vary from being more mineralized and containing less natural organics than supra-permafrost aquifers, to having a high degree of mineralization. The variation in water quality depends primarily on the source water, connections to other water bodies, and climatic temperature trend (decreasing temperatures tend to cause more water to freeze and increase the mineral concentration of the remaining unfrozen water) (Sloan and van Everdingen 1988).
Sub-permafrost aquifers tend to have water temperatures above 0º C. In the region of discontinuous permafrost, where permafrost tends to be thin, sub-permafrost aquifers commonly occur in unconsolidated deposits. Sub-permafrost aquifers located in alluvial deposits below river valleys are widely used sources of water supplies (Sloan and van Everdingen 1988).
Frozen ground (frozen either seasonally or permanently) does retard groundwater movement, but it is not impermeable–it is best described as a confining material with very low hydraulic conductivity. Unfrozen water can move through porous permafrost. Groundwater flow rates depend on the overall temperature of the system, the thermal gradient, and the available cross-sectional area of interconnected films of unfrozen water. When temperature decreases, that area becomes progressively smaller, lowering hydraulic conductivity. Fractures, on the other hand, can significantly increase the overall hydraulic conductivity of a frozen section of rock or soil. Segregated ice in permafrost, which can exist as lenses, layers, or ice wedges, can reduce hydraulic conductivity to the point were the ground is effectively impermeable (Sloan and van Everdingen 1988). Presumably, this occurs for the component of the groundwater flow that is perpendicular to the lens, layer, or wedge of segregated ice.
Lateral groundwater movement occurs in the active layer during the summer (frost-free) season, in taliks within the permafrost, and in unfrozen zones below the permafrost layer. Kane and Stein (1983) measured infiltration rates for frozen Fairbanks silt loam. They reported infiltration rates of 0.0001 centimeters per second (86 millimeters per day) under low soil moisture conditions and 0.000001 centimeters per second (0.86 millimeters per day) for more moist soil.
Contaminant Transport in Permafrost
While the ice of frozen soil restricts contaminant migration, water-soluble contaminants can degrade that ice (by depressing the pore water freezing point), producing unfrozen moisture that acts as a potential transport pathway. Nonaqueous-phase liquids (NAPLs), meanwhile, tend to migrate though the unfrozen moisture surrounding soil particles (McCauley 2000).
In general, contaminant transport in frozen soil follows the same principles as in unfrozen soil. Freezing water reduces the diameter of the flow paths and blocks some flow completely. But transport can still occur through the thin films of moisture surrounding soil particles, as well as through unfrozen water. The extremely low temperatures of the system affect the fluid's density and viscosity, tending to make it heavier and thicker. Contaminants in unfrozen water can depress the freezing point to prevent the formation of ice or melt ice within their flow path.
Andersland et al. (1996) concluded from field studies that ice-saturated soil barriers with temperatures below a contaminant's freezing point depression would effectively impede contaminant flow. Biggar et al. (1998) found that diesel fuel migrated through permafrost via air voids in unsaturated fill material, through soil fissures resulting from soil contraction during freezing, and beneath punctured synthetic liners.
McCauley (2000) measured the infiltration rate of a Diesel #2/Jet A-50 fuel mixture (heating oil) in frozen soil in a field test using a double ring infiltrometer (whose inner ring was 5 feet by 5 feet). Field conditions were expected to represent ice-saturated, frozen soil. The average infiltration rate was 0.000000043 centimeters per second (0.037 millimeters per day) over the test period of almost 40 days. During the test period the infiltration rate dropped from approximately 0.0000001 to 0.00000001 centimeters per second (0.086 to 0.0086 millimeters per day).
McCauley (2000) also measured the hydraulic conductivity of frozen and unfrozen soil samples in the laboratory using a mixture of Diesel #2/Jet A-50 fuel. For unfrozen soil samples the hydraulic conductivity varied little between the three soil types tested: approximately 0.001 centimeters per second (860 millimeters per day) for organic-rich silty sand, sandy silt, or silty-sand fill material. The same range of hydraulic conductivity was measured at approximately 35 percent saturation and 100 percent saturation. At low saturation levels (approximately 40 percent, corresponding to a volumetric water content of about 15 percent), the hydraulic conductivity of the frozen soil was between 0.0001 and 0.001 centimeters per second (86 to 860 millimeters per day). For all three soil types, the hydraulic conductivity of the frozen soil dropped dramatically (and almost linearly) with increasing saturation. At 100% saturation, the hydraulic conductivity of frozen soil samples was approximately 0.000000005 centimeters per second (0.004 millimeters per day). In these tests, the soil samples were completely saturated with clean water; under natural environmental conditions, complete ice saturation might be difficult due to increasing salt concentrations in the remaining unfrozen water.
Contaminant Transport on Fort Wainwright and Discontinuous Permafrost
The work of McCauley (2000) indicates that permafrost or seasonally frozen ground cannot be expected to act as a barrier to hydrocarbon transport, with the possible exception of when the frozen soil is fully saturated with ice. For relatively large areas, it is best to assume that frozen soil represents at best a leaky confining layer, until measurements indicate that the frozen soil is acting as a true barrier to contaminant transport. All of the Fort Wainwright contaminated sites would qualify as "large" for this analysis.
Lawson et al. (1998) summarized ongoing hydrogeological investigations conducted in the north-central area of the cantonment of Fort Wainwright. Most of their work was centered on the landfill that was built around 1950 on discontinuous permafrost. They used a variety of techniques, including 1) ground-penetrating radar to outline the three-dimensional locations of permafrost and groundwater in this area; 2) boreholes drilled at test sites to define subsurface material types; and 3) groundwater monitoring wells in thawed zones located above, below, and within the permafrost to gather information about the groundwater flow direction and velocity. Their results suggest that the extent and thickness of the permafrost varies greatly across the north-central cantonment area and can cause significant deviations from the regional groundwater flow patterns.
According to evidence presented by Lawson et al. (1998) and by works they reference, Fort Wainwright lies in a flood plain created by historical versions of the Chena and Tanana Rivers. Unconsolidated material lies above bedrock that slopes downward from Birch Hill toward the Chena River. The unconsolidated material just north of the Chena and between the Chena and the Tanana is believed to have originated as alluvial deposits, most likely in the form of braided streams. These sediments may influence the reported west-northwest regional groundwater flow direction. Groundwater flow just north of the Chena, from the north-central cantonment area of Fort Wainwright to Birch Hill, may be locally diverted from the regional flow path by a north-northeast to south-southwest-trending buried bedrock valley that extends from Birch Hill toward the Chena and the discontinuous permafrost.
Similar local diversions from the regional groundwater flow path are expected to be possible for all areas of Fort Wainwright, especially those north of the Chena River; however they are not expected to affect contaminant transport for long distances from the source area. The two greatest concerns are 1) the potential for contaminants released into a supra-permafrost aquifer to migrate into the sub-permafrost aquifer via a talik, causing undetected contamination of the sub-permafrost aquifer; and 2) undetected contamination in off-base drinking water wells located along the northeastern and northwestern base boundaries. The combination of information gained from Lawson et al. (1998) and McCauley (2000) suggests that local diversions from the regional groundwater flow pattern are more common north of the Chena River, because of that area's buried bedrock gullies and higher probability of permafrost and seasonally frozen soil. The same sources lead ATSDR to expect that most of the contaminants found at Fort Wainwright (fuels, solvents, and metals) can migrate through frozen soil.
References
Andersland OB, Wiggert DC, Davies SH. 1996. Frozen soil subsurface barriers: formation and ice erosion. J Contam Hydrol 23(2):133-47. Cited in McCauley CA. 2000. Fuel penetration rates in frozen and unfrozen soils: Bethel, Alaska. Master of Science thesis. Fairbanks: University of Alaska Fairbanks.
Biggar KW, Haidar S, Nahir M, Jarrett PM. 1998. Site investigations of fuel spill migration into permafrost. J Cold Reg Eng 12(2):84-104.Cited in McCauley CA. 2000. Fuel penetration rates in frozen and unfrozen soils: Bethel, Alaska. Master of Science thesis. Fairbanks: University of Alaska Fairbanks.
Ferrians OJ, Jr. 1965. Permafrost map of Alaska. U.S. Geological Survey Miscellaneous Geological Investigations Map I-445. Cited in Lawson DE, et al. 1998. Geological and geophysical investigations of the hydrogeology of Fort Wainwright, Alaska, Part II: north-central cantonment area. Hanover, New Hampshire: U.S. Army Corps of Engineers, Cold Regions Research and Engineering Laboratory. City of Publisher: name of publisher. (CRREL Report 98-6.)
