Document N2DVgEq8q2mdvLqomZOezxedp

PROPOSED IDENTIFICATION OF PERCHLOROETHYLENE AS A TOXIC AIR CONTAMINANT PART A REPORT STATE OF CALIFORNIA AM RESOURCES BOARD STATIONARY SOURCE DIVISION DECEMBER 1989 SL 038788 DRAM PRELIMINARY DRAFT PROPOSED IDENTIFICATION OF PERCHLOROETHYLENE AS A TOXIC AIR CONTAMINANT TECHNICAL SUPPORT DOCUMENT REPORT TO THE AIR RESOURCES BOARD ON PERCHLOROETHYLENE PART A PUBLIC EXPOSURE TO, SOURCES, AND EMISSIONS OF PERCHLOROETHYLENE IN CALIFORNIA Project Coordinator Robert V. Rood Contributors Richard D. Bode Dora Chang Richard Corey Steve P. Hui Mimi Jones Marcella Nystrom Reviewed by: Joan E. Denton, Ph.D., Manager Substance Evaluation Section Donald J. Ames, Assistant Chief Stationary Source Division Peter D. Venturini, Chief Stationary Source Division December 1989 SL 038789 DRAFT DRAFT REPORT TO THE AIR RESOURCES BOARD ON PERCHLOROETHYLENE Part A - Public Exposure To, Sources and Emissions of Atmospheric Perchloroethylene in California TABLE OF CONTENTS Page LIST OF TABLES ........................................................................................................ iii LIST OF FIGURES ...................................................................................................... iii I. INTRODUCTION .............................................................................................. A-l II. PHYSICAL PROPERTIES OF PERCHLOROETHYLENE ......................................... A-2 III. PRODUCTION, USES, AND EMISSIONS OF PERCHLOROETHYLENE ................. A-4 A. PRODUCTION OF PERCHLOROETHYLENE .................................................. A-4 B. USE AND EMISSIONS OF PERCHLOROETHYLENE ..................................... A-4 C. EMISSIONS TRENDS ............................................................................... A-13 D. POTENTIAL SOURCES OF INDOOR PERCHLOROETHYLENE ....................... A-14 IV. EXPOSURE TO PERCHLOROETHYLENE .............................................................. A-21 A. AMBIENT MONITORING FOR PERCHLOROETHYLENE ................................. A-21 B. AMBIENT CONCENTRATIONS OF PERCHLOROETHYLENE ........................... A-21 C. EXPOSURE TO PERCHLOROETHYLENE NEAR EMISSION SOURCES ............ A-33 D. INOOOR AIR EXPOSURE TO PERCHLOROETHYLENE ................................. A-36 E. OTHER ROUTES OF PERCHLOROETHYLENE EXPOSURE ............................. A-41 F. ESTIMATES OF TOTAL EXPOSURE FROM INDOOR AIR ........................... A-42 AND OTHER ROUTES V. ATMOSPHERIC PERSISTENCE AND FATE OF PERCHLOROETHYLENE ............... A-46 A. ATMOSPHERIC PERSISTENCE OF PERCHLOROETHYLENE ......................... A-46 B. FATE OF PERCHLOROETHYLENE IN THE ATMOSPHERE ........................... A-50 APPENDICES APPENDIX A -- METHODS FOR ESTIMATING USAGE AND EMISSIONS OF PERCHLOROETHYLENE IN CALIFORNIA APPENDIX B -- STANDARD METHODS OF ANALYSIS FOR TETRACHLOROETHYLENE SL 038790 ii DRAFT LIST OF TABLES AND FIGURES TABLES me II- 1 Physical Properties of Perchloroethylene .................................... A-2 III- 1 Sources, Usage and Emissions of Perchloroethylene .................. A-7 in California: 1987 IV- 1 Months Where At Least One Sample Was Collected and .............. Analyzed for Perchloroethylene: July 1988 through June 1989 A-23 IV-2 Minimum, Maximum, Median and Means of Perchloroethylene ........ A-26 Samples Collected during July 1988 through June 1989 IV-3 Summary of Intra-Basin Site Differences Based on ..................... A-28 Average Monthly Rankings and Monthly Means IV-4 Perchloroethylene Peak-to-Mean Ratios and ...................................... A-30 Coefficients of Variations (C.V.) IV-5 Mean Perchloroethylene Exposure Estimates: .................................... A-31 July 1988 through June 1989 IV-6 Lower Bound, Mean and Upper Bound Perchloroethylene ................... A-32 Concentrations for July 1988 through June 1989 IV-7 Overnight Indoor and Outdoor Air Levels of .................................... A-37 Perchloroethylene in California (6:00 pm to 6:00 am) IV-8 24-hour Concentrations of Perchloroethylene in ............................. A-39 Indoor and Outdoor Microenvironments IV-9 Estimated Doses of Perchloroethylene from Exposure ..................... A-43 through Different Media FISURES 1-1 Structure of Perchloroethylene ......................................................... III- l U.S. Production, Imports, Exports, and Use of ............................ Perchloroethylene IV- 1 Air Resources Board Toxics Network Monitoring Sites ................. IV-2 Mean Annual Perchloroethylene Concentrations ................................ Plotted using Extended Box Plots IV-3 Estimated Mean Annual Perchloroethylene Exposure ........................ IV- 4 Estimated Cumulative Perchloroethylene Exposure .......................... V- l Fate of Perchloroethylene ..................................................................... iii A-l A-5 A-22 A-2B A-34 A-35 A-52 SL 038791 INTRODUCTION Perchloroethylene is one of the family of chemicals known as chlorinated alkenes; chlorinated aliphatic hydrocarbon compounds containing a double bond. It is known by a variety of synonyms which include perchloroethene, tetrachloroethene, 1,1,2,2-tetrachloroethene, tetrachloroethylene, PERC, PCE, ethylene tetrachloride and Perclene (CAS No. 127-18-4). Perchloroethylene has the chemical formula CgCl^ and the chemical structure is shown in Figure 1-1. Figure 1-1 Structure of Perchloroethylene c\ /Cl c --c /\ Cl Cl Perchloroethylene has a wide number of uses in the industrial, governmental, and consumer sectors of the economy. It is used in California in a variety of products and processes, including dry cleaning operations, degreasing operations, adhesives, paints and coatings, aerosols, and specialty chemical production. Perchloroethylene is also used in California in printing inks, silicones, rug shampoos, and as a general solvent in laboratories. Approximately 19,000 tons of perchloroethylene were used in California in 1987. This resulted in approximately 18,000 tons being emitted into the atmosphere. This report is an evaluation of perchloroethylene uses, emissions, ambient and indoor concentrations, statewide population exposure, and atmospheric persistence and fate. The Air Resources Board (ARB) will consider the findings of this report together with the health effects findings of the Department of Health Services (DHS) to determine if perchloroethylene should be identified as a toxic air contaminant. SL 038792 II. PHYSICAL PROPERTIES OF PERCHLOROETHYLENE Perchloroethylene is a non-flammable, colorless, dense liquid at room temperature with an ethereal odor similar to chloroform. It is relatively insoluble in water (150 mg/L) but is miscible in alcohol, ether, chloroform, and benzene. Perchloroethylene is not known to contribute to either global warming or to the depletion of the stratospheric ozone layer. Some of the other physical properties of perchloroethylene are shown in Table II-l below. Table II-l Physical Properties of Perchloroethylene Property Value Reference Boiling Point (760 mm Hg) 3 Conversion (ppb to ug/m ) (ug/m to ppb) 121 C 3 1 ppb * 6*78 ug/m 1 ug/ m^ = 0.15 ppb Merck Index, 1983 Density, liquid (specific gravity) at 15C 1.6311 gm/cnf* at 20C 1.6230 gm/cm^ Merck Index, 1983 Merck Index, 1983 Molecular weight 165.83 Merck Index, 1983 Partition coefficient 1-octanol:water (25C) undecane:water (25C) blood:water (99F) 339-871:1, (409:1) 2700:1 - 31:1 U.S. EPA, 1985; Barbari and King, 1982 Barbari and King, 1982 U.S. EPA, 1985 Solubility, water (20C) 150 mg/L U.S. EPA, 1985 Vapor pressure (20C) (25C) 14 torr 19 torr U.S. EPA, 1985 U.S. EPA, 1985 A-2 SL 038793 References for Chapter TI Barbari, T. A.; C. J. King, 1982. "Equilibrium Distribution Coefficients for Extraction of Chlorinated Hydrocarbons and Aromatics from Water into Undecane." Environmental Science & Technology, 16(9):624. Merck Index, 1983, 10th Edition, Merck and Co., Inc., Rahway, New Jersey. U.S. EPA, 1985. "Health Assessment Document for Tetrachloroethylene - Final Report." EPA-600/8-82/005F, July 1985. 038794 A-3 DRAFT III. PRODUCTION. USES. AND EMISSIONS .OF-PERCHLOROETHYLENE Perchloroethylene is one of the most widely used chlorinated hydrocarbon solvents. It is used in a large variety of industrial, governmental and consumer activities. As a result of the extensive use of perchloroethylene as well as the manner in which it is used, substantial amounts of perchloroethylene are released to the atmosphere in California each year. A. PRODUCTION OF PERCHLOROETHYLENE Perchloroethylene is produced in the U.S. by five companies at eight facilities (U.S. EPA, 1985b). In 1985 U.S. production of perchloroethylene was 339,000 tons (C&E News, 1987). National production of perchloroethylene in 1985 increased 18 percent from 1984 production and is at approximately the 1975 production level (C&E News, 1987). The U.S. demand for perchloroethylene is forecasted to remain nearly the same through 1990 (CMR, 1986). However, since U.S. production of perchloroethylene depends on imports and exports as well as domestic demand, future production trends are not known. Figure III-1 indicates both production and use of perchloroethylene rose sharply from 1983 through 1985. In addition, Figure III-1 indicates that imports of perchloroethylene rose slightly, while exports dropped slightly during the same time period (C&E News, 1987; U.S. Department of Commerce, Bureau of the Census, FT 246, 1985; U.S. Department of Commerce, Bureau of the Census, FT 446, 1985). There is one producer of perchloroethylene in California (U.S. EPA, 1985b). The production capacity for this facility is estimated to be 25,000 tons per year (TPY). The facility reported stack emissions of 16 TPY and fugitive emissions of 0.45 TPY. The fugitive emissions were based on screening conducted by a private contractor, however, the EPA calculated the fugitive emissions to be 50 TPY based on a count of various equipment components and typical emission factors for the synthetic organic chemical manufacturing industry. Based on these two fugitive emission rates, the annual emissions from this facility range from 16.45 to 66 TPY (CMR. 1986; U.S. EPA, 1985b). B. USES AND EMISSIONS OF PERCHLOROETHYLENE Perchloroethylene has a wide number of uses in the United States in the industrial, governmental, and consumer sectors of the economy. These uses include dry cleaning operations, degreasing operations, production of chlorofluorocarbons, adhesives, aerosols, paints and SL 038795 A-4 DRAFT Figure III-1 U.S. Production, Imports, Exports, and Use of Perchloroethylene Production Imports Exports Use SL 038796 A-5 DRAFT coatings, and specialty chemical production. Included in Figure III-1 is annual perchloroethylene use for 1975-1985 (U.S. EPA, 1985b). Discussions of the methodologies used to estimate these usage and emissions data are included in Appendix A. Perchloroethylene is widely used in California, with approximately 19,400 tons consumed in the state in 1987 (AR8, 1989). This usage resulted in an estimated 17,600 tons per year of perchloroethylene emissions in the state. Table III-1 presents the 1987 California usage and emissions of perchloroethylene by source type in tons per year. Additional emissions may result from source types that have not yet been identified or for which data is not available. Perchloroethylene is used in California in the following products and processes: dry cleaning operations, degreasing operations, adhesives, aerosols, paints and coatings, specialty chemical production, printing inks, silicones, rug shampoos, and as a general solvent in laboratories. While a major use of perchloroethylene nationwide is as a chemical intermediate in the production of chlorofluorocarbons (CFCs), there is no chlorofluorocarbon production in California that relies on perchloroethylene as a substrate. Perchloroethylene is also handled during recycling, distribution for use, and disposal as a waste product. These operations may also result in emissions of perchloroethylene to the atmosphere. Emissions of perchloroethylene in California result primarily from stationary sources. Types of sources that have facilities in California with perchloroethylene emissions greater than 100 TPY include degreasing operations and possibly landfills. Five vapor degreasing facilities in California have been identified as having annual emissions of perchloroethylene that exceed 100 tons. Other source types that have individual facilities with emissions greater than 10 TPY include dry cleaning operations, perchloroethylene production, and possibly publicly owned treatment works (POTWs) and solvent reclamation. POTWs are municipal wastewater treatment facilities. Some of the uses in the miscellaneous category in Table III-1 may also result in significant emissions. Other source types that emit perchloroethylene include sanitary sewers, groundwater aeration, and surface impoundments. Perchloroethylene is also present in trace concentrations in emissions from waste oil combustion (U.S. EPA, 1986b). The ARB staff suspects that perchloroethylene may be emitted by motor vehicles, based on reports of chlorine emissions from diesel vehicles and the use of chlorinated hydrocarbons as diesel fuel additives (Pacyna, 1986; Tupa & Dover, 1984). Potential emissions sources not evaluated include on-site solvent recovery and industrial wastewater treatment. The major source categories are discussed below. A-6 -79I 03S DRAFT Table III-l Sources. Usage and Emissions of Perchloroethvlene in California:__ 1987 Source Type Direct Uses Tons Usage Dry Cleaning 12,857 Degreasing 3,318 Adhesives Formulation 335 Paints and Coatings 1,317 Miscellaneous 1,603 Total 19,430 Waste Management and Disposal Activities P0TWse Ground Water Aeration and Air Strippers Landfills k _ Emissions fTons/Yearl 11,314a 3,053b 335c 1,317C 1,603d 17,622 51 Not Available Not Available a. Assume 0.88 pounds of perchloroethylene is emitted for every pound used (Wolf and Myers. 1987). b. Assume 0.92 pounds of perchloroethylene is emitted for every pound used (U.S. EPA, 1985a). c. Assume 100 percent of the perchloroethylene used is emitted. d. Assume 100 percent of the perchloroethylene used is emitted. This estimate represents emissions from both the identified (961 tons per year) and unidentified miscellaneous sources. Some amounts of perchloroethylene in the miscellaneous category may actually be used in dry cleaning, degreasing, adhesive formulation, and paints and coatings. e. P0TW data is from 1986 inventory. 