PUBLIC HEALTH ASSESSMENT
PADUCAH GASEOUS DIFFUSION PLANT (U.S. DOE)
PADUCAH, MCCRACKEN COUNTY, KENTUCKY
APPENDIX C: HEALTH GUIDELINES, COMPARISON VALUES, AND EXPOSURE FACTORS
When a hazardous substance is released to the environment, people are not always exposed to it. Exposure happens when people breathe, eat, drink, or make skin contact with a contaminant. People can also be exposed to radioactive contaminants by irradiation--if they get close to the radioactive material and if the contaminants are present at high concentrations.
Several factors determine the type and severity of health effects associated with exposure to contaminants. Such factors include exposure concentration, frequency and duration of exposure, route of exposure, and multiplicity of exposure (i.e., the combination of contaminants and routes). Once exposure takes place, individual characteristics--such as age, sex, nutritional status, genetics, lifestyle, and health status--influence how that person absorbs, distributes, metabolizes, and excretes the contaminant. These characteristics, together with the exposure factors discussed above and the specific toxicological effects of the substance, determine the health effects that may result.
ATSDR considers these physical and biological characteristics when developing health guidelines. Health guidelines provide a basis for evaluating exposures estimated from concentrations of contaminants in different environmental media (soil, air, water, and food) depending on the characteristics of the people who may be exposed and the length of exposure.
ATSDR reviews health and chemical information in documents called toxicological profiles. Each toxicological profile covers a particular substance; it summarizes toxicological and adverse health effects information about that substance and includes health guidelines such as ATSDR's minimal risk level (MRL), EPA's reference dose (RfD) and reference concentration (RfC), and EPA's cancer slope factor (CSF). ATSDR public health professionals use these guidelines to determine a person's potential for developing adverse non-cancer health effects and/or cancer from exposure to a hazardous substance. ATSDR does not have guidelines for exposure to radioactive materials. Instead, the agency uses existing regulatory values and national or international recommendations.
An MRL is an estimate of daily human exposure to a contaminant that is likely to be without an appreciable risk of adverse non-cancer health effects over a specified duration of exposure (acute, less than 15 days; intermediate, 15 to 364 days; chronic, 365 days or more). Oral MRLs are expressed in units of milligrams per kilogram per day (mg/kg/day); inhalation MRLs are expressed in micrograms per cubic meter (µg/m3). MRLs are not derived for dermal exposure.
RfDs and RfCs are estimates of daily human exposure, including exposure to sensitive subpopulations, that are likely to be without appreciable risk of adverse non-cancer health effects during a lifetime (70 years). These guidelines are derived from experimental data and lowest-observed-adverse-effect levels (or no-observed-adverse-effect levels), adjusted downward using uncertainty factors. The uncertainty factors are used to make the guidelines adequately protective of public health. RfDs and RfCs should not be viewed as strict scientific boundaries between what is toxic and what is nontoxic.
For cancer-causing substances, EPA established the CSF [1]. A CSF is used to determine the number of excess cancers expected from maximal exposure for a lifetime.
Comparison values are estimated contaminant concentrations that are unlikely to cause detectable adverse health outcomes when these concentrations occur in specific media. Comparison values are used to select site contaminants for further evaluation. They are based on health guidelines. Comparison values are calculated using conservative assumptions about daily intake rates by an individual of standard body weight. Because of the conservatism of the assumptions and safety factors, contaminant concentrations that exceed comparison values for an environmental medium do not necessarily indicate a health hazard.
For nonradioactive chemicals, ATSDR uses comparison values like environmental media evaluation guides (EMEGs), cancer risk evaluation guides (CREGs), reference dose (or concentration) media evaluation guides (RMEGs), and others. EMEGs, since they are derived from MRLs, apply only to specific durations of exposure. Also, they depend on the amount of a contaminant ingested or inhaled. Thus, EMEGs are determined separately for children and adults, and also separately for various durations of exposure. A CREG is an estimated concentration of a contaminant that would likely cause, at most, one excess cancer in a million people exposed over a lifetime. CREGs are calculated from CSFs. Reference dose (or concentration) media evaluation guides (RMEGs) are media guides based on EPA's RfDs and RfCs.
EPA's maximum contaminant levels (MCLs) are maximum contaminant concentrations of chemicals allowed in public drinking water systems. MCLs are regulatory standards set as close to health goals as feasible and are based on treatment technologies, costs, and other factors.
For radiological contaminants, ATSDR uses information on radiation exposure and its effects prepared by federal agencies, including EPA, DOE, and the US Nuclear Regulatory Commission. The agency also uses other publicly available data sources and recommendations on radiation dose limits. The National Council on Radiation Protection and Measurements (NCRP), the International Commission on Radiological Protection (ICRP), and the United Nations Scientific Committee on the Effects of Atomic Radiation are a few of the sources.