Kane DL, Stein J. 1983. Master of Science thesis: Physics of snowmelt infiltration into seasonally frozen soils. Proceedings: advances in infiltration. Chicago: American Society of Agricultural Engineers. p. 178-87. Cited in McCauley CA. 2000. Fuel penetration rates in frozen and unfrozen soils: Bethel, Alaska. Master of Science thesis. Fairbanks: University of Alaska Fairbanks.
Lawson DE, Arcone SA, Delaney AJ, Strasser JD, Strasser JC, Williams CR, Hall TJ. 1998. Geological and geophysical investigations of the hydrogeology of Fort Wainwright, Alaska, Part II: north-central cantonment area. Hanover, New Hampshire: U.S. Army Corps of Engineers, Cold Regions Research and Engineering Laboratory. (CRREL Report 98-6.)
McCauley CA. 2000. Fuel penetration rates in frozen and unfrozen soils: Bethel, Alaska. Master of Science thesis. Fairbanks: University of Alaska Fairbanks.
Sloan CE, van Everdingen RO. 1988. Region 28, permafrost region. In: Back W, Rosenshein JS, Seaber PR, editors. The geology of North America, Volume O-2: Hydrogeology. Boulder, Colorado: The Geological Society of America.
USGS (U.S. Geological Survey). 1999. Ground water atlas of the United States: Alaska, Hawaii,
Puerto Rico, and the U.S. Virgin Islands. Washington: U.S. Geological Survey. (Pub. #HA 730-A) (http://capp.water.usgs.gov/gwa/ch_n/N-AKtext1.html 
)
Williams JR. 1970. Ground water in the permafrost regions of Alaska. U.S. Geological Survey Professional Paper 696. Cited in Lawson DE, Arcone SA, Delaney AJ, Strasser JD, Strasser JC, Williams CR, Hall TJ. 1998. Geological and geophysical investigations of the hydrogeology of Fort Wainwright, Alaska, Part II: north-central cantonment area. Hanover, New Hampshire: U.S. Army Corps of Engineers, Cold Regions Research and Engineering Laboratory. CRREL Report 98-6.
ATSDR defines an exposure pathway as having 5 parts:
When all 5 parts of an exposure pathway are present, it is called a Completed Exposure Pathway. Each of these 5 terms is defined in this Glossary.
APPENDIX C: LIST OF COMPARISON VALUES
Comparison values represent media-specific contaminant concentrations that are used to select contaminants for further evaluation to determine the possibility of adverse public health effects. The conclusion that a contaminant exceeds the comparison value does not mean that it will cause adverse health effects.
Cancer Risk Evaluation Guides (CREGs)
CREGS are estimated contaminant concentrations that would be expected to cause no more than one excess cancer in a million (10-6) persons exposed over their lifetime. ATSDR's CREGs are calculated from EPA's cancer potency factors (CPFs).
Maximum Contaminant Level (MCL)
The MCL is the drinking water standard established by EPA. It is the maximum permissible level of a contaminant in water that is delivered to the free-flowing outlet. MCLs are considered protective of public health over a lifetime (70 years) for individuals consuming 2 liters of water per day.
Environmental Media Evaluation Guides (EMEGs)
EMEGs are based on ATSDR minimal risk levels (MRLs) that consider body weight and ingestion rates. An EMEG is an estimate of daily human exposure to a chemical (in mg/kg/day) that is likely to be without noncarcinogenic health effects over a specified duration of exposure to include acute, intermediate, and chronic exposures.
Reference Media Evaluation Guides (RMEGs)
ATSDR derives RMEGs from EPA's oral reference doses. The RMEG represents the concentration in water or soil at which daily human exposure is unlikely to result in adverse noncarcinogenic effects.
APPENDIX D: ATSDR'S METHODS FOR DETERMINING IF A HEALTH HAZARD EXISTS
Overview of ATSDR's Methodology for Evaluating Potential Public Health Hazards
The health hazards that could plausibly result from exposures to contaminants detected in Fort Wainwright and the area are discussed in further detail in this appendix. It is important to note that public health hazards from environmental contamination happen only when (a) people are exposed to the contaminated media and (b) the exposure is at high enough doses to result in an effect.
Selecting Exposure Situations for Further Evaluation
As an initial screen, ATSDR evaluated available
data to determine whether contaminants were
accessible to the public or were above ATSDR's
comparison values. The majority of detected
contaminants were either not accessible to the
public or fell at or below comparison values and
were not evaluated further (see text box for a
description of comparison values). Exposure
situations with contaminants above comparison
values or that had insufficient environmental data
were deemed worthy of further evaluation. These exposure situations include the following:
Deriving Exposure Doses
After identifying potential exposure situations, ATSDR further evaluated exposures to contaminants in media considering information about exposures combined with scientific information from the toxicologic and epidemiologic literature. If necessary, ATSDR derived exposure doses, which are estimates of how much contaminant a person is exposed to on a daily basis. Exposure doses are expressed in milligrams per kilogram per day (mg/kg/day). This represents the amount of contaminant mass that an individual is assumed to inhale, ingest or touch (in milligrams), divided by the body weight of the individual (in kilograms) each day. When estimating exposure doses, health assessors evaluate chemical concentrations to which people could be exposed, together with the length of time and the frequency of exposure. Variables considered when estimating exposure doses include the contaminant concentration, the exposure amount (how much), the exposure frequency (how often), and the exposure duration (how long).
Using Exposure Doses to Evaluate Potential Health Hazards
If situations are identified where individuals would be expected to come into contact with site related contaminants, during the course of their normal activities, on a regular basis, ATSDR evaluates the potential for the exposure to cause a public health hazard. When evaluating potential health hazards, ATSDR analyzes the available toxicologic, medical, and epidemiologic data to determine whether exposures might be associated with harmful health effects (noncancer and cancer). As part of this process, ATSDR examines relevant health effects data to determine whether estimated doses are likely to result in harmful health effects. As a first step in evaluatingnoncancer effects, ATSDR compares estimated exposure doses to conservative health guideline values, including ATSDR's minimal risk levels (MRLs) and EPA's reference doses (RfDs). The MRLs and RfDs are estimates of daily human exposure to a substance that are unlikely to result in noncancer effects over a specified duration. Estimated exposure doses that are less than these values are not considered to be of health concern. To maximize human health protection, MRLs and RfDs have built in uncertainty or safety factors, making these values considerably lower than levels at which health effects have been observed. The result is that even if an exposure dose is higher than the MRL or RfD, it does not necessarily follow that harmful health effects will occur.
For carcinogens, ATSDR also calculates a theoretical increase of cancer cases in a population (for example, 1 in 1,000,000 or 10-6) using EPA's cancer slope factors (CSFs), which represent the relative potency of carcinogens. This is accomplished by multiplying the calculated exposure dose by a chemical-specific CSF. Because they are derived using mathematical models which apply a number of uncertainties and conservative assumptions, risk estimates generated by using CSFs tend to be overestimated.
If health guideline values are exceeded, ATSDR examines the health effects levels discussed in the scientific literature and more fully reviews exposure potential. ATSDR reviews available human studies as well as experimental animal studies. This information is used to describe the disease-causing potential of a particular chemical and to compare site-specific dose estimates with doses shown in applicable studies to result in illness (known as the margin of exposure). For cancer effects, ATSDR compares an estimated lifetime exposure dose to available cancer effects levels (CELs), which are doses that produce significant increases in the incidence of cancer or tumors, and reviews genotoxicity studies to understand further the extent to which a chemical might be associated with cancer outcomes. This process enables ATSDR to weigh the available evidence in light of uncertainties and offer perspective on the plausibility of harmful health outcomes under site-specific conditions.
Sources for Health-Based Guidelines
By Congressional mandate, ATSDR prepares toxicological profiles for hazardous
substances found at contaminated sites. These toxicological profiles were used
to evaluate potential health effects from contamination at Fort Wainwright.
ATSDR's toxicological profiles are available on the Internet at http://www.atsdr.cdc.gov/toxpro2.html
or by contacting the National Technical Information Service at 1-800-553-6847.
EPA also develops health effects guidelines, and in some cases, ATSDR relied
on EPA's guidelines to evaluate potential health effects. These guidelines are
found in EPA's Integrated Risk Information System (IRIS)–a database of
human health effects that could result from exposure to various substances found
in the environment. IRIS is available on the Internet at http://www.epa.gov/iris
. For more information about
IRIS, please call EPA's IRIS hotline at1-301-345-2870 or e-mail at Hotline.IRIS@epamail.epa.gov.
The contaminant 1,2-dichloroethane (1,2-DCA) was detected in the Shannon Park Baptist Church and the Steese Chapel Hall wells in 1991 at concentrations greater than ATSDR comparison values for drinking water. The primary exposure pathway of concern was past exposure to this VOC is through consumption of the private well water. No exposure via consumption is occurring now because the wells are not being used for drinking water. Because church members possibly drank water drawn from the church wells in the past, ATSDR evaluated the health effects from past ingestion exposure to 1,2-DCA in drinking water.