387gg A-7 DRAFT 1. Drv Cleaning About two-thirds (12,857 tons) of the perchloroethylene used in California in 1987 was used in the dry cleaning industry. For a dry cleaning plant to be efficient, it needs to recycle its used solvent. Most dry cleaning plants have several pieces of auxiliary equipment (e.g., filters, stills) used in the recovery and purification of perchloroethylene. Carbon adsorbers reduce process vent emissions by about 95 percent and refrigerated condensers reduce emissions by about 70 percent (U.S. EPA, 1989). For control of fugitive emissions, procedures for detecting, repairing, and preventing process leaks, and for minimizing the exposure of perchloroethylene-laden clothes to the atmosphere are used. The major use of perchloroethylene in California and nationwide is for dry cleaning. Dry cleaning occurs at coin operated facilities, commercial facilities, and industrial cleaners. Coin operated cleaners are usually part of coin operated laundry facilities offering "selfservice" dry cleaning to consumers in addition to ordinary washing machines and dryers. Commercial facilities are small neighborhood or franchise dry cleaning shops that clean soiled clothing for the consumer. Industrial dry cleaners are large cleaning plants which supply rental services of uniforms, mats, mops, and similar articles to business or industries (U.S. EPA, 1985b). Dry cleaning involves the cleaning of fabrics with nonaqueous organic solvents. The dry cleaning process involves three steps: 1) one or more solvent wash cycles; 2) a spin cycle to extract excess solvent; and 3) tumble drying (Morgan, 1985). Dry cleaning operations are typically either transfer operations in which clothes are manually transferred from the washer to the dryer, or dry-to-dry operations in which washing and drying occur in the same unit. Dry-to-dry operations tend to be used in newer installations and these units generally use less perchloroethylene per quantity of clothes cleaned than transfer units (Lauman, 1986). Steps can be taken to reuse some of the solvent through recycling operations. Due to the high cost of perchloroethylene, on-site solvent recovery is practiced at most large dry cleaning plants. Some of the waste not recovered on-site is sent for solvent recovery off-site (U.S. EPA, 1985b; DeVries, 1986). However, significant quantities of perchloroethylene wastes may be improperly disposed of as municipal waste and not sent for recycling, especially by some commercial dry cleaners (Lucks, 1987). The ARB staff estimates that the dry cleaning industry in California used an estimated 12,857 tons of perchloroethylene and emitted an estimated 11,314 tons of perchloroethylene to the atmosphere in 1987 (ARB, 1989). This emission estimate assumes that 0.88 pounds of perchloroethylene is emitted for every pound used for dry cleaning (Wolf and Myers, 1987). Based on a conversation with a solvent reclaimer for the dry cleaning industry, the staff estimates that there are approximately 4,200 dry cleaning facilities in the state (Lucks, 1987). A-8 DRAFT The major sources of emissions from dry cleaning operations are the dryer and the filter residues that result from solvent recovery, followed by the disposal of waste materials and leaks of vapor or liquid from the system (U.S. EPA, 1985b). 2. Degreasing The second most common use of perchloroethylene in California and nationwide is in degreasing operations. Degreasing is an integral part of many industrial categories such as automobile manufacturing, electronics, furniture manufacturing, appliance manufacturing, and textile, paper, plastic, and glass manufacturing. PEI Associates estimates that there are several thousand cold cleaning, vapor, and conveyor type degreasing units in operation in California (PEI, 1986). There are more than 350 vapor degreasers using perchloroethylene in California (ARB, 1987). A significant portion of the perchloroethylene used for degreasing is sent to solvent reclaimers (U.S. EPA, 1985b). Approximately 17 percent of the perchloroethylene used in California in 1987 was used in organic solvent cleaning (degreasing) operations. There are three basic types of degreasing equipment: cold cleaners, open top vapor degreasers, and conveyorized cleaners. In California, most of the perchloroethylene emissions from degreasing result from open top vapor degreasers. A large portion of emissions results from loss of solvent vapor from tanks due to diffusion and convection, and carry-out of solvent on cleaned parts. Other emissions result from equipment leaks and from solvent storage and handling. Control techniques include adding, equipment covers, increasing freeboard area, adding freeboard chillers, and providing drainage racks. Carbon adsorption may be used to recover solvent vapors. Operating practices that reduce solvent exposure to the atmosphere include keeping covers closed, fully draining cleaned parts, and maintaining moderate conveyor speeds and ventilation rates. Solvent degreasers function by immersing the part to be cleaned in the liquid or vaporized solvent contained in the degreasing tank. In a typical cold cleaning operation, dirty parts are cleaned manually by spraying and then by soaking in the tank. Open top vapor degreasers clean with the condensation of hot solvent vapor on colder metal parts. Conveyorized degreasers may operate with either cold or vaporized solvent. They are continuously loaded and are usually hooded or enclosed. After cleaning, the parts are either suspended over the tank to drain or placed on an external rack that directs solvents back into the tank (U.S. EPA, 1985b; U.S. EPA, 1977). It is assumed that approximately 0.92 pounds of perchloroethylene is emitted for every pound of fresh perchloroethylene used in degreasing (U.S. EPA, 1985a). Thus perchloroethylene emissions from degreasing operations in California during 1987 are estimated to be approximately 3,053 tons (ARB, 1989). 038800 SL A-9 DRAFT Solvent evaporation occurs with all types of degreasing equipment. The major sources of evaporation are the degreasing tank, the carry-out of solvent on cleaned parts, and the disposal of waste solvent. The amount of these emissions varies with the type of equipment used and the operating parameters. 3. Adhesive Formulation Less than two percent of the perchloroethylene used in California in 1987 was used in adhesive formulation. One hundred percent of the perchloroethylene used in adhesive formulation is assumed to be emitted into the atmosphere upon application of the adhesive. Perchloroethylene emissions from adhesive application in California during 1987 are estimated to be approximately 335 tons (ARB, 1989). 4. Paints and Coatings Less than seven percent of the perchloroethylene used in California in 1987 was used in paints and coatings formulation. One hundred percent of perchloroethylene used in paints and coatings formulation is assumed to be emitted to the atmosphere upon application of the paint or coating. Perchloroethylene emissions from paints and coatings in California during 1987 are estimated to be approximately 1,317 tons (ARB, 1989). 5. Miscellaneous About eight percent of the perchloroethylene used in California in 1987 was used in miscellaneous applications. There are not enough data to accurately identify the emissions from each source of perchloroethylene in the miscellaneous category. However, in addition to the source types described above, perchloroethylene is also used as a solvent and carrier in aerosol products, such as spray paints and cleaners; as a solvent in pharmaceutical and textile processing industries; as an ingredient in blended solvents; as a carrier solvent in printing inks, cleaners, polishes, and lubricants; as a recyclable dielectric fluid for power transformers; as a heat transfer medium; and as a pesticide intermediate. One hundred percent of the perchloroethylene used in the miscellaneous category is assumed to be emitted in the atmosphere. Thus, perchloroethylene emissions from the miscellaneous category in California during 1987 are estimated to be approximately 1,603 tons (ARB, 1989). About 961 tons were actually identified as miscellaneous in the survey; the remaining 642 tons which were not identified are assumed to be miscellaneous. This estimate may be lower if perchloroethylene usage in the miscellaneous category is identified for dry cleaning, degreasing, or other purposes. Also, industry has begun to use perchloroethylene as a dielectric fluid in transformers to replace polychlorinated biphenols (PCBs) (HSIA, 1987). A-10 DRAFT 6. Distribution. Recycling, and Disposal In addition to the uses discussed above, perchloroethylene is handled by distribution facilities, solvent reclaimers, POTWs, municipal landfills, ground water aeration treatment facilities, surface impoundments, and hazardous waste landfills. Emissions of perchloroethylene result from all of these handling operations. The perchloroethylene handled during distribution, recycling, and disposal operations has already been included in Figure 1 and on Table III-1. Almost all of the perchloroethylene shipped directly for use in the U.S., other than for use in chlorofluorocarbon (CFC) production, reaches the consumer through chemical distribution. Generally, sales of perchloroethylene to CFC producers are made directly and not through the chemical distribution system. Because there is no CFC production in California where perchloroethylene is used as a substrate, the ARB staff assumed that 100 percent of the perchloroethylene shipped directly into California is sold by chemical distributors. Distribution operations involve transport, storage, and repackaging of perchloroethylene. There are an estimated 500 regional distribution facilities in the U.S. (U.S. EPA, 1985b). An estimated 24,100 tons of perchloroethylene were sold through distribution facilities in California in 1985. These facilities emitted an estimated 7 tons of perchloroethylene in 1985. Emissions resulting from storage tanks were calculated using equations and assumptions about the throughput of perchloroethylene and the number and size of storage tanks (U.S. EPA, 1985b; U.S. EPA,- 1985c). There are approximately 20 commercial solvent reclaimers in California (CDHS, 1984). The three largest reclaimers process most of the solvents sent for solvent reclamation in California (DeVries, 1986). An estimated 1,940 tons of perchloroethylene were sent for recycling in California in 1985. An estimated 5 to 20 tons of perchloroethylene emissions resulted from solvent reclamation facilities in California during 1985. Emissions primarily result from the storage and handling of waste and reclaimed solvents, although some emissions come from distillation condensers at reclamation facilities. 7. Publicly Owned Treatment Works Publicly owned treatment works (POTWs) are municipal wastewater treatment facilities. POTWs are potential sources of perchloroethylene emissions because they may receive and treat commercial or industrial discharge containing perchloroethylene. Undetermined quantities of perchloroethylene are discharged to POTWs throughout California. Discharges of perchloroethylene to POTWs are assumed to result primarily from commercial and industrial sources. Perchloroethylene emissions from POTWs in California are estimated to be 51 TPY (Chang et al., 1987). Approximately 20 TPY of these emissions result from POTWs in Los Angeles County and 15 TPY from POTWs in Alameda County (Chang et al., 1987). While there is considerable uncertainty associated with these SL 038802 A-11 DRAFT estimates, the data are the most complete data available. In addition to the emissions from treatment plants, perchloroethylene is emitted in route to POTWs at pump stations and at certain high points in sewer lines. We do not have data on total emissions from sources such as pump stations or sewer lines. Major emitters of perchloroethylene were the East Bay Municipal Utility District Wastewater Treatment Facility (13 tons per year), San Oose-Santa Clara Water Pollution Control Plant (7.3 tons per year), and the Joint Water Pollution Control Plant (in L.A. County) (5.9 tons per year). 8. Landfills Landfills are potential sources of perchloroethylene emissions in California. However, there are presently not enough data available to estimate statewide perchloroethylene emissions from landfills. The quantity of perchloroethylene disposed of at municipal landfills is also undetermined. Perchloroethylene might be disposed of at municipal landfills as residues in discarded aerosol and paint cans, as well as by other means. Estimating landfill emissions is complicated by the variation in landfill contents, surface area, cover depth, and emissions may be large. Furthermore, emissions for some of the state's landfills are reduced by the use of gas recovery well systems. There are about 1,000 active landfills and 1,200 closed landfills in California (Barnickol, 1986). 