ATSDR uses standard or site specific intake rates for inhalation of air and ingestion of water, soil, and biota. Table C-1 presents the intake rates for groundwater, surface water, soil, and sediment that we used in estimating doses for PGDP. (The dose calculation equations, and our assumptions about exposure factors, are derived from the ATSDR Public Health Assessment Guidance Manual [2].) For screening purposes, ATSDR often uses the maximum contaminant concentration detected in a specific medium at a site to identify contaminants requiring specific exposure evaluations; using the maximum concentration results in a more protective evaluation. When unknown, the biological absorption of a substance within the human body is assumed to be 100%.
After estimating the potential exposure at a site, ATSDR identifies the site's "contaminants of concern" by comparing the exposures of interest with health guidelines, or contaminant concentrations with comparison values. As a general rule, if the guideline or value is exceeded, ATSDR evaluates exposure to determine whether it is of potential health concern. Sometimes additional medical and toxicological information may indicate that these exposures are not of health concern. In other instances, exposures below the guidelines or values could be of health concern because of interactive effects with other chemicals or because of the increased sensitivity of certain individuals. Thus additional analysis is necessary to determine whether health effects are likely to occur.
Exposure doses via ingestion are calculated on the basis of the following equation:
Dose (Ingestion) = (Chemical Conc. x IR x EF x ED) / (BW x AT)
where:
Chemical Conc. = concentration of each contaminant (in mg/g,µg/g, mg/L, or µg/L) IR = ingestion rate (in grams/day or liters/day) EF = exposure frequency in days per year ED = exposure duration in years BW = body weight in kilograms AT = averaging time in days
For soil and sediment doses, we take an additional step to determine exposure via dermal absorption, with the total dose being the sum of the ingestion dose and the dermal dose.
Dose (Dermal) = (Chemical Conc. x ABS x TSA x EF x ED) / (BW x AT)
where all factors are as above except:
ABS = a chemical-specific absorption or bioavailability factor (unitless) TSA = total soil adhered in milligrams (skin surface area x soil adherence value)
Once we have calculated the dose (in mg/kg/day) for a contaminant, we evaluate that contaminant's non-cancer and cancer health effects. For the former, we compare the dose with studies that have investigated the health effects of exposure to the contaminant. For the latter, we multiply the dose by the pathway-specific CSFs which are expressed in units of inverse dose--that is, (mg/kg/day)-1.
Excess Cancer Risk = Dose (mg/kg/day) x Cancer Slope Factor (mg/kg/day)-1
The excess cancer risk is the expected increase in cancer risk due to contaminant exposure. All of the uncertainties and health-protective exposure assumptions associated with the dose calcuations are included in the risk estimation, as well as the uncertainty in deriving the CSF. Excess cancer risks are described by the following categories [3]:
No increased risk less than 1 per 100,000 < 0.00001 No apparent increased risk 1 per 100,000 0.00001 Low increased risk 1 per 10,000 0.0001 Moderate increased risk 1 per 1,000 0.001 High increased risk 1 per 100 0.01 Very high increased risk more than 1 per 100 > 0.01
None of the excess cancer risk estimates necessarily indicate that exposure to carcinogenic contaminants will result in cancer in the exposed population.
References
Table C-1. Dose equations and factors used in calculating exposure doses at PGDP
| Dose Parameters | Groundwater | Surface Water | Soil | Sediment |
| Ingestion rate WKWMA workers Adults Children Pica children |
2 liters/day 2 liters/day 1 liter/day 1 liter/day |
0.5 liters/day 0.5 liters/day 0.5 liters/day None |
200 mg/day 50 mg/day 200 mg/day 2,000 mg/day |
100 mg/day None 100 mg/day None |
| Total soil adhered (for dermal contact) WKWMA workers Adults Children Pica children |
NA NA NA NA |
NA NA NA NA |
37,600 mg 9,400 mg 5,250 mg 3,000 mg |
37,600 mg 9,400 mg 5,250 mg 3,000 mg |
| Exposure frequency WKWMA workers Residents (adults and children) |
All groundwater exposures based on residential scenario | 12 days/year 12 days/year |
1.5 days/week 5.6 days/week |
0.75 days/week 12 days/year |
| Exposure area (location of stations used to determine 67th percentile concentration) |
All groundwater exposures based on specific well data | All surface-water stations outside of security fence | WKWMA workers: buffer zone stations
Residents: stations outside buffer zone |
All sediment stations outside security fence |
| Exposure duration | ~14 years (1974-1988) |
30 years (adult) 6 years (child) |
30 years (adult) 6 years (child) 3 years (pica child) |
30 years (adult) 6 years (child) |
| Body weight Adults Children Pica children |
70 kg 13 kg 10 kg |
|||
|
Averaging time
|
Adult: 30 years x 365 days/year Child: 6 years x 365 days/year Pica child: 3 years x 365 days/year 70 years x 365 days/year |
|||
| This table does not include information for the food and biota pathway; see the food and biota section of the public health assessment. | ||||
| Key: kg = kilograms; mg = milligrams; mg/day = milligrams per day; WKWMA = Western Kentucky Wildlife Management Area | ||||
APPENDIX D: ESTIMATION OF EXPOSURE DURATION FPR GROUNDWATER PATHWAY
Four residences near the northwest boundary of PGDP were exposed to trichloroethylene (TCE) and technetium 99, and possibly to lead, pentachlorophenol, and vinyl chloride via contaminated groundwater. The exposure occurred via ingestion of and dermal contact with groundwater, and inhalation of vapors from contaminated groundwater. Residents were provided with an alternate water source upon discovery of the contaminants in August 1988.