In estimating to what extent people might be exposed to contaminants, ATSDR used protective assumptions about how long people were exposed to contaminants and how much contaminated water they ingested each day. Because some uncertainty exists regarding how long the contaminants have been in the private wells–no sampling data prior to 1991–ATSDR conservatively assumed that an adult was exposed to the contaminant for the period the Shannon Park Baptist Church well was in use (1985-1991). (Less is known about the use history of the Steese Chapel Hall well.) In all likelihood, the wells have been contaminated for less than 7 years. Adults were assumed to drink about 2 quarts (2 liters) of tap water each day and to weigh (on average for male and female) about 150 pounds (or 70 kilograms). Children were assumed to drink about one quart (1 liter) of tap water each day and to weigh roughly 35 pounds (16 kg). ATSDR assumed that private well owners obtained all their daily fluids from this private well. This is likely another conservative assumption because individuals tend to get some of their liquid requirements from sources such as milk, juice, soda, and a variety of foods. Furthermore, ATSDR assumed that church members were exposed to the most contaminated water; therefore, ATSDR used the highest (or maximum) measured concentrations of contaminants in the private well. The highest detected concentrations of 5.86 ppb was actually reported in 1995, after use of the well was discontinued. These assumptions create a protective estimate of exposure, and together, allow ATSDR to safely evaluate the likelihood, if any, that contaminants in private well water could cause harm to its users.
ATSDR also reviewed the scientific literature to further evaluate potential health effects associated with exposure to 1,2-DCA-contaminated drinking water at the detected concentrations. Most of the toxicologic and health effects information reviewed by ATSDR came from experimental animal studies or from epidemiologic investigations (human data) that examined the relationship between these contaminants in drinking water supplies and various health effects.
Applying conservative assumptions allows ATSDR to estimate the highest possible exposure dose and determine the corresponding health effects. Although ATSDR expects that few, if any, residents were exposed to the highest contaminant concentrations, the "conservative" estimates are used to protect public health. ATSDR used the following equation and assumptions to estimate exposure doses:
where:
| C |
= Maximum concentration (milligrams of chemical per liter water [mg/L]). ATSDR assumed that people ingested water containing the highest detected levels of 1,2-DCA (5.86 ppb, which equals 0.00586 mg/L). This assumption is designed to overestimate exposures. Concentrations may have fluctuate over time. As such, people may have been exposed to water that contained the maximum detected concentrations, but they also may have been exposed to water with lower contaminant concentrations or may have consumed water free of contamination. |
| IR |
= Intake rate: 2 liters per day (L/day) for adults, 1 L/day for children. The intake rate represents the amount of liquids that a person would drink in a single day. The average adult drinks 1.4 liters of water a day and the average child, aged 3 or younger, drinks 0.6 liters of liquid each day (EPA 1997). This assumption overestimates exposures because people likely obtain water from sources other then their drinking water wells (e.g., prepackaged soda or juice; bottled water; or wells serving stores, businesses, or schools). |
| EF |
= Exposure frequency: 365 days per year (day/yr). For members of the church, ATSDR assumed that exposures occurred every day, although daily exposures are unlikely since most people are expected to visit the church during hours of worship. |
| ED | = Exposure duration: 7 years (yrs) to account for the time the well was installed (1985) until the well use was discontinued (1991) following the detection of 1,2-DCA. |
| BW |
= Body weight: 70 kilograms (kg), which equals 154 pounds, for adults and 16 kg, which equals 35 pounds, for children No site-specific information is available to characterize the average weight of people living at or near The Shannon Park Baptist Church. ATSDR reviewed the scientific literature and used the U.S. Environmental Protection Agency's (EPA) recommended default weight for an adult (70 kg) and child (10 kg) (EPA 1997). |
| AT |
= Averaging time: 2,555 days (7 yrs x 365 days/yr) for non-cancer effects to adults and children. In assessing non-cancer effects, the averaging time is equal to the exposure duration. In assessing cancer effects, the averaging time is equal to a person's life span. |
Noncancer Health Effects
ATSDR compared the estimated exposure doses to dose-based CVs to assess potential non-cancer effects. Dose-based CVs (referred to as minimal risk levels [MRLs] by ATSDR and reference doses [RfDs] by EPA) are contaminant-specific doses that are conservatively derived based on the health effects literature and are below the levels associated with adverse health effects.
Table D-1 summarizes the conservative doses of 1,2-DCA estimated for consumption of drinking water and the dose-based CVs. Doses for people drinking water from the church well between 1985 and 1991 were below dose-based CVs. Because of the conservative assumptions used to estimate a dose, the true dose is expected to be even much lower than the estimated dose. As such, possible past exposures to 1,2-DCA is not expected to result in adverse health effects for adults or children who drank water from church well in the past. No exposure has occurred since 1991 when use as a drinking water supply was discontinued.
Table D-1: Estimated Non-Cancer Exposure Doses
| Contaminant | Adult (mg/kg/day) |
Child (mg/kg/day) |
Intermediate Minimal Risk Level (mg/kg/day) |
| 1,2-DCA | 0.0001 | 0.0005 | 0.2 |
Cancer Effects
ATSDR estimated a theoretical excess cancer risk expressed as the proportion of a population that may be affected by a carcinogen during a lifetime of exposure. In assessing cancer risks, scientists assume that any exposure to a carcinogen could result in a possible cancer case. However, information about the likelihood of developing cancer is based on studies where animals or humans have been exposed to high concentrations of a contaminants, levels much higher than would occur as a result of environmental releases. This assumption that any contact with a carcinogen could lead to cancer is extremely conservative. Scientists assume that the theoretical cancer risk can never be zero, whereas the true or actual risk is unknown and could be as low as zero (EPA 1996).
The theoretical cancer risk for exposures to carcinogens from the church well water in the past was below 10-4 (1 additional cancer over background in a population of 10,000)–a level used as a guideline for exposure doses that are below levels of concern. Little information exists in the medical literature about the development of cancer in people who consumed 1,2-DCA. EPA has, however, classified 1,2-DCA as a probable human carcinogen based on evidence from laboratory animal studies (ATSDR 2001). In these studies, animals were fed large doses of 1,2-DCA, but at levels at least 16 times greater than those detected in the church well water. Based on this finding, ATSDR does not consider increased likelihood of developing cancer from 1,2-DCA a concern for people who consumed water from the church well. Furthermore, ATSDR strongly emphasizes that no public health hazard exist now because no one is or expected to in the future drink water from the well.
The Fort Wainwright coal-fired power plant supplies heat and electricity for the post, and burns approximately 300 tons of coal per day during the summer and up to 1,200 tons of coal per day in the winter. The power plant appears to be a little less than 2 miles east of the western boundary of the Fort Wainwright. The Chena school is a little over ½ mile west of the power plant. The nearest boundary of the closest on-post housing is also about ½ mile west of the power plant. Another on-post housing area is located a little over 1 mile north of the power plant, while an off-base housing area is a little over 1 mile to the northwest. The average wind direction in Fairbanks is from the southwest, or southwesterly, during the summer (June-August) and from the north, or northerly, during the rest of the year (CH2M Hill 1994).
The ADEC twice issued notices to Fort Wainwright (in 1994 and 1996) for air pollution violations (EPA 1999) and EPA issued a Notice of Clean Air Act Violation to the coal-fired power plant on March 10, 1999. At the time of the 1999 inspection, the plant was operating without adequate emission controls and functioning monitors (EPA 1999). Also, from the 1960s to 1993, the coal pile used for fuel at the plant was sprayed with waste petroleum fuel products, such as diesel, fuel oil, solvents, and lubricants from tanks, railroad cars, and drums while stored at the Coal Storage Yard. The oil was used to increase the British thermal unit content of the coal and ultimately improve the output of the coal fired-power plant (EPA 1996).
No ambient air monitoring data have been identified for Fort Wainwright. Two ambient air monitors were installed near the power plant and began operating in February 2003. The north monitoring station is approximately 480 feet northeast of the plant. The south station is approximately 1500 feet southwest of the plant. Both monitors are approximately 15 feet above ground level. Results indicate these sites meet National Ambient Air Quality Standards for sulfur dioxide (SO2), nitrogen oxides (NOx), carbon monoxide (CO) and PM10 (Siftar 2003a; personal communication). Ambient air monitoring information for carbon monoxide and particulate matter for the city of Fairbanks, west of the post, was available from the EPA AirData database for the years 1996 to 2001. The measured ambient air concentrations at the nearest station (located 1 mile west of the base) never exceeded the 1-hour EPA Air Quality Standard for carbon monoxide, and exceeded the 8-hour standard only four times in 6 years. Particulate matter never exceeded the 24-hour average EPA Air Quality Standard or the annual mean air quality standard in 6 years (EPA 2002a).