9. Groundwater Aeration and Air Stripping Towers In 1987, an EPA contractor identified 21 facilities with air strippers in California (U.S. EPA, 1987a). It is not known how many of these facilities actually treat contaminated groundwater containing perchloroethylene. Perchloroethylene emission estimates were available for only one facility in California. This facility had annual emissions of perchloroethylene of approximately 0.7 tons. There are presently no statewide perchloroethylene emissions data from air stripping towers. Perchloroethylene is also emitted from ground water aeration treatment facilities throughout the state. Total emissions from such facilities in California have not been determined. The ARB staff estimated that from 1 to 14 TPY of perchloroethylene emissions may result from groundwater aeration facilities in the South Coast Air Basin and the San Francisco Bay Area Air Basin. The upper estimate includes unknown quantities of other contaminants as well as perchloroethylene. This estimate is based on permit information on file at the South Coast Air Quality Management District and the Bay Area Air Quality Management District (SCAQMD, 1987; Lopez, 1987). The quantity of perchloroethylene emissions from different facilities varies widely, depending on the concentration of the contaminant in the groundwater, the flow rate of the water through the treatment process, and the hours of operation. A-12 0383 C. EMISSION TRENDS DRAFT In early 1986, the staff of the Chemical Marketing Reporter expected no growth in U.S. consumption of perchloroethylene through the year 1990. This zero growth in usage is accounted for by a projected decrease in the use of perchloroethylene for dry cleaning and degreasing, balanced by an increase in the use of perchloroethylene as a chemical intermediate (CMR, 1986). Two reasons for a near term decline of perchloroethylene usage in dry cleaning are 1) new commercial dry cleaners tend to use more efficient dry-to-dry equipment instead of transfer equipment and 2) industrial dry cleaners are increasingly using water and detergent solutions in place of perchloroethylene (Wolf, 1986; Lauman, 1986; DeVries, 1986). This loss is expected to be offset by growth in demand for trichlorotrifluoroethane (CFC-113) as a solvent in the electronics industry (CMR, 1986). Perchloroethylene is a chemical intermediate in CFC-113 production. While the net U.S. usage of perchloroethylene will remain fairly constant, ARB staff expects a decline in nationwide emissions of perchloroethylene through 1990. This is because a much larger percentage of the perchloroethylene used in dry cleaning and degreasing operations is emitted than from the use of perchloroethylene as a chemical intermediate. Emissions of perchloroethylene in California are expected to decline at a rate similar to the decline in U.S. emissions. Through the 1990's, however, perchloroethylene emissions in California may well increase due to projected population increases. The majority of perchloroethylene emissions result from dry cleaning operations and the use of dry cleaning services is related to the population. The population in California is projected to increase approximately 20 percent from 1984 to 2000 (U.S. Department of Commerce, Bureau of Census, 1985). Such an increase in population will probably result in a significant increase in perchloroethylene emissions, provided other factors such as level of control remain equal. Perchloroethylene may be increasingly used in the future as a dielectric fluid in transformers to replace environmentally persistent materials (HSIA, 1987). Although no data on projected usage are available, the staff expects the emissions from this use to be minimal. In addition to increases from CFC production, governmental regulation could result in small increases in perchloroethylene emissions from some source types. Efforts to control perchloroethylene residuals from sources such as dry cleaners might result in increased emissions from POTWs and solvent reclamation. In fact, the quantities of perchloroethylene received by solvent reclaimers have increased significantly over the past few years and the trend may continue. Concerns and restrictions on land disposal of halogenated solvents have prompted much of this trend. SL 038804 A-13 DRAFT D. POTENTIAL SOURCES OF INDOOR PERCHLOROETHYLENE 1. Building Materials and Consumer Products The National Aeronautics and Space Administration (NASA), has been compiling emission data for volatile organic compounds since 1975. The database covers more than 5,000 materials used in the space program. Some of these materials are commercially available and are used in homes and offices. Based on their analysis of this data base, Ozkaynak, et al. (1987) reported that perchloroethylene would be emitted from a number of common household products such as adhesives, foam, cosmetics and ink. Recent emission tests conducted in chambers by different investigators found no perchloroethylene emissions from a variety of building materials such as caulks, adhesives, stains, paints, waxes and furniture polishes (Berglund et al., 1987; Knoppel & Schauenburg, 1987; Tichenor, 1987; Wallace et al., 1987b, Girman et al., 1986 and Molhave, 1982). In a national survey, U.S. EPA (1987c) selected consumer products containing chlorinated solvents for chemical composition analysis. The samples were brand-name products selected randomly from six U.S. cities and represented 67 categories of household solvent products. The survey found perchloroethylene in products from 18 of the 67 categories. The three categories with the highest percentage of brand-name products containing some amount of perchloroethylene were: brake quieters/cleaners (58 percent), water repellents (25 percent) and fabric finishes (20 percent). Other categories with more than ten percent (11 percent - 17 percent) of perchloroethylene-containing products include: specialized aerosol cleaners, ignition wire driers, suede protectors, spot removers and spray cleaners. The concentration of perchloroethylene in individual products varies greatly. Some of the products contain as much as 90 percent perchloroethylene by weight. 2. Drv-Cleaned Mater_ia1s and Drv Cleaning-Related Activities Perchloroethylene is the primary solvent for dry cleaning clothing and fabrics. Freshly dry-cleaned garments probably serve as an emission source of indoor perchloroethylene, when brought home. According to a U.S. EPA (1987b) national survey of household solvent usage, the average frequency of household use of dry cleaning services is 0.96 times per month. In reviewing all of the TEAM study results, Wallace and Clayton (1987) concluded that perchloroethylene levels in air samples and breath samples were elevated for persons who were employed in or who had recently visited a dry cleaning shop. They also concluded that the main source of exposure appears to be wearing and storing dry-cleaned clothes. In following the personal activities of seven individuals, increased exposure of perchloroethylene was associated with visiting dry cleaning shops and laboratories using chemicals, and with using cleaning solvents (Wallace et al., 1987c). A-14 DRAFT 3. Vaporization from Water Sources Water can serve as a medium to carry pollutants from outdoor to indoor environments. Once in contact with air indoors, volatile and semi-volatile chemicals such as perchloroethylene can leave the water and enter the air. Human activities such as cooking, heating or showering with contaminated water can promote rapid vaporization of perchloroethylene from water. Industrial solvent-contaminated surface or ground water may, therefore, bring outdoor perchloroethylene indoors via the water supply. In California, surface water is generally free of perchloroethylene. In assessing ground water quality, the California Department of Health Services (CDHS, 1986) reported that about 7 percent of the wells for large public water systems were contaminated with perchloroethylene. About half of these wells contained water concentrations of perchloroethylene below 2 ug/1 (2 ppb). Water from wells with exceptionally high perchloroethylene concentrations are usually blended with water from wells with lower perchloroethylene concentrations to meet the water quality guideline of 4 ug/1 (4 ppb). Based on the information supplied by Spath (1987), the concentration of perchloroethylene in most California domestic water supplies is essentially below 1 ug/1 (1 ppb). Domestic water supply systems may be a source of indoor air emission of perchloroethylene. In the eastern states, some water distribution systems contained vinyl-toluene-lined asbestos cement pipes. The liner was sprayed on the inside of pipes with perchloroethylene-dissolved resin. High levels of perchloroethylene were found in water distributed by this type of system (Larson et al., 1983). The existence of this same water supply system in California is, however, unknown. 4. Other Factors that May Influence Indoor Concentrations Homes built on or near landfills containing perchloroethylene or related chlorinated hydrocarbons may accumulate perchloroethylene indoors. The rate of accumulation varies, depending heavily on the soil permeability, source strength, air exchange rate and structure of the house. Kliest et al. (1987) reported that houses built on polluted soil had almost four times higher perchloroethylene levels in the crawl-space than houses on clean soil. The relationship between the concentration of a pollutant in the crawl space and its concentration in the living space is still under investigation. Perchloroethylene from known emission sources may increase the indoor levels of nearby residential or commercial buildings. Of particular concern are small dry cleaning shops which usually locate in densely populated areas. In an EPA study of self-serve laundries with dry cleaning operations, similar perchloroethylene levels, around 10,000 mg/m (1500 ppb), were found in a laundry and in an occupied apartment directly above the laundry (Howie & Elfers, 1981). SL 038806 A-15 DRAFT Perchloroethylene is a registered pesticide in California for controlling wasps and hornets. However, the California Department of Food and Agriculture does not have any information on product types to indicate perchloroethylene use indoors (Formoli, 1987). As mentioned earlier, perchloroethylene was detected in the EPA analysis of consumer products used as pesticides (Wallace et al., 1987b). 5. Summary of Perchloroethylene Sources Elevated air concentrations of perchloroethylene in indoor environments appear to result from both indoor sources of perchloroethylene and mechanisms by which outdoor perchloroethylene is concentrated indoors. Consumer products appear to be a significant emission source of indoor perchloroethylene. In addition, dry-cleaned clothing is a likely additional source of indoor perchloroethylene, although the relative contributions from these two emission source categories cannot be quantified. In general, building materials may be a minor source of perchloroethylene. The possibility of perchloroethylene from outdoor sources becoming concentrated indoors requires more research. The use of water containing perchloroethylene indoors may increase indoor levels of perchloroethylene. The increased number of local dry cleaning shops in densely populated areas may also strongly influence perchloroethylene levels in nearby indoor environments. A-16 03*01 5^ References for Chapter III DRAFT ARB, 1987. "Data on Perchloroethylene Usage in Degreasers Retrieved from Emission Data System (EDS) using Source Classification Codes Data Base Year 1983." Emission Inventory Branch, Technical Support Division, Sacramento, CA. ARB, 1989. Data extracted from the halogenated solvent survey conducted in July and December, 1988. Emission Inventory Branch, Technical Support Division, Sacramento, CA. Barnickol, J., 1986. Personal communication. Telephone conversation between Ronald Rothacker of CARB and Jeff Barnickol of the California Water Resources Control Board. July 16, 1986. Berglund, B.; I. Johansson and T. Lindvall, 1987. "Volatile Organic Compounds from Building Materials in a Simulated Chamber Study." In: Indoor Air '87 Volume 1--Volatile Organic Compounds, Combustion Gases, Particles and Fibers, Microbiological Agents, pp.16-21. C 4 E News, 1987. Chemical and Engineering News. "Production by the U.S. Chemical Industry." June 8, 1987. CDHS, 1984. California Department of Health Services. Directory of Industrial Recvclers. California Waste Exchange. March 1984. CDHS, 1986. California Department of Health Services. Final Report on a Monitoring Program for Organic Chemical Contamination of Large Public Water Systems in California. Chang, D.; E. Schroeder and R. Corsi, 1987. "Emissions of Volatile and Potentially Toxic Organic Compounds from Sewage Treatment Plants and Collection Systems". Submitted under contract to the ARB. July 1987. CMR, 1986. Chemical Marketing Reporter. "Chemical Profile Perchloroethylene." February 3, 1986. DeVries, P., 1986. Personal Communication. Telephone Conversation between Ronald Rothacker of CARB and Paul DeVries of Roehl Disposal Corporation. September 22, 1986. Formoli, T., 1987. Personal Communication. Telephone Conversation between Steve Hui of CARB and T. Formoli of California Department of Food and Agriculture. October 6, 1987. Girman, J. R.; A. T. Hodgson and A. S. Newton, 1986. "Emissions of Volatile Organic Compounds from Adhesives with Indoor Applications." Environment International 12: 317-321. SL A-17 DRAFT HSIA, 1987. Halogenated Solvents Industry A11iance. "Perchloroethylene White Paper." August 1987. Howie, S. J.; and L. A. Elfers, 1981. "Ambient Perchloroethylene Levels Inside Coin-Operated Laundries with Dry Cleaning Machines on the Premises." EPA-600/4-82-032. Kliest, J.; T. Fast and J. S. M. Boleij, 1987. "The Relation between Soil Pollution and Indoor Pollution with Volatile Organic Compounds." In: Indoor Air '87 Volume 1--Vo 1 ati1e Organic Compounds, Combustion Gases, Particles and Fibers, Microbiological Agents, pp. 122-125. Knoppel, H.; and H. Schauenburg, 1987. "Screening of Household Products for the Emission of Volatile Organic Compounds." In: Indoor Air '87 Volume 1--Volatile Organic Compounds, Combustion Gases, Particles and Fibers, Microbiological Agents, pp.27-31. Larson, C. D.; 0. Love and G. Reynolds, 1983. "Tetrachloroethylene Leached from Lined Asbestos-Cement Pipe into Drinking Water." J. Am Water Works Assoc. 75(4):184-188. Lauman, G., 1986. Personal Communication. Telephone Conversation between Ronald Rothacker of CARB and George Lauman, Executive Director of the California Fabricare Institute, Cupertino, CA. September 10, 1986. Lopez, S., 1987. Correspondence from Sandra Lopez of the Bay Area Air Quality Management District to Rich Bradley of CARB. June 4, 1987. Data on emissions from groundwater aeration treatment facilities in the BAAQMD. Lucks, J., 1987. Personal Communication. Telephone Conversation between Ronald Rothacker of CARB and John Lucks of Safety Kleen Corporation. November 6, 1987. Molhave, L., 1982. "Indoor Air Pollution due to Organic Gases and Vapours of Solvents In Building Materials." Environment International 8:117-127. Morgan, D. L., 1985. Correspondence from Donald L. Morgan of Cleary, Gottlieb, Steen, and Hamilton to R. E. Rosensteel, EPA/ESED. March 1, 1985. Response for Halogenated Solvents Industrial Alliance Concerning Industrial Consumption Volumes of Perchloroethylene in 1983. Ozkaynak, H. et al., 1987. "Source and Emission Rates of Organic Chemical Vapors in Homes and Buildings." In: Indoor Air '87 Volume 1--Volatile Organic Compounds, Combustion Gases, Particles and Fibers, Microbiological Agents, pp. 3-7. A-18 SL 038809 DRAFT Pacyna, J. M., 1986. "Emission Factors of Atmospheric Elements." in Toxic Metals in the Atmosphere. Nriagu, J. 0., and Davidson, C. I., Ed. Wiley, New York. PEI Associates, Inc., 1986. "Occupational Exposure and Environmental Release Assessment of Methylene Chloride." April 12, 1986. SCAQMD, 1987. South Coast Air Quality Management District. Computer Printout from Chemical Unit Files on Groundwater Treatment by Air Stripping. October 16, 1987. Spath, 1987. Personal Correspondence, David Spath, staff of California Department of Health Services, (10/8/87 Memo). Tichenor, B. A., 1987. "Organic Emission Measurements via Small Chamber Testing. In: Indoor Air '87 Volume 1--Volatile Organic Compounds, Combustion Gases, Particles and Fibers, Microbiological Agents, pp. 8-13. Tupa, R. C.; and C. J. Dover., 1984. SAE Paper No. 841211. U.S. Department of Commerce, Bureau of the Census. "Statistical Abstract of the United States 1986." 