Very little groundwater monitoring took place before 1988, so monitoring data cannot be used to determine the duration of contaminant exposure. The rate of contaminant transport after 1988 has been used to estimate the annual rate of contaminant migration. The locations of the 100-microgram-per-liter TCE isocontours were qualitatively interpreted from monitoring data for 1988, 1991, and 1995. (This concentration was chosen not for health reasons but for better reliability in the data.) These contours were interpolated using maximum annual concentrations from residential and monitoring wells. The contouring procedure locates the line of equal concentration (100 µg/L) based on point values and the distances between adjacent values.
Figure D-1 shows isocontours for 1991 and 1995, which were drawn using ArcView overlaid on a site map. ArcView's map measurement tool was used to measure the plume progressions from 1988 to 1991 and from 1991 to 1995. These distance measurements divided by the number of years of plume progression (3 years and 4 years, respectively) yield a plume progression rate between 125 and 330 meters per year, depending on the time interval (1988-1991, 330 meters per year; 1988-1995, 207 meters per year; 1991-1995, 125 meters per year); see Table D-1.
Table D-1. Estimated plume migration rates based on plume locations for different time periods
| Time Period | Plume Progression | Annual Migration Rate |
| 1988-1991 (3 years) | 960 meters | 330 meters/year |
| 1988-1995 (7 years) | 1,450 meters | 207 meters/year |
| 1991-1995 (4 years) | 500 meters | 125 meters/year |
The largest uncertainty associated with estimating exposure duration is in interpreting the TCE isocontours. The 1988 contour is based on 21 annual data points (maximums at each well), which are irregularly distributed. Because data are limited, the resulting isocontour is a conservative estimate. The 1988 isocontour is approximately 1,200 meters (0.75 miles) downgradient of the residential well closest to the site boundary. The small number of data points used to interpret the plume location suggests that it had progressed at least that far but probably further. Also, the time between measured TCE concentrations was rounded to annual values. The data used to generate the contours are annual maximums, which occurred at approximately the same time each year.
Site personnel estimate plume migration at about 1 foot (30 centimeters) per day, which adds up to 110 meters per year. (This information came from a January 22, 1998, communication with Brad Montgomery of Bechtel Jacobs Company and a February 2, 1998, communication with Ross Miller of Geo Consultants, LLC.) This estimate is based on extensive flow modeling and the measured migration rate of tracers injected into the Regional Gravel Aquifer (RGA). Given the abovementioned uncertainty about the 1988 isocontour's location, we used a contaminant migration rate of 110 to 125 meters per year to evaluate the duration of contaminant exposure.
Table D-2 indicates the distance that the contaminant plumes have migrated beyond the affected residential wells. The 1991 and 1995 isocontours show that the plume has moved 2,200 meters (as of 1991) and 2,640 meters (as of 1995) downgradient of the wells. Dividing these distances by the annual migration rate of 125 meters per year provides an estimate of the total duration of plume migration. Subtracting the years of post-1988 migration from this total provides an estimate of the pre-1988 exposure duration. Using the 1991 and 1995 plume locations, the estimated pre-1988 exposure durations are 14.6 and 14.1 years (Table D-2). This estimate is for those wells closest to the site boundary (RW-002 and RW-113). Exposures for wells further downgradient would be of shorter duration. Also, this exposure duration is for TCE concentrations greater than 100 µg/L. Exposures at lower concentrations probably had a longer duration.
Table D-2. Estimated exposure durations, based on 125 meters per year migration rate and distance of plume migration downgradient of residential wells
| Plume Distance Beyond Residential Wells | Years Migration Past 1988 Exposure | Estimated Duration of Exposure (using 125 meters/year migration rate) |
| 1988: 1,200 meters | 0 years | 9.6 years |
| 1991: 2,200 meters | 3 years | 17.6 years - 3 years = 14.6 years |
| 1995: 2,640 meters | 7 years | 21.1 years - 7 years = 14.1 years |
TCE concentrations in the affected wells probably varied considerably over the exposure period. While there are no data for this period, well concentrations in the years after 1988 indicate significant variation in concentrations. (See Figure 4 in the main body of this report.) These variations are probably due to the changes of seasonal water levels in the Ohio River (i.e., river stages). Changes in the river stage directly affect both flow rate and direction in the RGA. In the case of well RW-017, high TCE concentrations correspond with times of lower river stages. Although we used maximum annual concentrations to calculate exposure doses (Table 5 in the main body of this report), ingested concentrations probably varied by a factor of two.