The EPA air monitoring results suggest that the air quality in the Fairbanks area is not adversely impacted by carbon monoxide or particulate matter emissions; however this monitor is located about 3 miles from the power plant and several residential communities are located within one mile of the power plant. ATSDR believes that the available air monitoring data does not directly evaluate the potential public health impact of air emissions from the Fort Wainwright coal-fired power plant during normal operations or while it was also burning waste oil. ATSDR is, therefore, concerned that residential communities located near the power plant may be exposed to higher levels of pollutants than that indicated by the Fairbanks air monitor, especially when the power plant is not operating with adequate emission controls and functioning monitors.
To complete its assessment, ATSDR reviewed background information and used a screening air modeling analysis to identify if the air quality in these communities could be impacted by the power plant to a greater extent than predicted by the Fairbanks air monitoring results. We reviewed background information on the emissions from coal fired power plants, and the measured effects of those emissions on the air quality of other communities surrounding coal fired power plants, to identify if any of the contaminants released could affect the local air quality at levels that could lead to public health concerns. Limited air modeling was also conducted to identify if the on-post, or nearby off-post residential communities would be expected to have significantly higher concentrations of particulate matter than those measured at the Fairbanks monitoring station.
Typical Coal-Fired Power Plant Emissions
Fossil fuels, predominantly coal, petroleum and gas, supply about 70% of the nation's fuel requirement for electricity generation (DOE 2002a), and general information about coal fired power plants is available from a variety of sources. The following sections provide some basic information about coal-fired power plant emissions and their effect on their local community.
A variety of gases and particulates are formed during coal combustion, the type and amount of contaminants released to the air depend on a combination of factors including the composition of the coal, coal combustion conditions and the type and condition of the air pollution control equipment. The major gases emitted from coal fired power plants, in terms of quantity, include sulfur dioxide (SO2), nitrogen oxides (NOx) and carbon dioxide (CO2) (DOE 2002b). Each of these gases are a concern for different reason. Emission of SO2 is a precursor to acid rain, NOx is a precursor to both acid rain and photochemical smog, and CO2 accumulation in the atmosphere is a potential greenhouse gas linked to concerns of global climate change (DOE 2002b). Other elements and compounds are also released, but at much lower concentrations. However, due to the vast amount of coal that is burned world wide, a significant amount of research has been devoted to identifying the other constituents and understanding their fate and transport in the environment.
One of these contaminants of greatest concern is mercury. Mercury enters the coal-fired power plant as a constituent of the coal. Approximately 25% of this mercury leaves the plant through the flue gases into the atmosphere (Meij et al. 2002). In the U.S., power plants are the greatest source of mercury emissions. A study commissioned by the Maryland Department of Natural Resources estimated that the flue gas contains an average mercury concentration of 0.010 milligrams per cubic meter (mg/m3); two-thirds in the Hg(II) form and one-third in the Hg(0) form (Maryland Department of Natural Resources 1999). There are currently no limits for mercury emissions from power plants, but in 2000 the EPA proposed emission reductions (EPA 2000).
Particulate emissions are another concern; however, properly functioning air pollution control systems can remove significant quantities of airborne particulate matter. Nielsen and Livbjerg (2002) reported that approximately 99.9% of the particles leaving the boiler in the flue gas were removed by the electrostatic precipitator and the desulfurization scrubber. The particles that did make it past the air pollution control equipment were very small, 50 to 80% of the particles were less than 2.5 um in diameter (PM2.5). Several of these particles contained metals that originated in the coal (i.e., aluminum, barium, calcium, cobalt, copper, iron, lead, manganese, nickel, phosphorus, silicon, vanadium and zinc). Llorens et al (2001) performed a mass balance on the metals found in the coal prior to combustion, and in the ash after combustion. They concluded that approximately half of the incoming metals were retained in the ash; mercury and selenium were emitted almost entirely to the atmosphere. Antimony, arsenic, boron, beryllium, cadmium, chromium, lead, lithium, molybdenum, tantalum, thallium, tin, uranium, vanadium, and wolfram were distributed to both the atmosphere and ash. However, it appears that the concentration of the metals is typically small and poses a far lesser public health concern than the particulate matter.
Various polycyclic aromatic hydrocarbons (PAHs) have also been identified in the flue gas following coal combustion (Katoh 2002). The specific type and amount of PAHs emitted appear to be due slightly to the type of coal burned and is largely dependent on the combustion conditions. Efficient combustion, characterized by sufficient oxygen, appears to reduce the amount of PAHs emitted (Revuelta et al. 1999; Mastral et al. 2001).
The information reviewed suggests that a variety of gases and particulates are formed during coal combustion process, but that properly operating combustion chambers and pollution control equipment can significantly reduce the amount of pollutants emitted from coal-fired power plants.
Typical Contribution of Coal-Fired Power Plant Emissions to Air Quality
Little information is available on the contribution of emissions from a coal-fired power plant on the air quality of its community. Suarez and Ondov (2002) analyzed air samples from Baltimore, Maryland. Their results suggest that although local industrial sources were the major contributors to air pollutants, some of the particulates could be linked to emissions from coal-fired power plants. Ramadan et al. (2000) characterized the fine and course particle-size fractions collected in air samples from Phoenix, Arizona. Results suggest that the major sources of the fine particulate matter were motor vehicles, vegetation burning (wood and biomass), and coal-fired power plants. The major constituent of the particles was carbon. Soil and construction dust appeared to be the primary sources of the coarse particles.
In both cases, the information reviewed did not specify the distance between the monitoring location and the nearest coal-fired power plant. Alternate sources indicate that one coal-fired power plant is located in Baltimore County and several others are located within a 30-mile radius of Baltimore (Clean Air Task Force 2002). No coal-fired power plants appear to be located in Phoenix, however four are located in other Arizona communities (Dresher 1976).
Mercury emissions from coal-fire plants are substantial. EPA estimated that in 1994 coal-fired power plants emitted approximately 33% of the total amount of mercury emitted to the air by U.S. industrial sources. In addition to industrial emissions of new mercury into the atmosphere, mercury is also emitted by natural sources and previously deposited mercury (from industrial and natural sources) is re-emitted. As a result, EPA estimated that U.S. utilities emit approximately 13 to 26% of the total amount of mercury found in the air in the United States.
Some studies indicate that mercury emissions from coal-fired power plants do not necessarily affect local communities, but do have a global affect (Maryland Department of Natural Resources 1999). These studies suggest that the emissions are spread over a very large area so that a community is exposed to a small amount of the contaminants emitted from its own coal-fired power plant and a small amount of the contaminants emitted from all of the other coal-fired power plants in the region and possibly the world. Results of these studies loosely suggest that local air quality can be affected by emissions from coal-fired power plants located both near and relatively far from the community.
Potential Health Hazards of Coal-Fired Power Plant Emissions on Local/ and the Global Community
EPA analyzed 67 different hazardous air pollutants believed to have a potential to affect public health (EPA 1998). They studied both local and distant releases and their potential to affect a community. EPA's evaluation of the inhalation pathway indicates that coal-fired power plants within the local area of a community (less than 31 miles from the power plant) have a theoretical hazard of 0.2 additional cancer cases per 1 million exposed persons per year. This theoretical increase is very small and essentially indicates that properly operating local coal-fired power plants do not represent a public health hazard.
Some of the emissions from a coal-fired power plant are believed to be dispersed far from the local community; which suggests that the local community is also potentially affected by emissions from coal-fired power plants located well beyond its borders. The EPA analysis considering the inhalation exposure to pollutants emitted from coal-fired plants within and outside of the local community concluded that the effect of long-range pollutant transport slightly increases the potential public health hazard. The theoretical hazard for both local and long range transport of coal-fired power plant pollutants was 1.3 additional cancer cases per 1 million exposed persons per year. In actuality, this increase is very small and not expected to result in measurable increase in cancer cases in most U.S. communities.
In addition to evaluating health hazards solely from inhalation exposure, EPA evaluated four of the high-priority pollutants for their multipathway (primarily inhalation and ingestion) exposure risk. High priority pollutants are those that are persistent and/or bioaccumulate, and are toxic by ingestion or are radioactive; where non-inhalation exposure pathways could represent a greater public health hazard than the inhalation exposure. Results of the initial screening evaluation (including both inhalation and non-inhalation exposure pathways) concluded that radionuclides, mercury, arsenic, and dioxins were the pollutants with the greatest multipathway public health hazard. Though not included in the multipathway analysis, cadmium and lead were also identified as potential multipathway hazards.