106th Edition. December 1985. U.S. Department of Commerce, Bureau of the Census. "U.S. Exports, Schedule B Commodity By County." FT 446/Annual, 1975-1985. U.S. Department of Commerce, Bureau of the Census. "U.S. Imports for Consumption and General Imports, TSUSA Commodity by County of Origin." FT 246/Annual 1975-1985. U.S. EPA, 1977. "Control of Volatile Organic Emissions from Solvent Metal Cleaning, OAQPS Guidelines." EPA-450/2-77-022. November 1977. U.S. EPA, 1985a. "Emission Test Report No. 85-HWS-l." Tests conducted at IT Corporation Benicia Facility, Benicia, CA. May 1985. U.S. EPA, 1985b. "Survey of Perchloroethylene Emission Sources." EPA-450/3-85-017, Research Triangle Park, NC. June 1985. U.S. EPA, 1985c. "Compilation of Air Pollutant Emission Factors, AP-42." Fourth Edition. September 1985. U.S. EPA, 1986b. "Compiling Air Toxics Emission Inventories. July 1986. U.S. EPA, 1987a. "Air Stripping of Contaminated Water Sources - Air Emission and Controls." Final Report prepared by Radian Corporation. EPA-450/3-87-017. Research Triangle Park, NC. SL 038810 A-19 DRAFT U.S. ERA, 1987b. "Household Solvent Products: A National Usage Survey." EPA-OTS 560/5-87-005. U.S. EPA, 1987c. "Household Solvent Products: A 'Shelf1 Survey with Laboratory Analysis." EPA-OTS 560/5-87-005. U.S. EPA, 1989. "Locating and Estimating Air Emissions from Sources of Perchloroethylene and Trichloroethylene." Revised Draft Report, April 4, 1989 prepared by Radian Corporation, Research Triangle Park, NC. Wallace, L. A.; C. A. Clayton, 1987. "Volatile Organic Compounds in 600 US Homes: Major Sources of Personal Exposure." In: Indoor Air '87 Volume 1--Volatile Organic Compounds, Combustion Gases, Particles and Fibers, Microbiological Agents, pp.183-187. Wallace L. A. et a 1., 1987b. "Emissions of Volatile Organic Compounds from Building Materials and Consumer Products." Atmospheric Environment 21(2): 385-393. Wallace, L. A. et al., 1987c. "The Influence of Personal Activities on Exposure to Volatile Organic Compounds." In: Indoor Air '87 Volume 1--Volatile Organic Compounds, Combustion Gases, Particles and Fibers, Microbiological Agents, pp. 117-121. Wolf, K., 1986. Personal Communication. Telephone Conversation between Ronald Rothacker of CARB and Dr. Katy Wolf of Rand Corporation, Santa Monica, CA. September 25, 1986. Wolf, K.; C. W. Myers, 1987. "Hazardous Waste Management by Small Quantity Generators - Chlorinated Solvents in the Dry Cleaning Industry." Rand Corporation. A-20 038^ DRAFT IV. EXPOSURE--LQ PERCHLQRflETHYLENE A. AMBIENT MONITORING FOR PERCHLOROETHYLENE The current toxics sampling network comprises 20 monitors statewide. Nine of these monitors are located in Southern California (south of Bakersfield), while the other 11 are located in the northern portion of the state. The statewide monitoring sites are indicated on Figure IV-1. Data used in the following exposure analysis were collected during the period of July 1988 through June 1989. The data for this period (hereafter referred to as "the study period") represent the most recent period for which the available data are known to be of consistent and verifiable quality. Individual 24-hour samples were collected in Tedlar bags and subsequently analyzed for perchloroethylene using gas chromatography with an electron capture detector. Standard operating procedures for sampling and analysis are provided in Appendix B. The monitor located in El Monte at the beginning of the study period was subsequently moved to Azusa where it stayed for approximately one month. It was then moved to Burbank where it stayed for the remainder of the study period. Because there are considerable missing data from El Monte, Azusa, and Burbank, data for these sites were not used in estimating a mean annual statewide exposure. The months during which data are available for other monitoring sites are summarized in Table IV-1. B. AMBIENT CONCENTRATIONS OF PERCHLOROETHYLENE 1. Study Period Data Collected The number of samples available per site during the study period range from 16 to 23, averaging 20.6 observations per site for the study period. Observations below the 0.01 parts per billion (ppb) limit of detection (LOD), are referred to as being partially observed data (POD). The only observation below the LOD was reported for the San Jose site on August 13, 1989. 038812 A-21 DRAFT Figure IV-X North Caul Uk* County IJJ1* * \ --H ..... *--i j Ir i i 1 i * T NoftkaUt I FUimii j-- : 1I SVaaclrlaamy#Yi<* V. i- AO* * \r^----- 1 J-fsT1.y^v-'----1** ?-]j CMoouunntitaaain Lah* Taiio* AIR RESOURCES BOARD TOXICS NETWORK MONITORING SITES (July 1988 - June 1989) UttnJ (7) SAN FRANCISCO (fl) MERCED (2) FREMONT (T3) SIMI VALLEY (5) RICHMOND (J) SAN IOSE (13) RUBIDOUX (Q LONG SEACN (s) CONCORD (fs) LA-NORTH MAIN (T) CITRUS HEIGHTS @ EL MONTE (7) MODESTO @ UPLAND (a) STOCETON (s) FRESNO (JO: BAJEERSFIELD 18) SANTA BARBARA 13) CHULA VISTA @ EL CAJON San Joaquin Vallay NivoDuthiaArtat t* llt HfAVt Lt*f * DC < IMf A ri, t CALItON*IA AIN lAMRS ^ i* Hitt S*n Diao 7 @ ------- A-22 Sti-i.rdrr Saurci 2,'>al,n 1 -*6> SL 038813 DRAFT TABLE IV-1 Months Where At Least One Sample Was Collected And Analyzed For Perchloroethylene3: July 1988 through June 1989 Site Location JASONDJFMAMJ Samples South Coast Air Basin Long Beach ooooooooo oo Los Angeles ooo oooooooo Rubidoux ooooooooo oo Upland oooooooooooo South Central Coast Air Basin Santa Barbara ooooooooooo Simi Valley o ooooooooo San Diego Air Basin El Cajon oooooooooooo Chula Vista ooooooooo oo San Francisco Bay Area Air Basin Concord oooooooooooo Fremont oooooo oooooo Richmond ooooooo oooo San Francisco oo oooooooo San Jose ooooooooooo San Joaquin Valley Air Basin Bakersfield ooooooooooo Fresno oooooooooooo Merced oooooo oo Modesto oooooooooooo Stockton ooooooooooo Sacramento Valley Air Basin Citrus Heights ooooo oooooo 23 18 23 23 17 23 22 21 21 23 23 16 22 18 23 15 23 19 19 a. A "o" indicates at least one sample was collected during the month. SL 038814 A-23 DRAFT Even though no value is reported when a concentration is below the LOD, important information is still available from its existence. Obviously, the fewer the data points below the LOD, the less significant their impact on the analysis. Because removing these values from the dataset would bias the sample statistics and exposure estimates, an estimated order statistic approach was used to calculate a replacement value for the POD.* 2. Site-Specific Concentrations The absolute range of concentrations sampled at each site during the study period is summarized in Table IV-2. Minimum, maximum, median and mean concentrations, and standard deviation are reported for each site. Mean perchloroethylene concentrations were calculated as the mean of available monthly means, because this approach provides equal weighting for each month when the number of samples per month varies. Standard deviations were calculated from monthly means so they could be compared with the annual mean concentrations. The site-specific ranges of the ambient perchloroethylene concentrations listed in Table IV-2 are plotted in Figure IV-2. Minimum concentrations range from POD at San Jose to 0.21 ppb at Los Angeles. Maximum concentrations range from 0.18 ppb at Bakersfield to 4.80 ppb at Simi Valley. Mean annual concentrations range from a minimum of 0.10 ppb at Citrus Heights and Bakersfield to a maximum of 0.70 ppb at Concord. Two relatively unusual concentrations were reported for Chula Vista; 13.00 ppb on March 6, 1989, and 14.00 ppb on April 4, 1989. values are more than 70 standard deviations from the mean of the remaining Chula Vista values. We believe that these values do not represent ambient concentrations and are possibly the result of influences from an unknown nearby source. These values were not included in the calculations of the various exposure estimates (e.g., site, basin and state-wide). Beth At all 19 monitoring sites, the median of the reported data is less than or equal to the mean of all values. A median concentration that is significantly different from the mean indicates the data has a non normal distribution. In the past, the distribution of the ambient toxic air contaminant data has generally been assumed to be log-normal. The perchloroethylene data were tested for log-normality using the 1. Estimating order statistics is our preferred method for estimating values for POD. This method was discussed by Gleit (1985) and uses an iterative algorithm for estimating the order statistics for a finite sample. A-24 SL 038815 T fCfl o Co 00 M00 5.0-r 4.5 4.0- 3.8- 3.0 Figure IV-2 MEAN ANNUAL PERCHLOROETHYLENE CONCENTRATION PLOTTED USING EXTENDED BOX PLOTS BASED ON JULY 19B8-JUNE 1989 DATA LEGEND -MAXIMUM -Sib QUANTILE -731b QUANTILE -MEAN -MEDIAN -EBtb QUANTILE -Bib QUANTILE -MINIMUM P erchloroethylene in Parts Per B illio n iCBOLA flS T A Monitoring Sites Throughout California A ft B/TSD: October 69 TABLE IV-2 DRAFT Minimum, Maximum, Median and Means of Perchloroethylene Samples Collected during July 1988 through June 1989 (PPb) Air Basin Site Location Minimum Maximum Median Mean3 Standard1 Cone. Cone. Cone. Cone. Deviatioi South Coast Air Basin Long Beach 0.13 Los Angeles 0.21 Rubidoux 0.05 Upland 0.10 Basin Summary 0.05 1.60 1.90 0.52 1.10 1.90 South Central Coast Air Basin Santa Barbara 0.06 Simi Valley 0.07 Basin Summary 0.06 0.33 4.80 4.80 San Diego Air Basin Chula Vista El Cajon Basin Summary 0.07 0.07 0.07 0.69 1.10 1.10 San Francisco Bay Area Air Basin Concord 0.07 3.30 Fremont 0.06 0.73 Richmond 0.03 0.41 San Francisco 0.08 0.97 San Jose POD 0.87 Basin Summary POD 3.30 * San Joaquin Valley Air Basin Bakersfield 0.04 Fresno 0.05 Merced 0.06 Modesto 0.04 Stockton 0.04 Basin Summary 0.04 0.18 0.68 0.39 0.83 0.26 0.83 Sacramento Valley Air Basin Citrus Heights 0.03 0.21 0.31 0.51 0.23 0.40 0.35 0.14 0.17 0.15 0.25 0.34 0.27 0.54 0.26 0.11 0.17 0.15 0.15 0.09 0.12 0.15 0.09 0.11 0.12 0.08 0.42 0.57 0.26 0.48 0.43 0.35 0.38 0.08 0.24 0.29 0.16 0.39 0.28 0.07 0.47 0.34 0.30 0.42 0.36 0.17 0.26 0.22 0.70 0.27 0.11 0.20 0.23 0.30 0.59 0.13 0.06 0.15 0.17 0.29 0.10 0.18 0.19 0.16 0.12 0.15 0.03 0.14 0.06 0.12 0.05 0.09 0.10 0.05 a. Basin Means are the mean of the site means. b. Basin Standard Deviations are pooled values of the standard deviations across sites within a basin. A-26 SL 038817 DRAFT Shapero-Wilk test (Shapero and Wilk, 1965). We found that perchloroethylene data are not distributed log-normally, although they are skewed like a log-normal distribution. 3. Basin-Wide Mean Concentrations The highest basin-wide mean concentration occurred in the South Coast Air Basin (0.43 ppb), followed by the San Diego Air Basin (0.36 ppb), San Francisco Bay Area Air Basin (0.30 ppb), South Central Coast Air Basin (0.28 ppb), and San Joaquin Valley Air Basin (0.15 ppb). The Sacramento Valley Air Basin was intentionally left out of the basin wide analysis as there is only one monitoring site in the basin (see Table IV-2). Air basins are identified to reflect similar geographic and meteorological conditions. Therefore, data are compared as site composites within each basin. Data for all sites within an air basin are ranked within each air basin. For example, the lowest value across all sites within an air basin has rank one. The Kruskal-Wal1 is test (Conover, 1971) is then used to determine whether within a basin there are data for a site that are significantly different than data for another site. If the Kruskal-Wal1 is test statistic is not significant, then the medians are not considered to be different. If the Kruskal-Wal1 is test statistic is significant, then Fisher's Least Significant Difference test (SAS Institute Inc., 1985) is used to determine the sites considered to be different and those that are similar. The most likely groupings for sites within each basin based on available ranked data are given in Table IV-3. If two sites within a basin have no associated letters in common, then the medians for each of the two sites are considered to be significantly different. It should be noted that test statistics based on rankings of the data rather than the actual reported value will sometimes lead to less than intuitive results. * 4. Peak-to-Mean Ratios Peak-to-mean ratios were calculated to provide possible insights into the nature of perchloroethylene emission release patterns. The technique is based on observations made for criteria pollutants. For example, carbon monoxide is a relatively inert criteria pollutant with generally widespread emissions. Over an annual period, peak-to-mean ratios for carbon monoxide tend to be fairly low, generally less than five. Another criteria pollutant, sulfur dioxide, may be emitted from widespread sources, but is also emitted from localized point sources. Peak-to-mean ratios for sulfur dioxide at sites influenced by localized sources tend to be much greater than five. Ratios between 50 and 90 have been seen at some locations. Based on what we know about the characteristics of the criteria pollutants, we can divide peak-to-mean ratios into low and high ratios. Generally, a low peak-to-mean ratio. SL 038818 A-27 DRAFT TABLE IV-3 Summary of Intra-Basin Site Differences Based on Average Monthly Rankings and Monthly .Means AIR BASIN Site Location Median9 Monthly Mean Groupings Rank Meanb Cone. Median Cone. South Coast Air Basin Los Angeles Long Beach Upland Rubidoux A A AB B South Central Coast Air Basin Simi Valley A Santa Barbara A San Diego Air Basin El Cajon Chula Vista A B San Francisco Bay Area Air Basin Concord A Fremont B San Jose B San Francisco B Richmond C San Joaquin Valley Air Basin Merced A Fresno AB Modesto B Stockton B Bakersfield B Sacramento Valley Air Basinc Citrus Heights 5.40 5.20 4.45 2.90 24.18 17.56 25.21 18.64 77.88 61.54 49.46 48.81 28.04 66.43 54.02 47.17 46.58 39.00 -- 0.64 0.42 0.46 0.26 0.44 0.15 0.42 0.27 0.75 0.29 0.23 0.21 0.11 0.18 0.18 0.16 0.12 0.10 0.10 0.51 0.31 0.35 0.23 0.17 0.14 0.34 0.25 0.54 0.26 0.15 0.17 0.11 0.15 0.12 0.09 0.11 0.09 0.08 a. Sites with the same letter are not considered to be significantly different. b. Means are computed as the mean of all observations. c. Since there is only one toxic monitoring site in the Sacramento Valley, no results are given for this basin other than the mean. A-28 SL 038819 DRAFT less than about ten, indicates either relatively constant and/or uniform emission sources or few emission sources but high, fairly constant background concentrations. A high peak-to-mean ratio, generally greater than about ten, usually indicates either intermittent and/or scattered emission sources or scattered emission sources with a highly variable background concentration. Peak-to-mean ratios for each site during the study period are given in Table IV-4. Ratios for 17 of the 19 sites are less than five. The two sites reporting data having peak-to-mean ratios of five or larger were Simi Valley and Modesto, with peak-to-mean ratios of 10.82 and 5.05, respectively. The high ratio reported for Simi Valley may be due to the influence of a nearby unknown source, however, without further data it is not possible to determine any specific reasons for the relatively high peak-to-mean. The relatively low ratios at all other sites suggest fairly consistent emission patterns across most of the state and throughout the year. Potential isolated large sources do not appear to impact the sampling network significantly. This is also apparent by comparing the Coefficient of Variation (C.V.) for each site. The C.V. is similar to the peak-to-mean ratio in that it is a unitless number that is used to compare the dispersive tendencies of distributions. Without more refined information regarding the distributional characteristics of perchloroethylene emissions, we cannot determine specifically the factors contributing to the observed distributional patterns. 