Figure D-1. TCE Isocontours (1991-95) and Contaminated Off-Site Well Locations
APPENDIX E: EXPOSURE TO AIRBORNE RADIONUCLIDES
Exposure doses to airborne radionuclides were estimated using the Clean Air Act Assessment Package--1988 (CAP88), a system developed by EPA [1,2]. CAP88 uses a modified Gaussian plume equation to estimate the average dispersion of radionuclides released from up to six sources. The sources can be either elevated stacks, such as a smokestack, or uniform area sources, such as a pile of uranium mill tailings. Plume rise can be calculated assuming either a momentum or buoyancy-driven plume. Assessments are done for a circular grid of distances and directions with a radius of 80 kilometers (50 miles) around the facility.
The program computes radionuclide concentrations in air, rates of deposition on ground surfaces, concentrations in food and intake rates to people ingesting food produced in the assessment area. Estimates of radionuclide concentrations in produce, leafy vegetables, milk, and meat consumed by humans are made by coupling the output of atmospheric transport models with the U.S. Nuclear Regulatory Commission Regulatory Guide 1.109 terrestrial food chain models.
Dose and risk estimates from CAP88 are applicable only to low-level chronic exposures, since the health effects and dosimetric data are based on low-level radionuclide intakes. The population estimates used in this evaluation are the 1980 Census data provided with the CAP88 model. In addition to population estimates, the model requires information on radionuclide emission rates, meteorological data, and agricultural data on consumption of locally grown food and dairy products. Radionuclide emission data were obtained from the annual site environmental monitoring reports.
The two meteorological data sets that were used in the evaluations are provided with the CAPP88 model. The 1950s emission years used 1960-1964 meteorological data; the 1996 emission year evaluation used 1989-1993 meteorological data. Agricultural input data, stack parameters, and source partitioning were based on information provided in the 1996 National Emission Standards for Hazardous Air Pollutants (NESHAP) report [3]. Four sources account for most PGDP radionuclide emissions: the C-310 stack, C-400 combined sources, the seal and wet air exhausts, and the C-710 laboratory. CAP88 places all sources at the center of the facility with respect to the surrounding population and varies only the height of the release. This evaluation used a zero plume rise factor based on emission temperature and velocity information in the NESHAP report.
The radionuclides evaluated include technetium 99, uranium 234, uranium 235, and uranium 238. The results reported in Table E-1 are for 1956 through 1959 (the years with the largest releases) and for 1996 (a recent year for which there is complete information). The uranium isotope releases were partitioned between the sources in the following proportions for all years:
Although these proportions may have changed with process and control operations, any variations in the sources had minimal effect on the estimated dispersion concentrations, because CAP88 locates all emissions at the same geographic point and because a zero plume rise was used.
Table E-1. Annual radionuclide emissions for selected isotopes and years [4,5,6,7]
| Year | Technetium 99 in curies (gigabecquerels) |
Uranium 234 in curies (gigabecquerels) |
Uranium 235 in curies (gigabecquerels) |
Uranium 238 in curies (gigabecquerels) |
| 1956 | 2.6 (96.2) | 1.62 (59.94) | 0.08 (2.96) | 3.50 (129.5) |
| 1957 | 4.8 (177.6) | 1.10 (40.7) | 0.05 (1.85) | 1.20 (44.4) |
| 1958 | 6.3 (233.1) | 1.09 (40.33) | 0.05 (1.85) | 1.16 (42.92) |
| 1959 | 5.1 (188.7) | 0.93 (34.41) | 0.04 (1.48) | 1.10 (40.7) |
| 1996 | 0.04 (1.48) | 0.003 (0.111) | 0.0001 (0.004) | 0.001 (0.037 |
In addition to the chronic or long-term process releases, accidental releases of UF6 have occurred throughout the operating history of the PGDP facility [8,9]. The largest reported accidental release occurred in 1960, when a cylinder ruptured releasing about 11,000 pounds (approximately 5,000 kilograms) of UF6 . This accident occurred in Building C-333 on November 17, 1960, at about 4:00 a.m. Another accidental release occurred during a fire at Building C-337 in December 1962. About 5,062 pounds (2,278 kilograms) of UF6 were released during the fire.
Acute airborne uranium hexafluoride (UF6) concentrations near PGDP from the 1960 and 1962 accidents were estimated using the RASCAL 3.0 air dispersion and dose model [10] and weather observations from the Paducah/Barkley Airport [11]. The RASCAL model (beta test version) provides a general assessment of potential uranium air concentrations following accidental releases. Due to the confluence of the water vapor from PGDP cooling towers with any airborne releases, atmospheric humidity is assumed to be similar to conditions of light precipitation.
Our data on weather conditions at the time of the 1960 release indicate a stable to very stable atmosphere (stability class F), very low wind speed from the northwest, and a temperature of 39oF (dry bulb) [11]. Under these release conditions and according to our modeling of this accident, an estimated uranium inhaled radiation dose of 1.5 rem (0.015 sieverts) and an estimated uranium inhaled chemical dose of 20 milligrams (mg) could have been received by the maximally exposed resident southeast of the site. The U.S. Nuclear Regulatory Commission's action level for intake of soluble uranium is 10 mg. (At this action level, residents may be instructed to evacuate or to stay indoors with windows closed.) A report assessing PGDP accidents [10] indicates that a 5-mg uranium dose can produce detectable, non-permanent kidney damage. The 1960 cylinder rupture could have resulted in inhaled exposure doses of 5 mg to 20 mg to people who lived approximately 2.5 miles (4 kilometers) from the release site. That includes off-site areas to the southeast of the site.