EPA has also used a variety of different modeling techniques to estimate the effect of mercury exposure from multiple pathways (inhalation and ingestion), from local and long range transport. Specific results of this evaluation are highly uncertain and may be more applicable to the lower 48 states than Alaska. The results suggest that most of the mercury emitted to the atmosphere is deposited more than 30 miles away from the source. Mercury deposition to the soil and terrestrial vegetation is predicted to occur, but at levels that do not result in hazardous exposures. The major concern with mercury emissions, is that modeling assessments indicate there is a plausible link between industrial mercury emissions and mercury found in fresh water fish.
Modeling Analysis of Ambient Air from Fort Wainwright Coal-fired Power Plant Emissions
Limited air modeling using SCREEN3, an EPA approved transport and dispersion screening model, was used to get a rough idea of where the highest ground level breathing zone concentrations would be expected to occur and if concentrations in the ambient air near the school or residential housing could be impacted by emissions from the power plant. The modeling was conducted assuming that the stack height was 80 ft (24.4 m) and the diameter was 6.85 ft (2.1 m). A sensitivity analysis was conducted to identify the effect of gas exit temperature, gas exit velocity, building downdraft, and rural or urban modeling conditions significantly changed the maximum predicted concentration and the predicted concentration at a location ½ mile downwind.
The results of the sensitivity analysis indicate that the maximum concentration and the concentration ½ mile downwind decreased as the gas exit temperature was increased from 370 to 450 K. These concentrations also decreased as the gas exit velocity was increased from 0.1 to 10 m/s. The use of urban as opposed to rural modeling conditions resulted in an increase in the maximum predicted concentration (and predicted that this concentration would occur closer to the stack) and a decrease in the concentration measured ½ mile downwind. Similar results occurred when building downdraft was included.
Information obtained from the National Emission Inventory (NEI) 1999 base year emissions data suggest that the exit temperature of the Ft Wainwright coal fired power plants is approximately 458 to 490 K and that the exit velocity is approximately 7.6 to 11.6 m/s.
Assuming emission factors from EPA's AP-42 (Fifth Edition, Volume I; Chapter 1, External Combustion Sources, Bituminous and Subbituminous Coal Combustion) for SOx, NOx, CO, and PM10, ambient concentrations of these pollutants were estimated for downwind sensors located ½ mile from the source and compared with the National Ambient Air Quality Standards (NAAQS). Results suggest that under some operating conditions the downwind concentrations could exceed NAAQS. Though not definitive, this analysis does suggest that under certain meteorological conditions, the outdoor air quality in the nearby residential area could be affected by emissions from the power plant, especially when the plant is operating without the appropriate pollution control measures.
On the basis of the available information, it is not possible to identify, with certainty, if the nearby residential area was periodically exposed to contaminants released by power plant; however, this review does indicate that periodic exposures to contaminants released from the stack were possible. Without knowing specific information about the type and quantity of contaminants present in the air in the residential neighborhood and the frequency and duration of exposure, it is not possible to determine whether possible releases from the Fort Wainwright coal-fire power plant were sufficient to cause health effects in the local community. Information reviewed suggests that emissions from the power plant, however, would not be expected to adversely affect the health the local population when the power plant is operating efficiently and properly using the approved pollution control equipment known to be functioning appropriately. Fort Wainwright is currently installing additional air pollution control equipment (Siftar 2003; personal communication). ATSDR concludes that while the past exposures to air pollutants potentially released by the coal-fired power plant are indetermenent; Fort Wainwrights efforts to both reduce and monitor the emissions from the power plant will help keep the emissions within state and federal regulatory limits and be protective of public health.
Coal ash was used for snow and ice treatment at Fort Wainwright for an unspecified period until 1992 (APVR-FW-DE-ENV, 1992). While no current or future exposures are expected, ATSDR wanted to further evaluate whether people could have been exposed to potential harmful constituents in coal ash when used as road grit in the past. The following evaluation of the use of the ash as a snow and ice road treatment was performed to identify the potential public health effects from this practice.
No specific environmental data exists concerning the concentration of ash resulting in the soil, ground water or air following its application to the roads. Information was found in EPA, Alaska Department of Environmental Conservation, industrial (coal and power plant), academic and other government web-sites that describe the nature, use, and effects of coal ash and suspension of road dust into the air.
Background Information on Coal Ash
Coal ash is a residue from coal burning in power plants. Currently, within the U.S. more than 75 million metric tons of ash is produced annually by coal burning power plants. Approximately 80% of that ash is deposited in landfills and surface impoundments (Bhumbla, 1996). Not surprisingly, universities, government agencies, and private organizations have been searching for ways to recycle the ash. Proposed and existing options include: inclusion in concrete as a building material, as a soil additive for agricultural purposes, re-filling mine shafts, as a bed for new road construction, and for use in snow and ice control.
The chemical composition and particle size distribution of the ash depends on the chemical composition of the coal fuel and where the ash is deposited in the power plant. Modern power plants typically have four sources of coal ash: fly ash, bottom ash, flue gas desulfurization sludge, and fluidized bed combustion waste. Typically, the coal is pulverized to the consistency of powder prior to combustion, and then it is blown into the boiler to be burned. The fine particulates, such as fly ash, rise with flue gases and either settle elsewhere in the system or flow out of the stack. The larger particles fall to the bottom and are removed as the bottom ash (CPM Inc., 2000). Power plants that use flue gas desulfurization or fluidized bed combustion processes to reduce sulfur emissions, tend to have high concentrations of sulfur in the ash from those locations.
About 16% of the coal ash produced by power plants is bottom ash (Kalyoncu, 2001). Bottom ash tends to resemble shattered glass and can be abrasive (Flygt, 2003). Several sources identified that bottom ash was used for snow and ice control (TFHRC, 2003; CPM Inc., 2000). The U.S. Department of Transportation, Federal Highway Administration acknowledged that bottom ash is used for snow and ice control at various areas around the country; however, they do not recommend the practice because the ash tends to have a low pH and may be corrosive to metals (DOT, 2002).
Coal Ash Use at Fort Wainwright
Though not specified in the documentation from Fort Wainwright, ATSDR believes that the source of the road grit was the bottom ash. Fort Wainwright documents did not identify how much coal ash was applied at a time, the frequency of applications over the winter months, or which roads were treated with the coal ash.
Fort Wainwright provided results of laboratory analyses of the coal composition, leach tests performed on fly ash and ash samples (this type of test is describe in more detail in the section evaluating the tar sites), and 'ash' composition. ATSDR used just the ash composition test to estimate the composition of the coal ash used as road grit. Table D-2 identifies the range of the various metals measured in the ash and compares the measured values to the common range of that metal in soil and the ATSDR comparison value.
Public Health Evaluation of Potential Past Exposure to Coal Ash at Fort Wainwright
ATSDR considered three potential public health concerns related to the previous practice of using the coal ash as road grit. First, potential contamination of the soil lying along the road due to metals in the ash and subsequent ingestion of that soil by Fort Wainwright residents. Second, infiltration of contaminants to the groundwater and potential ingestion with drinking water. Third, inhalation of contaminants due to suspension by vehicular traffic. To evaluate these concerns, ATSDR gathered background information about typical concerns associated with coal ash applied to ground surfaces to understand the theoretical potential of coal ash to contaminate ground water and affect public health. In addition, ATSDR used results of the previous laboratory tests of ash and estimated application rates to evaluate the potential of the coal ash used as road grit to affect public health.