5. Population Exposure Estimates Mean population exposure estimates were calculated using the study period perchloroethylene data. Exposures for the South Coast Air Basin and San Francisco Bay Area Air Basin were estimated by interpolating station values to census tract centroids. For the other air basins, a basin-wide mean concentration was estimated from the means for all sites in the basin. It was then assumed that all people in those basins with a sampling site (except Sacramento Valley Air Basin) were exposed to this estimated basin-wide mean concentration. Population data used in the exposure analysis represent 1980 census data updated to 1985 levels. The results of the exposure analysis, which are discussed below, are summarized in Table IV-5. Table IV-6 presents approximate 95 percent Bootstrap Confidence Interval bounds (Efron, 1982) for each mean annual exposure estimate. The overall statewide perchloroethylene exposure, weighted by population, is estimated to be 0.37 ppb. A total of 20,339,250 people are estimated to reside in the study areas, representing approximately 2. The Coefficient of Variation is the standard deviation divided by the mean, then multiplied by 100 so as to be expressed as a percentage. A-29 DRAFT Table IV-4 Summary of Perchloroethylene Peak-to-Mean Ratios and Coefficients of Variations fC.YJ (ppb) AIR BASIN Site Location Peak Cone. Mean3 Peak-to-Mean*3 C.V. Cone. Ratio South Coast Air Basin Long Beach 1.60 Los Angeles 1.90 Rubidoux 0.52 Upland 1.10 0.42 0.64 0.26 0.46 South Central Coast Air Basin Santa Barbara 0.33 0.15 Simi Valley 4.80 0.44 San Diego Air Basin Chula Vista 0.69 El Cajon 1.10 0.27 0.42 San Francisco Bay Area Air Basin Concord 3.30 0.75 Fremont 0.73 0.29 Richmond 0.41 0.11 San Francisco 0.97 0.21 San Jose 0.87 0.23 San Joaquin Valley Air Basin Bakersfield 0.18 Fresno 0.68 Merced 0.39 Modesto 0.83 Stockton 0.26 0.10 0.18 0.18 0.16 0.12 Sacramento Valley Air Basin Citrus Heights 0.21 0.10 3.85 2.97 1.97 2.39 2.24 10.82 2.55 2.62 4.42 2.54 3.73 4.63 3.79 1.79 3.72 2.14 5.05 2.25 2.20 86.7 71.4 52.9 51.6 52.6 216.0 60.0 71.0 99.8 63.2 71.4 101.2 94.7 41.7 95.3 55.0 107.8 47.1 51.6 a. Means are calculated as the sum of all observations for each site divided by the total number of observations at each site. b. Peak-to-Mean Ratios are calculated as the largest value reported for each site divided by the average of all observations reported. c. The Coefficient of Variation is the standard deviation divided by the mean, then multiplied by 100 so as to be expressed as a percentage A-30 SL 038821 Air Basin Table IV-5 DRAFT Mean Perchloroethylene Exposure Estimates July 1988 through June 1989 (PPb) Estimated Mean Population South Coast Q.44a 10,092,133 South Central Coast 0.28b 925,822 San Diego 0.36b 2,135,872 San Francisco Bay Area 0.25a 4,394,374 San Joaquin Valley 0.15b 1,901,243 Sacramento Valley0 OrH o 889,806 Overall Population-Weighted Exposure 0.37 20,339,250 a. Population-weighted exposure estimate b. Mean for all sites within the basin c. Exposure estimates are for Sacramento County residents only. A-31 Table IV-6 DRAFT Lower Bound, Mean and Upper Bound Perchloroethylene Concentrations for July 1988 through June 1989a (PPb) Air Basin Site Location Lower Bound Site Mean Upper Bound South Coast Air Basin Long Beach Los Angeles Rubidoux Upland 0.29 0.40 0.22 0.37 South Central Coast Air Basin Santa Barbara 0.12 Simi Valley 0.20 San Diego Air Basin Chula Vista El Cajon 0.22 0.30 San Francisco Bay Area Air Basin Concord 0.4S Fremont 0.21 Richmond 0.08 San Francisco 0.13 San Joseb 0.15 San Joaquin Valley Air Basin Bakersfield 0.08 Fresno 0.12 Merced 0.16 Modesto 0.10 Stockton 0.09 Sacramento Valley Air Basin Citrus Heights 0.07 0.42 0.60 0.26 0.48 0.16 0.39 0.30 0.42 0.70 0.27 0.11 0.20 0.23 0.10 0.18 0.19 0.16 0.12 0.10 0.64 0.84 0.31 0.62 0.19 0.62 0.40 0.57 1.04 0.34 0.14 0.30 0.34 0.12 0.27 0.23 0.22 0.15 0.13 a. Upper and lower bounds represent 95 percent Bootstrap Confidence Interval bounds (Efron, 1982) for each mean annual exposure estimate. b. One value below the LOD was reported for the San Jose site. The single observation was estimated to be .006 ppb. Using zero or the LOD as a replacement value, after rounding to two decimal places, gave the same annual mean exposure estimate for San Jose. A-32 SL 038823 DRAFT 80 percent of the state's population. Basin-specific, populationweighted mean concentrations vary from a minimum of 0.15 ppb in the San Joaquin Valley Air Basin (excluding Sacramento County's 0.10 ppb) to a maximum of 0.44 ppb in the South Coast Air Basin. Figure IV-3 shows the total number of people exposed to various mean annual perchloroethylene concentrations (rounded off to the nearest 0.05 ppb). Figure IV-4 represents the same data as in Figure IV-3, but plotted as the cumulative population exposed to an estimated mean perchloroethylene concentration. The overall geographic mean perchloroethylene concentration was 0.29 ppb. This value is approximately 22 percent lower than the population-weighted exposure estimate of 0.37 ppb and indicates that the highest concentrations of perchloroethylene tend to be in the areas of higher population density. C. EXPOSURE TO PERCHLOROETHYLENE NEAR EMISSION SOURCES To assess the impact of perchloroethylene emission sources on nearby population, the ARB used emissions information in conjunction with meteorological data to estimate population exposure in census tracts surrounding eight facilities that emit perchloroethylene. The following text is a discussion of the data and modeling methodology used to estimate population along with the exposure results. The concentrations presented in this section do not represent the total ambient air exposure of the people living in the vicinity of the modeled sources. The concentrations reported are those which would exist if the modeled facility was the only perchloroethylene emission source affecting the nearby population. In reality, there are other emission sources, large and small, located in the basin which contribute to the total perchloroethylene concentration that the modeled population is exposed to. Two 41 kilometer grids with 1 kilometer receptor spacing were utilized to model emissions from eight facilities in the South Coast Air Basin. One grid contains only those facilities near City of Industry and the other grid contains only those facilities near Burbank. The emission sources were approximately centered on each grid and the two grids only slightly overlap. For each grid, the ISCST air quality model was used to predict above ambient annual average perchloroethylene concentrations. Meteorological data input into the model were obtained from the nearest available meteorological station to each facility. Meteorological data for the City of Industry case were obtained from Los Angeles International Airport records for 1978. Meteorological data for the Burbank case were obtained from Burbank Airport records for 1964. These years of meteorological data were selected because they were the most recent years available and they represent poor years in terms of pollutant dispersion in the South Coast Air Basin. SL 038824 A-33 P o p u la tio n in M illio n s DRAFT Figure IV-3 ESTIMATED MEAN ANNUAL PERCHLOROETHYLENE EXPOSURE BASED ON JULY 1S08-JUNE 1980 DATA ARB/TSD:September 6 8 8 C0 sI XSi V* aw CotOOos ' Figure IV-4 ESTIMATED CUMULATIVE PERCHLOROETHYLENE EXPOSURE BASED ON JULY 1988-JUNE 1989 DATA DRAFT ARB/TSD: October 89 Estimates of cumulative impacts from the Burbank and the City of Industry grids are not provided in this analysis. Results indicate that perchloroethylene impacts from the modeled facilities are minimal on the edges of both grids. Thus, the cumulative impacts are not different from these results. The modeling results were used to calculate the maximum annual average exposure. For City of Industry, over 2,000 people are estimated to be exposed to an annual average concentration of 6 ppb perchloroethylene. This exposure occurred at a receptor approximately two kilometers northeast of one of the modeled facilities. For Burbank, over 600 people are estimated to be exposed to an annual average concentration of 3 ppb perchloroethylene. The estimated annual average exposure (weighted by population) to perchloroethylene for people living near City of Industry is 0.07 ppb for a population of approximately 2,691,000 people. For Burbank, the estimated annual average exposure (weighted by population) to perchloroethylene is approximately 0.03 ppb for a population of approximately 3,228,000 people. 0. INDOOR AIR EXPOSURE TO PERCHLOROETHYLENE The best data for estimating indoor air exposure for perchloroethylene come from the Total Exposure Assessment Methodology (TEAM) studies conducted by the Environmental Protection Agency (EPA) during 1980-85 (referred to as TEAM 80/84 in this report; Wallace, 1987; U.S. EPA 1987a,b) and by the EPA and ARB during 1987 (referred to as TEAM 87 in this report; Pellizzari et al., 1989). The TEAM 80/84 studies provided personal sampling data whereas TEAM 87 provided indoor fixed-site sampling data as well. Limited indoor concentration data are also available from a few commercial building studies and from a few European studies. 1. Personal Air Sampling Data The TEAM 80/84 studies monitored a total of 240 participants in the Los Angeles and Contra Costa areas. Each participant carried a personal air sampler and was monitored for two consecutive 12-hour sampling periods. In the Los Angeles area, field studies were conducted in February 1984 and May 1984. In the February study, the average (arithmetic mean) 24-hour personal air concentration of 3 perchloroethylene was 16 ug/m (2.40 ppb). In the May study, the average 24-hour personal air concentration of perchloroethylene was 15 ug/m^ (2.25 ppb). The study in the Contra Costa area was conducted in June, 1984, and the average 24-hour personal air concentration was 5.6 ug/m (0.84 ppb). In Table IV-7, the median and maximum concentrations of the overnight (12-hour) personal air samples are presented along with the median and maximum concentrations of concurrent and colocated outdoor ambient air samples (Pellizzari, et al., 1986). The overnight A-36 SL 038827 DRAFT Table IV-7 Overnight Indoor and Outdoor Air Levels of Perchloroethvlene in California 16:00 om to 6:00 ami (PPb)3 Number of Malched Samples Indoorb Outdoor Los Angeles (Feb./84) Los Angeles (May/84) Contra Costa (June/84) 25 23 10 Median Maximum 1.34 14.1 0.26 8.4 0.32 1.32 Median 1.1 Maximum 5.1 0.2 0.72 0.04 0.84 a. Original data reported in ug/m , conversion factor: 1 ug/m =0.15 ppb b. Overnight personal sampler data Source: Pellizzari et al., 1986 SL 038828 A-37 DRAFT personal sampler data provide a good estimate of indoor air exposure at home since, during this sampling period (6:00 pm - 6:00 am), most participants remained in their homes. The data indicate that median indoor air levels are generally higher than median outdoor air levels. In addition, the data indicate that extremely high levels of perchloroethylene could be found in some homes. The TEAM 80/84 studies also included other geographical areas besides California. For example, studies were conducted during three seasons in New Jersey. The weighted average overnight air exposure for O the three seasons were 11, 9.0 and 13 ug/m (1.65, 1.35 and 1.95 ppb), respectively. The corresponding average outdoor air concentrations were 3.7, 4.0, and 1.9 ug/m3 (0.56, 0.60, and 0.29 ppb), respectively (Wallace, 1987). The TEAM 87 study (Pellizzari et al., 1989) was designed as a follow up study of the TEAM 84 study for California. The same Los Angeles areas studied in 1984 were revisited and some of the original participants were included in the TEAM 87 study. A total of 51 and 43 persons participated in the January, 1987 and June, 1987 field studies, respectively. In the January study, the average (arithmetic mean) 24hour personal air concentration of perchloroethylene was approximately 12 ug/m (1.80 ppb). In the June study, the average 24-hoour personal air concentration of perchloroethylene was about 13 ug/m (1.95 ppb). These two-season concentrations are very similar to those obtained in the TEAM 84 study for the Los Angeles areas. 