According to accident records, this release occurred on November 17, 1960, at approximately 4:00 a.m. At that time of day and year, it is unlikely that nearby residents would be outside, where exposure to the maximum concentrations would occur. Air temperatures were in the 30s, so windows and doors would have been shut--very little exposure to residents inside their houses probably occurred. Additionally, this exposure scenario assumes that 62% of the UF6 cylinder content was vented from the building over a 1-hour period and became airborne. Notes from accident summaries suggest that a considerable portion of the UF6 remained in the liquid phase and was recovered [9].
Estimated uranium air concentrations and doses from the 1962 fire are much lower than from the 1960 cylinder accident. The explosion and fire that caused this release resulted in much greater atmospheric dispersion and much lower air concentrations and doses. Off-site uranium air concentrations from this accident probably did not present a health hazard to the surrounding community.
In addition to the documented 1960 and 1962 accidents, the community had concerns about two other potential incidents: a 3-day UF6 release on March 15 through 17, 1970; and a large accidental release sometime in 1969 or 1970. A Union Carbide memorandum contained reference to a 3-day UF6 (March 15-17, 1970) that was detected via on-site air monitoring inside the building [8]. This memorandum also indicated that the average gross alpha air monitoring results for the perimeter east location from October 1969 to May of 1970 were higher than normal. However, the individual weekly air monitoring results indicated that this average was elevated for a different time period than March 15-17, 1970 [12]. Also, the plant's original report for this incident indicated that a total of 15 grams of uranium was released inside the building and eventually released through the building ventilation system. This amount of uranium would not have an adverse impact off site. For the second concern (an accident that occurred in 1969 or 1970 when houses to the southeast turned black and trees died), the site accident records for the 1969/70 time frame do not report any events capable of producing significant off-site uranium or hydrogen fluoride concentrations.
However, an extensive review of the weekly air monitoring data indicates that there were several periods of elevated gross alpha and gross beta (presumably, uranium and technetium 99) concentrations at perimeter air monitors during the 1969 and 1970 timeframe [12]. There is some indication that the site investigated elevated gross beta levels to the north of the plant during this time, but there was no explanation of the cause. Due to the limited information available on these specific events, ATSDR cannot evaluate potential exposure doses off site. However, monitoring data do indicate that some type of release event(s) occurred that are not reflected in the accident reports reviewed.
At this time, it is not possible to determine if nearby residents were actually exposed to hazardous concentrations of uranium from any of these accidental releases. However, this analysis does estimate that potentially hazardous releases have occurred and that rupture of a UF6 cylinder represents potentially hazardous conditions for residents living adjacent to PGDP. In addition, the air dispersion models suggest that significant concentrations of uranium may have been deposited in off-site areas. Currently, we have no reports of health effects related to these accidents; however, if data become available suggesting that health effects did occur, we will re-evaluate the need for followup activities.
References
APPENDIX F: EXPOSURE TO AIRBORNE HYDROGEN FLOURIDE
During the uranium enrichment processes at PGDP, uranium hexafluoride (UF6) is released into the air. The UF6 reacts rapidly with water in the air to form particulate uranium and fluorides, and also hydrogen fluoride gas (HF) [1]. HF is the most abundant form of atmospheric fluoride and reacts with atmospheric water to form hydrofluoric acid aerosols. Airborne particulate fluorides have low solubility and are removed from the atmosphere through dry and wet deposition.
Releases of UF6 (with atmospheric conversion to HF) occurred both as long-term releases due to process operations and as short-term releases due to accidents. Long-term (chronic) exposure to HF is evaluated based on correlation of annual UF6 releases with measured site perimeter HF concentrations. Short-term (acute) HF exposures are evaluated using accident records and air dispersion modeling.
Estimated uranium releases and ambient air monitoring results have been reported consistently throughout PGDP's operational history; fluoride releases and HF ambient air concentrations have not. Evaluation of potential HF exposures to nearby residents presents several problems: no reporting of HF release quantities or ambient air monitoring during the period of highest potential fluoride and HF emissions (1956), changes in sampling locations, and changes in the data reported (e.g., annual medians vs. means). Consequently, evaluation of chronic HF exposures during the period of highest potential emissions requires estimation or modeling of HF emissions from periods of consistent data reporting.
Uranium emissions are a good proxy for prediction of chronic or long-term HF ambient air concentrations. Ambient airborne HF concentrations were measured at several locations for the years 1961 to 1970. Comparison of mean ambient airborne HF concentrations from these locations with uranium emission estimates for the same years provides a correlation coefficient of 0.8863, which indicates a strong positive relationship between uranium emissions and measured HF concentrations at the perimeter north monitoring site (HF concentrations increase proportionately with increases in uranium emissions). That relationship is plotted in Figure F-1.