Table D-2. Concentration of Metal in Coal Ash (ppm)
| Metal | Measured Concentration by EPA 6010 | Measured Concentration by EPA 200.7 [ppm] or EPA 7471 | Common Range for Soils 2 | ATSDR CV for Soil |
| Arsenic | 21.9 | 13.0 | 1 - 50 | 20 3 |
| Barium | 2543 | 4037 | 100 - 3000 | 4000 3 |
| Cadmium | 1.5 | 1.7 | 0.01 - 0.70 | 50 3 |
| Lead | 59.6 | 61.9 | 2 - 20 | 400 4 |
| Mercury | 2.5 | 5.02 1 | 0.01 - 0.3 | 6.7 5, 6 |
| Selenium | <5.3 | <4.4 | 0.1 - 2 | 300 3 |
| Silver | <2.1 | <1.8 | 0.1 - 5 | 300 3 |
| 1 Mercury
was analyzed by EPA method 7471; all others by EPA 200.7. 2 EPA 1987. 3 ATSDR's reference dose media evaluation guide (RMEG) for a child. 4 EPA, Region 9, Preliminary Remediation Goals (PRGs) 5 Arizona Department of Environmental Quality, soil remediation levels. 6 The WHO permissible tolerable weekly intake is 5 µg/kg; assuming a 20 kg child was eating this material, he would be able to consume up to 20 grams (g) of ash ([5 µg mercury /kg bodyweight * 20 kg bodyweight]/[5 ug mercury g/g ash]). The density of bottom ash is approximately 2.5 grams per cubic centimeter 7; 20 g of ash would be approximately 50 cubic centimeters (about 1.7 ounces, 10 teaspoons, or a little over 3 tablespoons). 7 TFHRC. 2003. |
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Potential Ingestion Exposure to Ash Metals in the Soil
ATSDR considered the potential for the metals measured in the ash to be ingested by children playing near a roadway that had been treated with bottom ash as road grit. Concentrations of the metals are below their ATSDR comparison value (CV) for that material in soil. Both ATSDR and U.S. Environmental Protection Agency (EPA) do not have a CV for elemental mercury in soil. The concentration of mercury in the ash is below the reported soil remediation level for Arizona. In addition, significant quantities of the ash would need to be consumed on a regular basis to exceed the World Health Organization's (WHO) standard mercury ingestion. Ash transported off of the road surface by rain or vehicular traffic is expected to ultimately end up in the post's storm water discharge system or land on neighboring property landscape. Once the ash was added to the roadway the effective concentration of the metals would be reduced as the ash mixed in the surrounding soil. Therefore, ATSDR concludes that use of the bottom ash as road grit material did not constitute an ingestion hazard for post residents or visitors.
Potential Exposure to Ash Metals due to Groundwater Contamination
ATSDR considered the potential for metals measured in the ash to infiltrate to the underlying groundwater, potentially contaminating the groundwater in that area. Again, because the metal concentrations measured in the ash are within ATSDR CVs for soil, it is not expected to have changed the groundwater concentration for these metals. ATSDR concludes that use of the bottom ash for road grit did not cause a drinking water hazard.
Potential Exposure to Ash Metals due to Inhalation
ATSDR considered the potential for the metals measured in the ash to become airborne following vehicle traffic on the roads. This evaluation was performed using the format described by EPA to estimate the particulate concentration following vehicle traffic on a paved road. ATSDR assumed the most conservative, yet realistic, exposure scenario which considered air concentrations along a two-lane road that had recently been treated with bottom ash. Under this scenario, sufficient material would have been available and passing car traffic ground the grit to a small enough size that would permit it to become airborne. Other assumptions used in our evaluation include: the roads were dry, grit is not held to the road surface by water or ice; traffic was relatively heavy for a short period of time (vehicle speed of 35 miles per hour, one car every 2 seconds for 15 minutes), similar to what might be expected near a road during rush-hour; all vehicles were similar to a 2003 Ford F-150, 4×4 regular cab style side truck, 4.6L V-8, 5-speed manual transmission, curb weight of 4431 lbs (weighing about 2.2 tons) (Ford, 2003); the road had two lanes (one for each direction); vehicle traffic is primarily in one direction for the 15 minute time period; the road does not have a parking lane or curb, the road surface is even with the landscape; there are no openings to the storm water drainage system along this section of road.
EPA estimates the particulate emissions from vehicle traffic on a dry paved road by:
E = k (sL/2)0.65 (W/3)1.5, where E = the particulate emission factor [g/VMT] k = particle size multiplier for particle size of interest [g/VTM] sL = road surface silt loading [g/m2] W = average weight of vehicles traveling the road [tons] VMT = vehicle miles traveled
This emission factor includes the emissions from the exhaust, brake wear, and tire wear that were released during test while developing this formula. ATSDR applied this equation using a low value of the silt loading to simulate clean road conditions and then a higher silt loading simulate worst-case road conditions following ash application. Ideally ,silt loading would be estimated based on tests of the on-post roads during the conditions simulated; however, that information is not available. Other researchers though have looked at the relationship between measured silt loading, air concentrations of PM, and the resulting emission factor for a variety of locations. Clean paved roads appear to have a silt loading of about 0.1 to 1 gram per square meter (g/m2); uncleaned city roads appear to have a silt loading of 1.0 to 3.6 g/m2 while uncleaned industrial roads can be over 10 g/m2. ATSDR assumed that clean road conditions at Fort Wainwright could be simulated by using a silt loading of 0.5 g/m2 while the silt loading following ash application could be simulated by a silt loading of 20 g/m2.
Table D-3 shows the values used for the calculation of the emission factor (EF) for airborne particulate material with an aerodynamic diameter of 10 micrometer (µm) (or PM10) or less, for both conditions and the resulting values. The emission factor (EF) calculated for the ash treated roads assumes that all of the applied bottom ash is immediately ground into silt-sized particles. If this were true, the ash would not be a useful method of improving traction on snow and ice covered roads. In addition snow melt, rain and traffic is expected to transport a fraction of the applied material directly to the landscape on the side of the road. This combination of factors is expected to reduce the EF value for the ash treated road; however we were not able to identify from the literature a reasonable reduction.
Table D-3. Calculation of the PM10 Emission Factors for Clean and Ash Treated Roads following the EPA estimation procedure
| Variable | Clean Road | Ash Treated Road |
| k [g/VMT] | 7.3 | 7.3 |
| sL [g/m2] | .5 | 10 |
| W [tons] | 2.2 | 2.2 |
| E [g/VMT] | 1.9 | 13.0 |
A literature search was able to identify several researchers who have measured the PM10 concentration along a road, some have attempted to quantify the change in the measured concentration with distance from the road. Laxen et al. (2002) measured PM10 concentrations adjacent and at varying distances, up to 50 meter (m), from the edge of the road. For single lane roads the results suggest a decreasing trend from approximately 30 microgram per cubic meter (µg/m3) near the road to approximately 20 µg/m3 at a distance of 50 m from the road. The results for larger roads, with more lanes, showed a relatively constant concentration of approximately 25 µg/m3 from the roadside to approximately 95 m from the road. Researchers in England reported PM10 measurements taken near roadsides for six different roads in the Newcastle upon Tyne area. The mean 24-hr PM10 concentrations varied from a low of 15 µg/m3 to a high of 85 µg/m3. The Newcastle upon Tyne area has several significant sources of particulate emissions, therefore the concentrations presented here are a combination of background sources, vehicle emissions and suspension of road silt (Newcastle City Council, 2000).
Fitz (1998) measured the PM10 concentrations upwind and downwind of a road, before and after street cleaning to investigate the effect of silt loading on local air quality. The differences between the upwind and downwind concentrations were near the uncertainty of the method, suggesting that the increase in PM10 concentration due to the vehicle traffic was not discernable.
The Alaska Department of Environmental Conservation (ADEC) measured PM10 in a parking lot of the Lemon Creek Valley in Juneau between November 1993 and April 1994. The purpose of the monitoring was to check if PM10 released from wood-burning stoves and vehicle traffic on nearby sanded roads could impact the air quality of that area. Those results indicate that ambient air quality was within all National Ambient Air Quality Standards (NAAQS), suggesting that neither wood-burning stoves or sand treated roads adversely impacted the local air quality (ADEC 1996).
Results from these studies suggest that for areas with a large background sources of PM10, the reduction in PM10 concentration from the road is not as significant for areas with fewer background sources. In areas with few background sources of PM10, well traveled roads can be significant local sources of PM10. Compared to many areas, Fort Wainwright has only a few background sources of PM10, notable sources include the power plant and active flight line. It is expected that the background PM10 concentration on Ft Wainwright would be similar to those measured in Lemon Creek Valley, that had an average PM10 concentration of 7 µg/m3 and a maximum 24-hr concentration of 51 µg/m3 (the second highest 24-hr concentration was 21 µg/m3). These values are within the PM10 24-hr concentration ranges reported for other areas near roadways without grit treatment; 15 to 85 µg/m3. All of these values are well within the NAAQS 24-hour standard of 150 µg/m3.
Therefore, although it is not possible to know the actual PM10 concentration near the road following ash treatment, it is likely that the PM10 concentration was within NAAQS levels. ATSDR also considered the composition of the PM10 by assuming the percent of metals in the suspended particulate matter would be similar to that measured in the coal ash. Table D-4 shows the percentage of each metal measured in the coal ash and the estimated concentration of the metal in the air near the roadway, assuming that a roadway PM10 concentration of 50 ug/m3. The final column of the table shows the ATSDR CV for that metal in air.