2. Fixed-Site Air Sampling Data The TEAM 87 study also provides fixed-site monitoring of indoor air concentrations of perchloroethylene (Pellizzari et al., 1989). Air sampling devices were placed in two indoor locations (kitchen and living room) of the homes belonging to people who participated in the personal sampling program. Air samples were collected during the same two seasons. The 24-hour average concentrations of perchloroethylene in the kitchen area were 6.72 ug/m (1.01 ppb) in winter and 2.27 ug/m (0.34 ppb) in summer. The 12-hour average concentrations of 3 perchloroethylene in the living area were 5.27 ug/m (0.79 ppb) in 3 winter and 2.75 ug/m (0.41 ppb) in summer. The 24-hour outdoor average concentrations were 4.41 ug/m3 (0.66 ppb) in winter and 1.74 ug/m3 (0.26 ppb) in summer. Table IV-8 supports the conclusion that indoor air concentrations of perchloroethylene were generally higher than outdoor concentrations during the two sampling seasons. It also indicates that indoor and outdoor air concentrations of perchloroethylene were generally higher during winter than summer months. The average indoor concentrations of the kitchens and the living areas were very similar. A-38 SL 038829 DRAFT Table IV-8 24-Hour Average Concentrations of Perchloroethylene in Indoor and Outdoor Microenvironments (PPb)a Winter Mean Median Summer Mean Median INDOORS Kitchen Living Areab 1.01 0.66 0.79 0.66 0.34 0.32 0.41 0.38 OUTDOORS 0.66 0.53 0.26 0.21 a. Original data reported in ug/m3 ; conversion factor: 3 1 ug/m = 0.15 ppb b. 12-hour daytime value only. Source: Pellizzari et al., 1989 SL 038830 A-39 DRAFT In studies of indoor air quality in 10 public-access buildings, Wallace et al. (1987a) reported that perchloroethylene was one of the 24 compounds most frequently found out of a total of over 200 identified organic compounds. The 3-day concentration means of perchloroethylene 3 in three newly completed buildings ranged from undetectable to 7 ug/m (1.05 ppb) while the 3-day concentration means for seven old buildings ranged from 1 to 6 ug/m'* (0.15 to 0.90 ppb) (Sheldon et al., 1988). Data are also available from several European studies. Although such data may not be representative of California indoor concentrations due to differences in consumer products and life styles, the European data nonetheless confirm the general presence of measurable levels of perchloroethylene in indoor environments. In a study of 14 homes and one office building in Italy, De Bortoli, et al. (1986) detected indoor 3 perchloroethylene concentrations ranging from 3 to 47 ug/m (0.457.05 ppb). Lebret, et al. (1986) reported that 30 percent of 300 Dutch homes sampled had indoor air levels of perchloroethylene greater than 3 the detectable limit of 2 ug/m (0.30 ppb). However, the median levels of perchloroethylene sampled in different age-group homes were generally below 2 ug/m3 (0.30 ppb). Krause, et al. (1987) have also reported preliminary results from a 500-home study in Germany. The observed 3 indoor levels of perchloroethylene ranged from less than 1.0 ug/m and 3 up to 617 ug/m (0.15 to 92.55 ppb) with an average (geometric mean) 3 concentration of 12 ug/m (1.8 ppb). 3. Summary of Indoor and Personal Air Exposure Perchloroethylene is an ubiquitous indoor air pollutant. The indoor air concentrations of perchloroethylene are generally higher than outdoor air concentrations. TEAM studies conducted in California provide the most representative sampling data for assessing Californians' indoor air exposures. Based on the TEAM 87 data, average 3 residential indoor air concentrations range from 2.75 to 6.72 ug/m (0.41 to 1.01 ppb). Based on personal air sampling data from the TEAM 87 and TEAM 80/84 data, Californians' average personal air exposure to 3 perchloroethylene ranges from 5.6 to 16 ug/m (0.84 to 2.40 ppb). 4. Contribution to Total Exposure Section 39660.5(d) of the Health and Safety Code states "the state board shall identify the relative contribution to total exposure to the contaminant from indoor concentrations, taking into account both ambient and indoor air environments." Based upon available data and assuming Californians spend 80 to 90 percent of their time indoors (Robinson, 1977), indoor inhalation may be the major route of exposure to perchloroethylene. A comparison of simultaneous indoor and outdoor perchloroethylene concentrations was conducted as part of to.- TEAM 80/84 A-40 SI 38832 DRAFT study. The results of this comparison (Table IV-7) indicate that indoor perchloroethylene concentrations can be more than 11 times greater than outdoor concentrations. However, this may not always be the case since indoor concentrations of perchloroethylene are dependent upon the use of dry cleaners and consumer products containing perchloroethylene (see Chapter III., Section D. Potential Sources of Indoor Perchloroethylene). E. OTHER ROUTES OF PERCHLOROETHYLENE EXPOSURE 1. Water Ingestion The major source of drinking water for California is surface water which does not have detectable perchloroethylene concentrations. The detectable limit of perchloroethylene in water is 0.5 ug/1 (0.5 ppb). Although perchloroethylene is the most frequently found contaminant in ground water, only six percent of the wells for large public water systems contain perchloroethylene levels above 0.5 ug/1. This yields a potential exposed population of about 461,000 persons (CDHS, 1986). For small public water systems, less than 0.6 percent of the supplying wells have detectable perchloroethylene levels (Spath, 1987). Information on private well contamination is very limited. In a limited testing of suspected private wells in Santa Clara County, about eight percent were contaminated with organic chemicals (Hinman et al., 1986). This percentage represents a high estimate since suspected rather than random wells were chosen for testing. In addition, the number of persons supplied by private wells is much less than those supplied by public water systems. Based on this information, perchloroethylene exposure through drinking water is judged to be minimal ordinarily. 2* Esod Ingestion Perchloroethylene is not one of the compounds that have been measured routinely in U.S. food and food products. Using a new technique for V0C analysis in food, Entz and Hollifield (1982) detected low levels of perchloroethylene in different fish from various U.S. waterways and in a variety of jellies and sauces from a food processor in Pennsylvania. However, these data do not provide quantitative estimates of perchloroethylene in food. A more comprehensive analysis of different food groups was conducted in 1975 by British researchers. They reported that perchloroethylene levels were generally low, with highest levels in margarine and olive oil at 7 ug/kg (McConnell et al., 1975). Based on these British data, Gilbert et al. (1982) estimated that perchloroethylene exposure in the United States via food ingestion was 1.2 ug/day. SL 038832 A-41 DRAFT The range of total daily intake of perchloroethylene via food consumption in Europe was estimated at 87.4 to 160 ug. These estimates were the results from market basket surveys in European countries as summarized in a Dutch criteria document (Ministerie van Volkshuisvesting, 1984). These high estimates of perchloroethylene in food may not be applicable to California since food sources and food consumption patterns are different between Europeans and Californians. F. ESTIMATES OF TOTAL EXPOSURE FROM INDOOR AIR AND OTHER ROUTES Estimates of the presented daily dose of perchloroethylene from different environmental media are provided in Table IV-9. The presented dose represents an amount of a chemical presented to an individual (host) without consideration of any subsequent biological interactions between the host and the chemical. From the Table, indoor air exposure appears to be the most significant medium of exposure to perchloroethylene. 1. Indoor Air Data from the 24-hour air monitoring in the kitchen area were used to calculate the presented daily dose from residential indoor air exposure. In addition, 24-hour personal air monitoring data were used to estimate the total air exposures. Total air exposures include exposures through residential and non-residential indoor air, plus outdoor air. 2. Food The estimated daily dose of perchloroethylene through food ingestion is highly uncertain since no California or U.S. data are available. In addition, European data are themselves not consistent and are not considered likely to represent Californians' food consumption habits. Therefore, the exposure,estimate through food ingestion may be high; the relevance of the European food market survey data is unknown. 3. Drinking Water The relative contribution of drinking water to daily exposures of perchloroethylene appears to be minimal. The quality of data used for making the estimates is relatively good. SL 038833 A-42 DRAFT Table IV-9 Estimated Doses of Perchloroethylene from Exposure through Different Media Media Average Presented Daily Dose (inhaled or ingested! References AIR Residential Indoor Air Personal air data 90 uga 246 ugb EQflB British data 1.2 ug European food market survey (87-160 ug) WATER-F0R DRINKING PURPOSES Surface Water Ground Water Supply Large public water system Small public water system 1 ug 0.6 ug 1 ug Pel 1 izzari et al., 1989 Pel 1izzari et al., 1989; U.S. EPA, 1987b Gilbert et al., 1982; McConnell et al., 1975 Ministerie van Volkhuisvesting, 1984 CDHS, 1986; Spath, 1987 CDHS, 1986; Spath. 1987 CDHS, 1986; Spath, 1987 Calculations: (Assume the average person inhales 20 m of air daily) a. Let C1= the average (arithmetic mean) of 6.72 and 2.27 ug/m3 the means of 24-hour kitchen air concentrations from the TEAM 87 winter and summer studies. When Cl = 4.5 ug/m3, presented daily dose = Cl x 20 m3 = 90 ug b. Let C2= the average (arithmetic mean) of 16, 15, 13, 12 and 5.6 ug/m? the means of 24-hour personal air concentrations from the TEAM 87 and TEAM 80/84 summer and winter studies. Personal air exposures include exposures through residential and non-residentia 1 indoor air, plus outdoor air. When C2 = 12.32 ug/m3, presented daily dose = C2 x 20 m3 = 246 ug 3S34 A-43 References for Chanter IV DRAFT CDHS, 1986. Calif. Dept, of Health Services. Final Report on a Monitoring Program for Organic Chemical Contamination of Large Public Water Systems in California. Conover, W. J., 1971. "Practical Nonparametric Statistics," John Wiley and Sons Inc., pp. 256-263 De Bortoli, M. et al., 1986. "Concentrations of Selected Organic Pollutants in Indoor and Outdoor Air in Northern Italy." Environ. International 12:343-350. Efron, B., 1982. "The Jackknife, The Bootstrap and Other Sampling Resampling Plans," SIAM Monograph No. 38 Entz, R. C. ; H. C. Hollifield, 1982. "Headspace Gas Chromatographic Analysis of Foods for Volatile Halocarbons." J. Agric. Food Chem. 30:84-88. Gilbert, D. et al., 1982. "Exposure and Risk Assessment for Tetrachloroethylene." U.S. Environmental Protection Agency Report. (EPA/400/4-85/015) Gleit, A., 1985. "Estimation for Small Normal Data Sets with Detection Limits," Environ. Sci. Technol., li (12); Page 1201 Hinman, K.; D. Schwartz and E. Soffer, 1986. "Santa Clara Valley Integrated Environmental Management Project." Revised Stage I Report. U.S. Environmental Protection Agency Report. Krause, C. et al., 1987. "Occurrence of Volatile Organic Compounds in the Air of 500 Homes in the Federal Republic of Germany." In: Indoor Air '87--Proceedings of the 4th International Conference on Indoor Air and Climate, Berlin (West), 17-21 August 1987. Volume 1, pp.102-106. Lebret, E. et al., 1986. "Volatile Organic Compounds in Dutch Homes." Environ. International 12:323-332. McConnell, G.; D. M. Ferguson and C. R. Pearson, 1975. "Chlorinated Hydrocarbons and the Environment." Endeavour 34(121):13-18. Ministeri# van Volkshuisvesting, 1984. "Criteriadocument over Tetrachlooretheen." Publikatlereeks Lucht 32. Pellizzari, E. D. et al., 1989. "Development and Implementation of Exposure Assessment Procedures for Toxic Air Pollutants in Several Los Angeles County, CA Communities." Final Report. ARB Contract no. A5-174-33. A-44 SV DRAFT Pellizzari, E. D. et al., 1986. "Comparison of Indoor and Outdoor Residential Levels of Volatile Organic Chemicals in Five U.S. Geographical Areas." Environ. International 12:619-623. Robinson, J. P., 1977. "How Americans Use Time: A Social-Psychological Analysis of Everyday Behavior," Praeger, New York. SAS Institute Inc., 1985. SAS/Stat Guide for Personal Computers, Version 6 pp. 221 - 227 Shapero, S. S. and M. B. Wilk, 1965. "An Analysis of Variance Test for Normality (Complete Samples)," Biometrica, Page 591 Sheldon, L. et al., 1988. "Indoor Air Quality in Public Buildings," Volume I and II. U.S. Environmental Protection Agency Report. (EPA/600/6-88/009a,b). Spath D., 1987. Personal correspondence with David Spath, staff of California Department of Health Services, (10/8/87 Memo). U.S. EPA, 1987a. "The Total Exposure Assessment Methodology (TEAM) Study: Elizabeth and Bayonne, New Jersey, Devils Lake, North Dakota and Greensboro, North Carolina: Vol. II Parts 1 and 2." U.S. Environmental Protection Agency Report. (EPA/600/6-87/002b) U.S. EPA, 1987b. "The Total Exposure Assessment Methodology (TEAM) Study: Selected Communities in Northern and Southern California: Vol. III." U.S. Environmental Protection Agency Report. (EPA/600/6-87/002c) Wallace, L. A., 1987. "The Total Exposure Assessment Methodology (TEAM) Study: Summary and Analysis: Vol. I." U.S. Environmental Protection Agency Report. (EPA/600/6-87/002a) Wallace, L. A. et al., 1987a. "Volatile Organic Chemicals in 10 PublicAccess Buildings." In: Indoor Air '87--Proceedings of the 4th International Conference on Indoor Air and Climate, Berlin (West), 17-21 August 1987. Volume 1, pp.188-192. St 038 836 A-45 DRAFT V. ATMOSPHERIC PERSISTENCE AND FATE OF PERCHLOROETHYLENE The atmospheric persistence of a pollutant is its tendency to remain in the atmosphere in its original form. In analyzing human exposure to perchloroethylene, its persistence is important for two reasons: 1) if the removal time is long compared to the time needed to advect (disperse by wind) the pollutant across an air basin, the concentration throughout the basin can be inferred from measurements at specific locations; and 2) if attenuation of concentration in the plume from a source occurs mostly by dispersion of the plume rather than by chemical or physical removal (i.e. if the pollutant is persistent), routine modeling procedures like Gaussian modeling can be used to estimate the local effect of the source. If removal is fast (i.e. if the pollutant is not persistent), much more complicated modeling may te needed to estimate local effects. A. ATMOSPHERIC PERSISTENCE OF PERCHLOROETHYLENE There are chemical and physical mechanisms that operate to remove pollutants from the troposphere (the lower part of our atmosphere). These mechanisms include: photolysis (degradation by solar radiation), photo-oxidation (reactions with reactive species found in polluted atmospheres), adsorption on particles that fall out of the air (dry deposition), and wet deposition from interaction with fog or rain (wash out). Chemical mechanisms appear to be the dominant force responsible for removing perchloroethylene from the atmosphere. Two commonly used measures of persistence are half-life (t,,2) and lifetime (r). Half-life is defined as the time required for the concentration of a reactant to fall to one-half of its initial value, whereas lifetime is defined as the time it takes for the reactant concentration to fall to 1/e