The strong correlation of uranium emissions and HF concentrations at the perimeter north station in the years where both data sets are available allows for the prediction of HF concentrations from uranium emission data in the years for which no HF monitoring data are available. These HF concentrations are predicted using the linear regression forecasting function in the computer program EXCEL (version 7.0a). Figure F-2 shows the relationships between uranium emissions (in curies per year), estimated and monitored HF concentrations at the perimeter north and one mile north stations, and measured fluoride concentrations in grass samples near the perimeter north station.
The perimeter north station consistently had the highest concentrations of both particulate uranium and HF. The perimeter north station was closer to the fluoride processing facility than other stations [2], and was downwind of the processing facility with respect to the prevailing south-southwest winds [3]. Therefore, it was assumed that this station would have had the highest concentrations of HF during the year of highest release.
All of the measured parameters show a strong relationship to uranium emissions and to the estimated HF concentrations. Figure F-2 shows measured and estimated HF ambient concentrations in relation to the Kentucky ambient air standard for average annual HF exposure (500 parts per billion, or ppb) [4] and the ATSDR provisional long-term guidance value of 12 ppb. None of the measured or estimated HF concentrations exceed the Kentucky ambient air standard.
ATSDR has established a provisional guidance value of 10 micrograms per cubic meter (12 ppb) for annual average air concentrations of HF [5,6]. HF concentrations below 12 ppb (annual average value) are not likely to cause adverse health effects. This guidance value is more than 100 times lower than an exposure concentration that caused mild irritation to the eyes and noses of human volunteers exposed for 10 days [1]. None of the measured or estimated HF concentrations at the one north sampling station exceeded this guidance value (Figure F-2).
Some of the estimated HF concentrations at the perimeter north station did exceed the ATSDR guidance value; the maximum value was 28 ppb (Figure F-2). The maximum annual HF emission occurred in 1956, which is the period of maximum uranium emissions. Because HF concentrations at the perimeter north station are consistently higher than at other locations, this station represents a worst-case exposure scenario. It is important to point out that no off-site residents live at the perimeter security fence. The nearest houses are closer to the one mile north and east stations than to the perimeter stations. Consequently, the concentrations at the nearest house would have been closer to the concentrations estimated by the one north station (Figure F-2) than to the concentrations at the perimeter north station. The estimated annual average HF concentrations at these points of exposure are below levels of health concern.
There is some uncertainty associated with deriving HF concentrations from uranium emissions. One measure of this uncertainty is the standard error, which is represented by error bars on the predicted HF concentrations in Figure F-2. The error bars, which show the predicted maximum and minimum HF values, do not significantly change the predicted HF concentrations with respect to the ATSDR and Kentucky health guidance values. Note that the largest standard errors occur between 1965 and 1968, the period with the highest variability and lowest uranium emissions.
In addition to the chronic or long-term process releases, accidental releases of UF6 and HF (estimated from reported UF6 releases) have occurred throughout the operating history of the PGDP facility [7,8]. The largest reported accidental release occurred in 1960, when a cylinder ruptured releasing about 11,000 pounds (approximately 5,000 kilograms) of UF6. This accident occurred in Building C-333 on November 17, 1960, at about 4:00 a.m. Another accidental UF6 release occurred during a fire at Building C-337 in December 1962. About 5,062 pounds (2,278 kilograms) of UF6 were released during this fire. Many other smaller releases have occurred, but these were at least an order of magnitude smaller than the 1960 release and less than 30% of the size of the 1962 release.
Airborne UF6 and HF concentrations surrounding PGDP from the 1960 and 1962 accidents were estimated using the RASCAL 3.0 air dispersion and dose model [9] and weather observations from the Paducah/Barkley Airport. The RASCAL model (beta test version) is used to provide a general assessment of potential HF and uranium air concentrations following accidental releases. Due to the confluence of the water vapor from PGDP cooling towers with any airborne releases, atmospheric humidity is assumed to be similar to conditions of light precipitation.
Our data on weather conditions at the time of the 1960 release indicate a stable to very stable atmosphere (stability class F), very low wind speed from the northwest, and a temperature of 39ºF (dry bulb) [10]. Under these release conditions, short-term hazardous HF concentrations (6 parts per million, or ppm; 15-minute Short Term Exposure Limit) could have extended more than 1 kilometer (0.6 miles) from the release site (Building C-333) toward the southeast. This means that the estimated HF concentrations could have been at hazardous levels immediately off site. Estimated concentrations of more than 30 ppm, which is considered immediately dangerous to life/health, extended more than 500 meters (1,640 feet) from Building C-333 and would not have reached the off-site community. Table F-1 summarizes the air dispersion analysis.
According to accident records, the 1960 release occurred on November 17, 1960, at approximately 4:00 a.m. At that time of day and year, it is unlikely that nearby residents would be outside, where exposure to the maximum concentrations would occur. Air temperatures were in the 30s, so windows and doors would be shut--very little exposure to residents inside their houses probably occurred.