Note that the results shown in this table are purely estimates, no actual measured data exist. It is merely an attempt to identify if ground ash from treated roads could have had excessively high concentrations of metals. Typically, the CV for a particular chemical represents the average concentration people can be exposed to for long periods of time, on a regular basis, where toxicological data indicate that adverse health effects are not likely. The CV presented for lead does not follow this definition and is not truly appropriate for this analysis. The value is from the Occupation Health and Safety Administration (OSHA) guidelines. It provides the concentration that healthy adults, working 8-hr shifts for 5-days per week, can be exposed to without adverse health effects. There is no CV for residential lead air concentration, the goal is to reduce and eliminate residential lead exposure because lead can affect mental and physical development of children. The estimated lead concentration in the air near the road is over 16,000 times less that the OSHA limit. While it is always best for children and adults to avoid unnecessary lead exposure, this estimated concentration is not likely to have caused adverse health effects.
Table D-4. Estimated Metal Concentration in Air Near Ash Treated Roads
Assuming PM10 Concentration of 50 µg/m3
| Metal | Ash Composition 1 [%] |
Estimated Air Concentration [µg/m3] |
ATSDR CV |
| Arsenic | 0.002 | 0.001 2 | 0.0002 3 |
| Barium | 0.4 | 0.2 | 0.52 4 |
| Cadmium | 0.0002 | 0.0001 | 0.0006 3 |
| Lead | 0.006 | 0.003 | 50 7 |
| Mercury | 0.00005 | 0.00025 | 0.2 5 |
| Selenium | <0.0005 | <0.00025 | 18 6 |
| Silver | <0.0002 | <0.001 | 18 6 |
| 1 Highest
of the EPA 6010, EPA 200.7 or EPA 7471 method, converted to percent from
ppm by: Percent = (ppm/1,000,000)*100. 2 Estimated Air Concentration = (50 ug/m3)*(.002/100%) = 0.001 µg/m3. 3 ATSDR's cancer risk evaluation guide (CREG). 4 EPA Region 9, PRG. 5 EMEG/MRL. 6 EPA Region 3, RBC. 7 OSHA; note this is an occupational 8-hr time weighted average exposure limit, not necessarily protective for residential exposures. |
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All of the other estimated metal concentrations are well below ATSDR's CV except for arsenic. Given that the estimated concentration is a result of high vehicle traffic, the concentration will decrease with distance from the road. The CV typically describes a concentration that an individual could live with for long periods of time without any adverse health effect. Short periods of time spent at higher concentrations also do not lead to adverse health effects. Again, while it is best to avoid unnecessary arsenic exposure, this estimated concentration is not likely to have caused adverse health effects. As a result of this evaluation, ATSDR concludes that use of bottom ash from the coal power plant did not present an inhalation health hazard for the residents of Fort Wainwright.
Four sites have been identified on the main cantonment area that contain tar apparently emerging through the soil surface. The public health concerns of these sites are centered on understanding 1) what harmful materials exist in the tar areas, 2) whether on- or off-post drinking water wells could be affected by the tar, and 3) whether on-post residents could incur exposure to the tar at levels of adverse health effects. The following discussion presents an overview of tar and then address each health concern individually.
Tar Sites at Fort Wainwright
The four tar sites are known as the Southgate Road, Glass Park, Chena River, and the Power Plant tar sites. Table D-5 describes the location of these sites. Each tar site appears to encompass several hundred-thousand square feet of area. Glass Park and Chena River sites are located along the river, near both obvious industrial areas and/or lightly developed areas with pedestrian access. The other two sites are both located south of the runway; approximately 1,000 feet apart. (Fort Wainwright, 1991).
Table D-5. Location of Tar Sites at Fort Wainwright
| Tar Site | Location |
| Southgate Road | Located west of the South Post soccer field on the South Post parade field, near Southgate Road. |
| Glass Park | Located near the western boundary of the main cantonment area, south of the Chena River near Building 4040. |
| Chena River (Golf Course Site) |
Located northwest of the Post Golf Course, on the north bank of the Chena River. |
| Power Plant (Railroad Tracks) |
Located west of the Coal-Fired Power Plant cooling pond, next to the railroad tracks. |
Material at the Tar Sites
The tar at the four sites described in Table D-5 is believed to be the result of previous disposal of construction-related material, specifically road tar (USACE, 1995). It is unclear, however, as to when the tar was disposed of at the sites and what, if any, other material was also disposed at these sites. A search of the Hazardous Substances Data Bank (National Library of Medicine, 2003) indicates that same basic tar-like material can be used for either roofing or road paving construction. This material can be identified by a variety of different names including: asphalt, asphalt cement, bitumen, petroleum roofing tar, pitch, road asphalt, and road tar. While the term tar will be used in this document to refer to this asphalt material, it really is a misnomer. In actuality, true tars are produced by destructive distillation of coal, oil or wood, and asphalt is a residue of fractional distillation of crude oil (National Library of Medicine, 2003). These true tars are not used in construction and are not likely present at the Fort Wainwright tar sites.
Depending on the distillation process, three different types of asphalt may be obtained: paving asphalts, roofing asphalts, and asphalt-based paint. Most of the asphalt produced in the U.S. is used for paving and roofing. The major constituents of asphalt include: aliphatic compounds, cyclic alkanes, aromatic hydrocarbons, and heterocyclic compounds containing nitrogen, oxygen, and sulfur atoms (NIOSH 2000). Based on the available information, it appears that the basic composition of the tar is similar for both paving and roofing applications, although the two materials are different and not interchangeable.
When used for road paving the resulting asphalt material consists of approximately 95% (by weight) of aggregate (crushed stone, gravel or sand) and 5% asphalt cement (tar) that acts as a glue to hold the pavement together (NAPA, 2002). Typically, locally available aggregate material is used for paving operations (NAPA, 2002). No materials associated with construction debris, such as cans of primer, paint, paint thinner, and other wall preparations, nails, lumber, asbestos, and dry wall, have been identified as being discovered in, or near, any of the tar pits.
On the basis of the available information, ATSDR concluded that tar (asphalt cement) would be the major contaminant source and was the only material considered during the evaluation of the four known on-post tar sites.
Potential Impacts of Tar to Drinking Water Supplies
Drinking water wells are located within a mile of two tar sites. Two heavily used post drinking water wells are located about 4,700 feet to the west of the Power Plant tar site and abut 4,500 feet northwest of the Southgate Road tar site. A lesser used well is located 1,000 feet north of the Power Plant tar site and a drinking water well used by Burger King is located about 500 feet south from the Southgate Road tar site. The general direction for groundwater flow for this portion of the post is towards the northwest.
In 1992, tar samples were collected from each of the four main tar sites and analyzed by the Toxicity Characteristic Leaching Procedure (TCLP). The resulting leachate was analyzed for semi-volatile organic compounds (SVOCs), volatile organic compounds (VOCs), pesticides, 8 RCRA metals (arsenic, barium, cadmium, chromium, lead, mercury, selenium, and silver), and polychlorinated biphenyls (PCBs). The sample collected from the Southgate Road Tar Site showed trace amounts of the VOCs ethylbenzene, toluene, and total xylenes; however, SVOCs, pesticides, RCRA metals, and PCBs were not detected. Samples from the Glass Park Tar Site, the Chena River Tar Site, and the Power Plant Tar Site all were non-detect for SVOCs, VOCs, pesticides, and PCBs. Barium was detected at trace levels in these three sites (as high as 0.48 part per million [ppm]; below TCLP regulatory limits and ATSDR's CV for drinking water) and selenium was detected at trace levels (0.074 ppm; below TCLP regulatory limits but slightly above ATSDR's drinking water CV for children [0.050 ppm]) at the Power Plant Tar Site (USACE-AK 1992). The measured concentrations were well below the TCLP regulatory limits of 100 and 1 ppm, respectively (40CFR261.24).
Results of those tests were used to determine that the tar would not leach contaminants to the underlying groundwater (USACE, 1995). Based on the TCLP results, the Army, EPA and Alaska Department of Environmental Conservation (ADEC) concluded that there was "no evidence that a potential source of contamination exists at these sites" and signed the No Further Action (NFA) recommendation in 1994, with the comment that future action at these sites should be coordinated with the Solid Waste/Pollution Prevention program of ADEC (USACE, 1995).
TCLP is a laboratory test to simulate leaching from a material that could occur in a municipal landfill. This test is not designed to identify the constituents of the material itself, but identify the probable components of the leachate resulting from a short-term exposure (18 +/- 2 hours) to a low to moderate pH solution. The TCLP test is based on a scenario that the greatest potential for the material to contaminate groundwater, following disposal in a municipal landfill, would occur if the material were disposed with other wastes that could acidify the infiltrating rain water. Under this scenario, acidified infiltration water would be expected to generate the greatest concentration of contaminants in the leachate. If the contaminant concentration of the laboratory derived leachate is within the recommended standards, it is assumed that the material will not release significant amounts of contaminants that could affect local drinking water under the scenario described (EPA Method 1311; Kimmell, 1999). If the laboratory derived leachate exceeds the recommended standards, the material must be treated and disposed of as a hazardous waste.