of its initial value (where e * 2.718), or approximately 37 percent of its original value (Finlayson-Pitts and Pitts, 1986). For the rest of this discussion we will describe the persistence of perchloroethylene in terms of lifetime. A-46 DRAFT 1. Chemical Removal Mechanisms The atmospheric lifetime of perchloroethylene is related to the rate constant for the photo-chemical reactions that occur and the concentration of the reactants involved in the atmospheric reactions. For a second order reaction of perchloroethylene with atmospheric oxidants (e.g. hydroxyl radical, ozone, or N03 radical), the following relationship holds: perchloroethylene + B --k-2-> products, then r - 1/ (k^ [B]) (where k2 is the 2nd order reaction rate constant, and [B] is the concentration of the atmospheric oxidant). Of the chlorinated ethenes, perchloroethylene is the least reactive to electrophilic attack. This reduced reactivity is believed to be a consequence of the electron-inductive effect of the four chlorine atoms, which reduce the electron density about the double bond and, by doing so, reduce the reactivity of the double bond (U.S. EPA, 1985). The four chlorine atoms also provide steric protection to the double bond which would also decrease the reactivity of the double bond. However, Howard (1976) reports that the effect of the four chlorine atoms may not completely account for the unusually low reactivity of perchloroethylene to electrophilic attack. Halogenated ethenes such as C2F3C1, C2FC13, and sterically hindered ethenes such as tetramethylethene do not demonstrate the dramatic decrease in reactivity relative to other ethene derivatives as does perchloroethylene. Howard speculates that the bonding in (^Cl^ may differ from other ethenes. The electrons from the chlorine atoms In this planar symmetrical molecule could be involved in the pi electron system of the double bond, which could have a major effect on the nature of the bonding and the reactivity of the molecule. There are three principal photo-chemical reactions that can affect the atmospheric persistence of perchloroethylene. These are: 1) attack during daylight hours by hydroxyl radicals (OH radicals); 2) attack at night by NO^ radicals; and 3) attack by ozone (0^) (Finlayson-Pitts and Pitts, 1986). SL 038838 A-47 DRAFT a. Reaction with QH Radicals The principal mechanism for perchloroethylene removal from the atmosphere appears to involve reaction with OH radicals. The lifetime of perchloroethylene as a consequence of this reaction (tq^) is inversely proportional to the atmospheric concentration of OH radicals and the reaction rate constant, kOH The reaction can be expressed as follows: tqh - 1/ kH COH] or (k0H [OH])-1 Atkinson (1986) reviewed the work of several investigators studying the kinetics and mechanics of OH radical reactions. Based on this work Atkinson produced a formula for calculating the kOH for perchloroethylene: /C0H . 9.64 x ID'12 -<1209)/! (where T is the absolute temperature in K). Using this formula, Atkinson estimated the rate constant for the reaction of perchloroethylene with OH radicals to be 1.67 x 10 -13 cm3 molecule-1 second-1 at 298K. Using a comparable rate constant of 1.7 x 10-13 cm3 molecule-1 second-1 at 300K, Singh et al. (L9dl) calculates 1 lifetime for perchloroethylene in the troposphere of 68 days. This calculation was based on 12 hours per day of sunlight and a 24-hour hydroxyl radical concentration of 1 x 10^ molecules per cubic centimeter, which has been reported as a reasonable estimate (Cupitt, 1983). The OH radical concentration may be somewhat lower during winter months. By using a seasonally averaged OH radical concentration of 4 x e 10 molecules per cubic centimeter one can calculate a lifetime of 292 days. Conversely, OH radical concentrations can be somewhat higher in heavily polluted atmospheres (i.e. higher criteria pollutant concentrations that are involved in OH radical formation). These results indicate that, depending on atmospheric conditions, the lifetime of perchloroethylene, as a result of its reaction with OH radicals, may range from slightly greater than two months to almost one year. b. Reactions with Nitrate Radicals and Ozone The lifetime of perchloroethylene as a result of its reaction with nitrate radicals and ozone can be determined in the same manner as was done above with OH radicals, r = 1/ (kg [B]). Based on a nitrate A-48 SL 038839 DRAFT reaction rate with perchloroethylene of <6 x 10"*^ cm"* molecule-* sec"* and a 12 hour nighttime nitrite radical concentration of 10 ppt (2.4 x o3 10 molecule/cm ), the calculated lifetime of perchloroethylene with respect to reaction with nitrate radicals is >4.4 years (Atkinson, 1989). Based on an ozone reaction rate with perchloroethylene of <2 x 10 -23 cm3 molecule -1 sec -1 (Mathias et a!,f 1974) and a tropospheric ozone concentration of 1 x 10 12 molecule/cm3 (Cupitt, 1980), the lifetime of perchloroethylene as a consequence of its reaction with ozone is >1000 years (Atkinson, 1989). Both nitrate radicals and ozone chemical reaction removal processes are too long to compete with the OH radical reaction. c. Other Reactions Several chamber studies indicate perchloroethylene is more reactive than expected from calculations of its reactions with OH radicals (Dimitriades, et al., 1983). This has been studied by researchers interested in perchloroethylene1s contribution to ozone/oxidant problems in the urban atmosphere. Dimitriades argues that the smog chamber reactions of perchloroethylene are dominated by chlorine atom substitution rather than OH radical attachment and that the chlorine atom reactions are the reason for the increased reactivity of perchloroethylene in the smog chambers. Although this reaction occurs under laboratory conditions, Dimitriades contends that this reaction does not occur at a high enough rate in the atmosphere to affect the reactivity of perchloroethylene. Other hydrocarbons present in the atmosphere will react much more rapidly with available chlorine atoms, effectively scavenging the atoms and preventing chlorine atom-initiated photooxidation from being a major degradation process. In addition, in real urban atmospheres, other organic gases are present at concentrations several hundred times as high as perchloroethylene. The source of the chlorine atoms in the chamber studies was not elucidated by Dimitriades. 2. Physical Removal Mechanisms We have not found estimates of the rates for the physical removal of perchloroethylene from the atmosphere. However, Cupitt (1980), has estimated the lifetime of ethylene dichloride under conditions of removal by rain washout, dry deposition, and adsorption on aerosols (that fall out) as 390 years, 13 years, and 25 years, respectively. Physical removal is dependent on several physical properties of a substance, including the polarity (dipole moment), solubility in water, adsorptivity on particles (e.g., on carbon), and its vapor pressure. Since perchloroethylene is comparable to ethylene dichloride in all of these properties, we have assumed that perchloroethylene has similarly SL 38840 A-49 DRAFT long removal times for these removal processes. Therefore, chemical removal mechanisms will be the predominant factors influencing the persistence and fate of perchloroethylene. B. FATE OF PERCHLOROETHYLENE IN THE ATMOSPHERE The reaction of perchloroethylene with OH radicals is predicted to occur as shown below and illustrated in Figure V-l (Graedel, 1978). C12C:CC12 + OH.......... > H0CC12CC12 H0CC12CC12 + 02-........ > H0CC12CC1202 H0CC12CC1202+ NO.......... > H0CC12CC120 + N02 HOCC12CC120 ............ > HOCC12 + C0C12 H0CC12 + 02.......... > C0C12 + H02 Lillian (1975) predicted that perchloroethylene should lead to the formation of large quantities of phosgene in the atmosphere. Singh (1976) investigated the environmental significance of the production of phosgene from perchloroethylene in the atmosphere. Singh estimated that the photooxidation of perchloroethylene could result In phosgene levels in the low ppb range in urban areas under adverse meteorological conditions. The low reactivity of perchloroethylene determined in the more recent smog chamber studies by Dimitriades (1983) suggest that only trace levels of phosgene would be expected to be formed (U.S. EPA, 1985). In addition, the photooxidation of perchloroethylene is believed to lead to the production of other potentially toxic compounds in the atmosphere. Singh (1977) predicted that the major environmental impact of chloroethenes in general, and perchloroethylene in particular, is likely to be the decomposition of these compounds into highly toxic species. Chamber studies conducted by Gay (1976) found that the chlorinated photooxidation products of perchloroethylene are hydrogen chloride, phosgene, trichloroacetyl chloride. On the other hand, no phosgene was detected from perchloroethylene degradation during smog chamber studies conducted by Dimitriades in 1983. Trichloroacetyl chloride can further decompose to form carbon tetrachloride (Singh, 1977). Singh estimated that as much as eight percent by weight of the atmospheric perchloroethylene could eventually be converted into carbon tetrachloride. However, this reaction is believed to occur through a chlorine atom substitution process that is not likely to occur at a substantial rate in the atmosphere as opposed to a smog chamber. This chlorine substitution reaction is illustrated in Figure V-l. A-50 A & DRAFT Finally, a possible reaction pathway for the atmospheric oxidation of perchloroethylene to oxalyl chloride has been suggested and is presented below (Howard, 1976): OH C12C:CC12........... > CClgCClgOH cci2cci2oh + o2. . . . . . . . > o2cci2cci2oh 02CC12CC120H NO........... > COCICC120H + N02 + Cl OH COCICC120H ............. > C0C1C0C1 H20 Cl SL 038842 A-51 DRAFT FIGURE V-1 FATE OF PERCHLOROETHYLENE Reactions with OH Radicals ci ci OH Cl Cl Perchloroethylene OH 1 Cl --C --C--ci 1t 11 Cl Cl OH-Chloroalkene adduct (1) o2 Cl -- C--CI phosgene (3) + HO. OH 0 u2 1 II c --Cl + Cl --C --Cl 1 phosgene (3) Cl O \ O OH 11 Cl-- c--c -- 1l Cl Cl NO > NO ' o OH 11 Cl -- c--C -- 11 Cl Cl HydroxyChloroAlkoxy Radical (2) Reaction Products via Chlorine Substitution Cl ci \/ c=c Cl/ \ Cl Perchloroethylene + Oi Cl ROB/11/B9 o \ Cl 1 -- c1 --Cl 1 Cl NO NOz 0 - 1o -- o Cl 1 1 -C --Cl I Cl C11 l A-52 11 0 Cl II 1 Cl + ci --c--c --Cl 11 Cl trichloroacetylchloride SL 0388A3 % % References for Chapter V DRAFT Atkinson, R., 1986. "Kinetics and Mechanisms of the Gas-Phase Reactions of the Hydroxyl Radical with Organic Compounds Under Atmospheric Conditions." Chemical Reviews. 86:69-201. Atkinson, R., 1989. Memo to Richard Corey of the California Air Resources Board regarding the atmospheric chemistry of vinyl chloride, trichloroethylene, and perchloroethylene. July 20, 1989 Barbari, T. A.; C. J. King, 1982. "Equilibrium Distribution Coefficients for Extraction of Chlorinated Hydrocarbons and Aromatics from Water into Undecane." Environmental Science & Technology. 16(9):624- Cupitt, L. T., 1980. "Fate of Toxic and Hazardous Materials in the Air Environment." EPA-600/3-80-084, PB 80-221948. Cupitt, L. T., 1983. Personal communication between Dr. L.T. Cupitt of EPA's ESRL and Dr. D.A. Gray of Syracuse Research Corp., March 1983. Dimitriades, B. et al., 1983. "Photochemical Reactivity of Perchloroethylene: A Hew Appraisal." Journal of the Air Pollution Control Association, 33: 375. Finlayson-Pitts, B. J., and J. N. Pitts Jr., 1986. "Atmospheric Chemistry: Fundementals and Experimental Techniques." John Wiley & Sons (pub). Gay, B. W. et al., 1976. "Atmospheric Oxidation of Chlorinated Ethylenes." Environmental Science and Technology, 10: 58. Graedel, T. E., 1978. Personal communication between Dr. Graedel, Bell Laboratories, and H. B. Singh, SRI, Oct. 1978. Howard, C. J., 1976. "Rate Constants for the Gas-Phase Reactions of OH Radicals with Ethylene and Halogenated Ethylene Compounds." Journal of Chemical Physics, 65: 148. Lillian, E. et al., 1975. "Atmospheric Fates of Halogenated Compounds." Environmental Science and Technology, 9: 1042. Mathias, et al., 1974. Canadian Journal of Chemistry. 52, 3852, 1974. Merck Index, 1983, 10th Edition, Merck and Co., Inc., Rahway, New Jersey. Singh, H. B. et al., 1976. "Phosgene in the Ambient Air." Nature, 264: 428. Singh, H. B. et al., 1977. "Urban-Nonurban Relationships of Halocarbons, SFg, N2O and Other Atmospheric Trace Constituents." Atmospheric Environment, 11: 819. SL 388AA A-53 draft Singh, H. B. et al., 1981. "Measurements of Some Potentially Hazardous Organic Chemicals in Urban Environments." Atmospheric Environment, 15: 601. U.S. EPA, 1985. "Health Assessment Document for Tetrachloroethylene - Final Report." EPA-600/8-82/005F, July 1985. A-54 SL 0388A5 DRAFT APPENDICES SL 038846 DRAFT Appendix A Methods for Estimating Usage and Emissions of Perchloroethylene in California SL 038847 DRAFT APPENDIX A METHODS FOR ESTIMATING USAGE AND EMISSIONS OF PERCHLOROETHYLENE IN CALIFORNIA The usage and emission estimates for a couple of source categories are based on the estimated shipments of perchloroethylene into California. The ARB staff estimated total shipments as a combination of two types of shipments: 1) direct shipments for use in degreasing and dry cleaning operations, and 2) shipments of perchloroethylene incorporated into products for other uses. Domestic manufacturers estimate that they shipped 19,850 tons of perchloroethylene into California in 1985 for use in degreasing and dry cleaning (Morgan, 1986). In addition, an undetermined quantity of perchloroethylene imported from foreign countries was used in California in degreasing and dry cleaning. This quantity was assumed to be proportional to the ratio of the quantity used in California for degreasing and dry cleaning to the quantity produced and used in the U.S. This quantity is estimated as follows: o 70,000 tons imported into the U.S. in 1985; (U.S. Department of Commerce, Bureau of Census, FT 246, 1985). o 328,000 tons produced and used in the U.S. in 1985 (i.e., U.S. production minus exports); (C&E News, 1987), (U.S. Department of Commerce, Bureau of Census, FT 446, 1985). o 19,850 tons shipped from the U.S. producers in 1985 for degreasing and dry cleaning use in California; so 4,240 tons. Thus, the total quantity of perchloroethylene shipped directly into California in 1985 is estimated to the nearest hundred tons to be: 19,850 tons + 4,240 tons 24,100 tons. Additional perchloroethylene that was incorporated into products was also shipped into California. The ARB staff assumes that the perchloroethylene in adhesives, aerosols, paints and coatings, and