The explosion and fire that caused the 1962 release resulted in much greater atmospheric dispersion and much lower air concentrations and doses. This release was modeled using a fire scenario in RASCAL 3.0, which did not analyze HF dispersion. However, HF in a fire is atmospherically unstable and very unlikely to undergo significant atmospheric dispersion. Off-site HF air concentrations from this accident probably did not present a health hazard to the surrounding community.
In addition to the documented 1960 and 1962 accidents, there were community concerns about two other potential incidents: a 3-day UF6 release on March 15 through 17, 1970; and a large accidental release sometime in 1969 or 1970. A Union Carbide memorandum contained reference to a 3-day UF6 (March 15-17, 1970) that was detected via on-site air monitoring inside one of the buildings [7]. This memorandum also indicated that the average gross alpha air monitoring results for the perimeter east location for the period from October 1969 to May of 1970 were higher than normal. However, the individual weekly gross alpha air monitoring results indicated that this average was elevated for a different time period than March 15-17, 1970 [11]. The airborne fluoride results from the perimeter east location for March 1970 were not elevated. Also, the plant's original report for this incident indicated that a total of 15 grams of uranium was released inside the building and eventually released through the building ventilation system. This amount of uranium hexafluoride would not have an adverse impact off site. For the second concern (an accident that occurred in 1969 or 1970 when houses to the southeast turned black and trees died), the site accident records for the 1969/70 time frame do not report any events capable of producing significant off-site uranium or hydrogen fluoride concentrations.
However, an extensive review of the weekly air monitoring data indicate that there were several periods of elevated fluoride concentrations at perimeter air monitors during the 1969 and 1970 timeframe [11]. There is some indication that the site investigated elevated hydrogen fluoride levels to the east of the plant later in 1970, but there was no explanation of the cause. Due to the limited information available on these specific events, ATSDR cannot evaluate potential exposure doses off site. However, monitoring data do indicate that some type of release event(s) occurred that are not reflected in the accident reports reviewed.
At this time, it is not possible to determine if nearby residents were actually exposed to hazardous concentrations of uranium from any of these accidental releases. However, this analysis does indicate that potentially hazardous releases have occurred and that rupture of a UF6 cylinder represents potentially hazardous conditions for residents living adjacent to PGDP. Currently, we have no reports of health effects related to the reported accidents; however, if data become available suggesting that health effects did occur, we will re-evaluate the need for followup activities.
Table F-1. Summary of RASCAL 3.0 model for assessing accidental releases of HF from November 17, 1960, UF6 cylinder rupture
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References
APPENDIX G: AIRBORNE TRICHLOROETHYLENE DISTRIBUTION AND POTENTIAL OFF-SITE EXPOSURE
Past operations at PGDP involved large quantities of trichloroethylene (TCE) as an organic solvent and degreaser. Although significant amounts of TCE were released into the groundwater system, most TCE from operational processes volatilized into the atmosphere [1]. To determine if those airborne releases present a potential for inhalation exposure to nearby residents, we conducted a TCE air dispersion analysis using the Industrial Source Complex (ISC3) model [2].
The ISC3 model uses meteorological data to generate air concentration averages, for periods from 1 hour to 1 year, for any location surrounding an air emission source. ATSDR's analysis used 1989 meteorological data from the nearby Paducah Municipal Airport (Barkley Airfield). According to data available to us, the largest annual release of TCE to the atmosphere (62,826 kilograms, or 138,845 pounds) occurred in 1986 [3]. The dispersion from this release was modeled as a single source from Building C-400--a vent 5 meters (16 feet) off the ground with a diameter of 25 centimeters (10 inches)-- and the annual release proportioned over the entire year.
For this analysis, we assumed conservative dispersion with no chemical degradation or photochemical breakdown of TCE. (Typically, TCE is estimated to degrade in the atmosphere with a chemical half-life of 3 to 7 days [4].) Under these conservative assumptions, the maximum airborne TCE concentration is 112 micrograms per cubic meter (µg/m3) for a 1-hour averaging period, and 3 µg/m3 for a 1-year averaging period, at a location 1 kilometer (0.6 miles) north of Building C-400 (that is, off site). Some animal studies have shown carcinogenic effects from TCE; however, ATSDR and EPA are re-evaluating TCE's carcinogenic effects on humans. Until TCE's carcinogenicity for humans is determined, minimal risk levels (MRLs) for non-cancerous effects are used to screen for contaminants of concern. The MRL for TCE is 10,920 µg/m3 for acute exposures (1 to 14 days) and 546 µg/m3 for intermediate exposures (15 to 365 days) [4,5]. The estimated TCE air concentrations during the highest TCE release year are two orders of magnitude lower than the MRLs and below levels of health concern. Consequently, we did not choose TCE as a contaminant of concern for airborne releases at PGDP.