ATSDR recognizes the value of the TCLP test to categorize a material as hazardous waste and predict its potential to contaminate infiltrating water under certain circumstances. However, ATSDR also recognizes the limits of this test's ability to estimate the material's ability to contaminate the underlying groundwater at a specific site. There are several site specific characteristics that could influence how much of a compound could be released by the tar to infiltrating water, and how much could actually reach the underlying water table. Site characteristics that could influence the amount of chemical released by the tar to the infiltrating water include: the amount of tar at the site, amount of infiltration (during a particular rainfall, or snowmelt event, and that occurring over the course of a year), infiltration rate, and infiltrating water characteristics (i.e., pH). Other site characteristics would be expected to influence the amount of released chemical that actually reaches the water table. These include: site geology (i.e., whether infiltrating water flows straight down to the water table or meanders slowly through the soil, and if the released chemicals are sorbed to the soil particles so that their movement is retarded in comparison to the infiltrating water), and depth from the tar to the water table. Given the difficulties in using the TCLP results to identify and estimate the concentration of contaminants in the groundwater beneath, and downgradient, of the tar sites ATSDR would prefer to evaluate results of groundwater sampling from monitoring wells located beneath or just downgradient of the tar sites when assessing potential health hazards.
Limited groundwater monitoring data are available for these sites as follows:
Glass Park Tar Site. A monitoring well near the Glass Park tar site (AP-12/5525) appears to be located between the tar site and Building 4051. An underground storage tank (UST 300) exists near the building, but no environmental data for this site was identified. The expected groundwater flow direction for this area is toward the river. Therefore, the monitoring well is likely upgradient from the tar site (groundwater would likely flow first past the well and then towards the tar). At this site, the river surface is about 10 feet below the edge of the park; basically the tar site is on the top of a small cliff overlooking the river. Samples taken from this well during June and July of 1990, suggest barium concentrations in the groundwater could be above ATSDR's comparison values (CVs) for drinking water. However, the source of the barium is not known and could be due to natural conditions or influenced by the sources other that the tar site. This water is not used for drinking so there is no public health concern with the elevated levels of barium or other contaminants.
Chena River Tar Site. One monitoring well appears to be located within the tar site (AP-26/5538), another appears to be located between the tar site and the river bank (AP-27/5539). The Beacon Tower Landfill is located about 1,000 ft northeast of the tar site. This area was reportedly used as a landfill during the 1950s and 1960s. A NFA recommendation was signed in June 1992 because site inspections did not find enough evidence to indicate that the landfill actually existed. If the landfill does not exist, contaminants measured in these two wells are presumably due to natural processes or are influenced by the tar. The well believed to be within the tar site was sampled for metals, volatile organic compounds and semivolatile organic compounds in 1989; arsenic, barium, and lead were detected above ATSDR CVs. The well believed to be outside of the tar site had no contaminant detected above ATSDR CVs. This water is also not used for drinking so there are no public health concerns associated with the metal concentrations measured.
Power Plant and Southgate Tar Sites. These sites are located within 1 mile of drinking water supply wells. No monitoring wells are located within, or immediately downgradient of the Power Plant or Southgate Road tar sites. The closest downgradient wells appear to be approximately 2,000 to 3,000 feet away and have the potential to be impacted by other sources, such as the power plant coal storage yard. Given the results collected at the other tar sites, it seems likely that the Power Plant and Southgate Road tar sites also have the potential to increase the concentration of some metals in underlying groundwater. (It is important to note, however, that concentrations of metals in the groundwater could be the result of natural processes in this area.) Based on the relative position of these two tar sites and general information about the pumping rates used by the supply wells (provided in the draft wellhead protection plan), ATSDR believes while the dominant groundwater flow direction is likely still toward the northwest, significant variations to that direction can occur both temporally and spatially. If the contaminants are infiltrating from the tar, however, it is likely that the concentrations will decrease rapidly as a result of dispersion, or mixing, of the tar-contaminated water with distance from the site.
Overall, there is not enough available information to determine whether or to what extent contaminants are leaching from the Power Plant or Southgate Road tar sites (site within a mile of the drinking water supply wells) at a level that could affect groundwater quality. The Fort Wainwright Department of Public Works, however, has an active program to test the on-post drinking water supply. These results indicate that the post drinking water has met safe drinking water standards. Based on the expected results of routine drinking water quality testing program, ATSDR concludes that the tar sites are not a public health hazard for the post's drinking water supply.
Potential Hazards from Contact with Material at the Tar Sites
Public access to these sites is not restricted by fences, and signs are not posted to limit activities at the tar sites. Some sites, such as the Southgate Road tar site, are even easily accessible by pedestrians. ATSDR reviewed the toxicologic literature on health effects from contact with tar to identify if routine contact with the tar could be expected to cause health effects. Laboratory tests of raw roofing tar applied to the skin of mice yield conflicting results. One study concluded that the roofing tar led to malignant skin tumors in the treated mice, yet the other study concluded roofing tar was not carcinogenic. However, these tests were performed to identify potential hazards to occupational exposure of the tar material. Limited data are available that describe potential human health effects resulting from dermal contact with tar and are limited to the exposure of workers to the tar, particularly, molten tar.
Occupational exposure to the fumes have been reported to cause dermatitis, acne, photosensitization, and dermal melanosis (POISINDEX(R), 1993). The primary immediate hazards for workers are burns to the skin by direct contact with the molten tar, and nose and throat irritation due to inhalation of fumes from the hot tar. A few studies did find non-dermal issues, such as a slight elevation among asphalt workers for respiratory diseases (Randem et al., 2003a), lung cancer (Randem et al., 2003b, Stucker et al., 2003), or stomach cancer (Stucker et al., 2003); though these researchers are quick to point out that the potential increases are very small and could be confounded by other lifestyle factors such as smoking. Other studies suggest that there is no significant adverse health effects from these occupational exposures (POISINDEX(R), 1993, Bergdahl and Jarvholm, 2003). Once the tar cools, these types of exposures are no longer possible. Cooled tar is relatively non-toxic (POISINDEX(R), 1993).
The cooled tar in the Fort Wainwright tar sites is not in a form that would allow it to release significant amount of vapors or coat the skin following dermal contact. Repeated, long term contact with this material such that the tar could coat the skin is not likely given the limited accessability of the general public to this material. ATSDR noted during site visits that the tar material will stick to certain items; however, individuals could walk across the tar and touch the tar without getting residual material on shoes or skin. Due to the climate, snow cover and heavy clothing would prevent direct contact with the material for months at a time. During the time when direct contact is possible; it is expected to be occasional and short term. Although the material is easily accessible, it is expected that people will not have continued, direct contact often enough to cause health effects. Therefore, ATSDR concludes the tar sites do not present a public health hazard to on-post employees, residents or visitors. However, ATSDR supports ADEC's NFA request that all future actions involving the tar sites be coordinated with the Solid Waste/Pollution Prevention Program of ADEC.
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References for Coal Ash
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ATSDR. 1999. ATSDR Toxicological Profile for Mercury. Atlanta: U.S. Department of Health and Human Services.
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South Coast Air Quality Management District under Contract 96018, January. 98-AP-RT4H-005-FR. Riverside, CA: Center for Environmental Research and Technology. Accessed on June 26,
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References for the Tar Sites
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Fort Wainwright. 1991. Groundwater Monitoring Well Location Map.
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APPENDIX E: RESPONSE TO COMMENTS
ATSDR received comments from two different sources for the Public Comment Public Health Assessment for Fort Wainwright. All of the comments provided clarification of, or editorial changes to the text. ATSDR incorporated those comments into the text.
ATSDR did receive one comment from a community member, it is addressed below.
Comment
A number of years ago, I believe in 1978 I was employed on the army base. One day, we were directed to load up a huge flatbed truck that contained probably over 50 and less than 100 50 gallon drums containing unknown substances. We then pick axed them open and pitched them over the edge of a hill on the seemingly deserted part of the base. At the time, being just a young person I just did it. Now I'm concerned about the environmental damage caused by those chemicals and if my health could be affected by exposure to those chemicals.
Response
The Environmental Office at Fort Wainwright, in coordination with the Alaska Department of Environmental Conservation and the EPA, are continually working to characterize and remediate historic dumping sites on Fort Wainwright. Based on the information provided, I can't be certain if the site you described was covered in the public health assessment.
ATSDR discussed your concerns with the Fort Wainwright Environmental Office. We have been assured that if you contact them, they will be able to go through their records and maps with you. They will be able to identify if the area you describe has been investigated and if so, the results of the investigation. If the site has been investigated, they will likely know the types of chemicals that were disposed of and what your occupational exposure may have been. If the site has not yet been investigated they will be able to use your information to evaluate if future environmental investigations need to be considered.