References
APPENDIX H: AIRBORNE HEXAVALENT CHROMIUM DISTRIBUTION FROM THE PGDP WATER COOLING TOWERS AND POTENTIAL OFF-SITE EXPOSURE
Isotopic diffusion operations at PGDP generate excess heat, which is released to the environment through four cooling systems. In these systems, heat exchangers transfer heat to cooling waters, which in turn release the heat to the atmosphere through 14 water cooling towers (located in four discrete areas). Until 1993, a chromium solution was added to the cooling waters to prevent corrosion [1]. This caused hexavalent chromium to be released to the atmosphere at the water cooling towers.
Although annual chromium emissions have been calculated based on the quantities of chromium compounds added to the cooling system, the airborne chromium concentration has never been measured at on-site or off-site locations. DOE has measured and modeled chromium deposition in surrounding soils and plants, and found that chromium concentrations are at background levels for locations more than 1,500 meters from the cooling towers [2,3]. However, the security fence to the east and north of the easternmost cooling systems is less than 500 meters from the towers, and the closest residence is about 1,000 meters from the towers.
Because inhalation of hexavalent chromium can be toxic, we estimated potential exposures to airborne concentrations of chromium using the ISC3 air dispersion model [4]. This model uses site-specific meteorological data (in this case, from Barkley/Paducah Municipal Airport) and a Gaussian air dispersion equation to estimate contaminant concentrations.
The 1992 chromium release of 2,015 kilograms per year (0.064 grams per second, or g/sec) was
used as the emission rate for the dispersion calculations [5]. The 1992 chromium release was the
highest annual emission on record, and thus represents the most conservative source term for
evaluation of public health affects. To characterize local weather, we used the most recent complete
meteorological data set (a 1990 hourly data set) from the EPA SCRAM Web site
(http://www.epa.gov/ttn/scram/ ) for the Barkley/Paducah Municipal Airport weather station.
The chromium emissions were modeled as four sources, based on information from the study of cooling tower drift at PGDP [2]. Relative locations of the cooling towers were derived from Figure 1 of that study. Chromium concentrations at breathing height were estimated for a 5,000-meter polar grid, with potential receptors located at 500-meter intervals along 16 transects (every 22.5 degrees).
The source-specific release rates and source dimensions were based on a total annual emission rate of 0.064 g/sec, allocated between the four sources [2]. The four sources were modeled as volume sources, with release heights of 8 meters and lateral dimensions of 25 meters (towers 1 and 2) and 75 meters (towers 3 and 4). Release rates were estimated as:
The model used regulatory default options and dry settling/deposition to estimate all chromium concentrations [6]. Chromium particles in the cooling tower drift have diameters of 5 to 50 microns. However, these particles are contained within water droplets that have diameters of 100 to 1,300 microns [3]. A particle size distribution of 100 to 999 microns (with a 700-micron mean) was used in calculating settling velocities.
Maximum chromium concentrations were calculated for each potential receptor for 1-hour, 8-hour, 24-hour, and 1-year averaging periods. Because wind directions and speeds change so drastically over a year, these maximum concentrations represent the highest estimated concentrations for each time period for each location. Averages for 1-hour periods are significantly higher than the 8-hour, 24-hour, or 1-year averages.
The results of this modeling indicate that dispersed hexavalent chromium air concentrations on site and off site were lower than the health guidelines for intermediate and chronic exposures [7]. The intermediate minimal risk level (MRL) for inhalation of particulate hexavalent chromium is 0.5 micrograms per cubic meter (µg/m3). The intermediate and chronic MRL for inhalation of dissolved hexavalent chromium as an aerosol is 0.1 µg/m3. The highest estimated 1-hour, 8-hour, 24-hour, and 1-year average air concentrations were on site, between the four cooling towers; they are listed in Table H-1. The maximum estimated off-site concentrations were about 500 meters north-northeast of the cooling towers, outside the security fence, and are listed below. The closest residence is approximately 1,000 meters east of the easternmost cooling system, and the estimated maximum concentrations for this location are listed below. The results of this air dispersion model are in agreement with chromium distribution studies that found no air-dispersed chromium in soil or vegetation samples beyond 1,500 meters of the cooling towers [2,3].
Table H-1. Maximum estimated airborne hexavalent chromium concentrations
| Exposure Time | Maximum On Site | Maximum Off Site | Maximum at Closest Residence |
| 1-hour maximum | 0.0215 µg/m3 | 0.005 µg/m3 | 0.0011 µg/m3 |
| 8-hour maximum | 0.0108 µg/m3 | 0.0025 µg/m3 | 0.0005 µg/m3 |
| 24-hour maximum | 0.0067 µg/m3 | 0.0007 µg/m3 | 0.0005 µg/m3 |
| 1-year maximum | 0.0009 µg/m3 | 0.0003 µg/m3 | 0.0004 µg/m3 |
The results of this modeling study, which uses conservative assumptions for settling, dispersion, and emission rates, did not find any areas where exposure to airborne chromium exceeded health guidelines. The distribution of these airborne concentrations is supported by measurements of deposited chromium in soil and vegetation samples. Therefore, airborne hexavalent chromium was not selected as a contaminant of concern at PGDP.
References
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