miscellaneous uses was incorporated into products. Usage in these categories is estimated below in the section on "Other Uses" to total 2,500 tons. This estimate is for 1983 and the ARB staff assumes the same quantity for 1985. This estimate accounts for imported as well as domestically produced perchloroethylene. Thus, the total quantity of perchloroethylene used in California in 1985 is estimated to the nearest hundred tons to be: Total Usage: 24,100 tons + 2,500 tons - 26,600 tons. DRY CLEANING: The quantity of perchloroethylene used in dry cleaning in California in 1985 is estimated by summing 1) the quantity shipped into California by domestic producers for use in dry cleaning and 2) an estimate of the 0388A8 Appendix A-l DRAFT quantity of imported perchloroethylene used in California for dry cleaning. Domestic manufacturers estimate that they shipped 14,930 tons of perchloroethylene into California in 1985 for use in dry cleaning (Morgan, 1986). The quantity of perchloroethylene imported into California from foreign markets for use in dry cleaning operations is estimated as: o 70,000 tons imported into the U.S. in 1985; (U.S. Department of Commerce, Bureau of Census, FT 246, 1985). o 328,000 tons produced and used in the U.S. in 1985 (i.e., U.S. production minus exports); (C&E News, 1987), (U.S. Department of Commerce, Bureau of Census, FT 446, 1985). o 14,930 tons shipped in 1985 from U.S. producers to California for dry cleaning; so 3,190 tons. Thus, the total quantity of perchloroethylene shipped into California in 1985 for dry cleaning is estimated to the nearest hundred tons to be: Dry Cleaning Usage: 14,930 tons + 3,190 tons = 18,100 tons. To estimate emissions, the ARB staff used information that 0.88 pound of perchloroethylene is emitted for every pound of fresh perchloroethylene used in dry cleaning in the U.S. (Wolf & Meyers, 1987). Thus, an estimate of perchloroethylene emissions from dry cleaning rounded to the nearest hundred tons is: Dry Cleaning Emissions: (18,100 tons) (0.88 lb./1b.) = 15,900 tons. A quick check of this emission estimate was made by using a more general approach. One of the studies cited above includes an emission estimate for dry cleaning in the U.S. of 126,500 tons in 1985 (Wolf & Meyers, 1987). The ARB staff adjusted this estimate by the proportion of the U.S. population in California, i.e. 11 percent (U.S. Department of Commerce, Bureau of Census, 106th Ed., 1985). Estimated emissions using this approach are 13,900 tons. Given the uncertainty that is introduced by using a ratio of populations, this estimate compares well with the original estimate of 15,850 tons of perchloroethylene emissions. The ARB staff allocated perchloroethylene emissions to different sectors of the dry cleaning industry by using EPA estimates of U.S. usage in 1983. The EPA estimates are: (Bath/EPA, 1986). Coin-Operated Usage: 23,060 tons per year; Commercial Usage: 91,100 tons per year; Industrial Usage: 14,360 tons per year. For example, the estimate rounded to the nearest hundred tons for commercial dry cleaning is: Calif. Commercial Emis. * = 11,300 tons. Appendix A-2 SL 038849 DEGREASING: DRAFT The quantity of perchloroethylene used in degreasing in California in 1985 is estimated with the same type of approach used above to estimate the use of perchloroethylene in dry cleaning. Usage is estimated as the sum of 1) the quantity shipped into California by domestic producers and 2) an estimate of the quantity of imported perchloroethylene used in California for degreasing. Domestic manufacturers estimate that they shipped 4,925 tons of perchloroethylene into California in 1985 for use in degreasing (Shimp, 1986). The quantity of perchloroethylene imported to the U.S. for use in degreasing operations in California is estimated as: o 70,000 tons imported into the U.S. in 1985; (U.S. Department of Commerce, Bureau of Census FT 246, 1985). o 328,000 tons produced and used in the U.S. in 1985 (i.e., U.S. production minus exports); (C&E News, 1987), (U.S. Department of Commerce, Bureau of Census FT 446, 1985). o 4,925 tons shipped in 1985 from U.S. producers to California for degreasing; so * 1,050 tons. Thus, the total quantity of perchloroethylene shipped into California in 1985 for degreasing is estimated to the nearest hundred tons to be: Degreasing Usage: 4,925 tons 1,050 tons * 6,000 tons. The ARB staff estimated emissions by using a typical factor for the proportion of perchloroethylene usage that is emitted. An EPA report estimated that 0.92 pound of perchloroethylene is emitted for every pound of fresh perchloroethylene used in degreasing in the U.S. (EPA, 1985). Thus, degreaser emissions rounded to the nearest hundred tons are estimated as: Degreasing Emissions: (5,940 tons) (0.92 lb./lb.) = 5,500 tons. OTHER USES: The EPA reported U.S. usage of perchloroethylene for adhesives, aerosols, paints and coatings, and miscellaneous in 1983 (EPA, 1985). Population fractions were used to estimate California usage from U.S. usage for these categories. California's population was estimated to be 11 percent of the U.S. population in 1985 (U.S. Department of Commerce, Bureau of Census, 106th Ed., 1985). Thus, California usage was estimated as 11 percent of U.S. usage for these categories, or 2,500 tons. Emissions were estimated as less than or equal to usage. The emissions estimate for the use of perchloroethylene in specialty chemical production is based on information supplied by Dow Chemical in Pittsburg, California, to an EPA contractor (Radian Corp., 1986). The Dow Chemical facility is the only known facility using perchloroethylene in chemical production in California. No usage information was available for the facility. Appendix A-3 DISTRIBUTION: DRAFT An estimated 24,100 TPY of perchloroethylene were sold through distribution facilities in California. This quantity includes the above estimates of the perchloroethylene used in California for dry cleaning and degreasing. Emissions from perchloroethylene distribution primarily result from storage in large tanks during distribution. Additional perchloroethylene usage in California was assumed to be incorporated in products and therefore not stored in large tanks. EPA data on the estimated emissions from national distribution were used in estimating the California emissions (EPA, 1985). Emissions were calculated as: (24,100 TPY distributed in rA\ CA) (55 TPY emitted in U.S.) (I78;000`TPra7sfrTBufia"Tn0:S.) 7 TPY. SOLVENT RECLAMATION: An estimated 1,940 TPY of perchloroethylene were sent for recycling in California in 1985. The quantity of perchloroethylene sent for recycling in California was estimated by using: 1) the quantity of halogenated solvent reported on hazardous waste manifests as sent to recyclers, and 2) the estimated percent of perchloroethylene contained in the halogenated solvents. The California Department of Health Services reported that 9,685 tons of halogenated solvents were sent for solvent reclamation in California in 1985 (Radian Corp., 1986). A representative of one of the largest solvent reclaimers in the State estimated that 20 percent of the halogenated solvent sent for solvent reclamation is perchloroethylene (OHS, 1986). Thus, the quantity of perchloroethylene sent for solvent reclamation was jstimated as: (9,685 TPY) (.20) = 1,940 TPY. Industry representatives estimated emission losses for a solvent as volatile as perchloroethylene to be one quarter to one percent of the perchloroethylene received (DHS, 1986), (Schneider, 1986a), (Schneider, 1986b), (O'Morrow, 1986). Thus, perchloroethylene emissions are estimated as: (1,940 TPY) (.0025 to .01) = 5 to 20 TPY. POTWs: See the discussion on POTWs in the body of the report. MUNICIPAL LANDFILLS: The estimate of perchloroethylene emissions from a typical municipal landfill is based on 1) a mean surface gas concentration of perchloroethylene of 0.3 parts per million, 2) a typical volume of landfill gas emitted per square foot of landfill surface per day of 0.8 SCFT/ft/day (SCFT stands for standard cubic foot), and 3) an estimated surface area of a landfill of 40 acres. The Gas Research Institute reported a mean Appendix A-4 sv DRAFT concentration of 0.3 parts per million (volume) of perchloroethylene for 83 surface gas samples from nine landfills in the U.S. (Roehl Disposal Corp., 1986). This same reference estimates that a typical volume of methane emitted from a landfill surface is 0.4 SCFT/ft*/day. ARB staff assumed that landfill gas is approximately 50 percent methane. ARB staff also assumed a typical size of a municipal landfill to be 40 acres. HAZARDOUS WASTE LANDFILLS: Emission estimates for individual hazardous waste landfills are presented based on three methods of estimation, including: EPA estimates based on source monitoring at three hazardous waste landfills; ARB staff estimates of emissions from a hazardous waste landfill of median size; and EPA estimates of modeled emissions. The EPA reported mass emission rates for three California landfills, including two active and one inactive landfill. The perchloroethylene emission rates were as follows: 3.64 TPY for a 215,000 ft* active landfill; 0.03 TPY.for a 16,000 ft* area of another active landfill and 0.30 TPY for a 7,200 ft* section of the same landfill; and 0.003 TPY for a 32,400 ft* inactive landfill (G.R.I., 1982), (EPA, 1984). ARB staff developed emission estimates for a typical landfill based on the above EPA source monitoring data and data on the median size of a landfill. Data on the median size (971,450 ft*) of a hazardous waste landfill in the U.S. were taken from a article published in the Journal of the Air Pollution Control Association (EPA, 1985). The range of perchloroethylene emissions from an active landfill of median size is estimated from the above data as: Low estimate: (971.450 ft2) = 1.8 TPY. (16,000 ft*) High estimate: (971,450 ft2) . 16.4 TPY. (215,000 ft*) The emission estimate for an inactive landfill of median size was calculated in a similar manner. SURFACE IMPOUNDMENTS: See the discussion on surface impoundments in the body of the report. SL 038852 Appendix A-5 DRAFT REFERENCES I. Bath, D. Blake, 1986, Radian Corporation, memorandum to Mark Meech of EPA/OAQPS entitled: "Documentation of Emissions and Exposure Model Inputs for the Perchloroethylene Dry Cleaning Source Category." May 8, 1986. 2. California Department of Health Services, 1986. Computer printout from hazardous waste information system giving disposal method information for 1985. April 11, 1986. 3. Chemical and Engineering News. "Production by the U.S. Chemical Industry," June 8, 1987. 4. DeVries, P., 1986. Personal communication. Telephone conversation between Ronald Rothacker of CARB and Paul DeVries of Roehl Disposal Corporation. September 22, 1986. 5. Gas Research Institute. Landfill Methane Recovery Part II:__ fiai. Characterization. December 1982. 6. Howie, R., 1986. Personal communication. Telephone conversation between Ronald Rothacker of CARB and Reese Howie of Radian Corporation, Research Triangle Park, NC. July 22, 1986. 7. Morgan, D.L., 1986. Correspondence from D.L. Morgan of Cleary, Gottlieb, and Hamilton to Dale Shimp of CARB. October 28, 1986. Response to request for dry cleaning and degreasing uses of perchloroethylene in California in 1985. 8. O'Morrow, K., 1986. Personal communication. Telephone conversation between Ronald Rothacker of CARB and Ken O'Morrow of Oil and Solvent Process Co., Azusa, CA. June 9, 1986. 9. Schneider, P., 1986a. Personal comuunication. Telephone conversation between Ronald Rothacker of CARB and Peter Schneider of Romic Chemical Corporation, East Palo Alto, CA. June 9, 1986. 10. Schneider, P., 1986b. Personal communication. Telephone conversation between Ronald Rothacker of CARB and Peter Schneider of Romic Chemical Corporation, East Palo Alto, CA. October 11, 1986. II. U.S. Department of Commerce, Bureau of rensus. U.S. Imports for Consumption and General Imports. TSUSA Commodity bv Country of Origin. FT 246/Annual 1985a. 12. U.S. Department of Commerce, Bureau of the Census. U.S. Exports. Schedule B Commodity Bv Countv. FT 446/Annual, 1985b. 13. U.S. Department of Commerce, Bureau of the Census. Statistical Abstract of the United States 1986. 106th Edition. December 1985c. Appendix A-6 SL 038853 DRAFT U.S. Environmental Protection Agency (EPA), 1984. Emission Test Report No. 85-HWS-2. Tests conducted at Chemical Waste Management, Inc. Kettleman Hills Facility, Kettleman Hills, CA. November 21, 1984. U.S. EPA, 1985a. Emission Test Report No. 85-HWS-l. Tests conducted at IT Corporation Benicia Facility, Benicia, CA. May 1985a. U.S. EPA, 1985b. Survey of Perchloroethvlene Emission Sources. EPA450/3-85-017, Research Triangle Park, NC. June 1985b. Vogel, Gregory A., 1985. "Air Emission Control at Hazardous Waste Management Facilities," Journal of the Air Pollution Control Association. May 1985. Wolf, K. and C.W. Myers. Hazardous Waste Management bv Small Quantity Generators - Chlorinated Solvents in the Drv Cleaning Industry. Rand Corporation. June 1987. 03885A Appendix A-7 DRAFT APPENDIX B Standard Methods of Analysis for Tetrachloroethylene SL 038855 DRAFT Method No. ADDL002 October 16, 1986 Revision: 3J Approved: liZZ Page T of 14 Pages METHOD NO. ADDL002 STANDARD OPERATING PROCEDURE FOR THE DETERMIKATION OF VOLATILE ORGANICS IN AMBIENT AIR USING TEKAX TRAP PRECONCENTRATION GAS CHRUMA1OGRAPHY AND TARDEH PHOTOIONIZATION/ELECTRON CAPTURE DETECTORS 1.0 SCOPE This document describes a procedure for the determination of volatile halogenated hydrocarbons and aromatics having a boiling point of less than 12Q'C. This procedure is based on documents received from the ARB Haagen-Smit Laboratory, El Monte, as well as EPA Method T01. 2.0 SUMMARY OF PROCEDURE Ambient air is continuously sampled and collected in a Tedlar bag over a 24 hour period and immediately sent to the laboratory for analysis. A sample from the bag is drawn through a sampling valve attached to a . Tekmar LSC-2 Tenax Sample Concentrator (see Figure I) with a vacuus pump at 50 cc/min for four minutes (total sample volume: 200 cc). The organic constituents are trapped on Tenax and when the collection is complete, the Tenax is purged with 40 cc of helium to remove any trapped moisture. The sample is then thermally desorbed onto the head of the GC column. The GC column is temperature programmed and component peaks Si 03*8S6 Appendix B-l eluting from the column are sequentially detected and quantified, first by a photoionization detector (PID) and then by an electron capture detector (ECO). The components are 1 dent*.flee oasec cn retention tines. Positive Identification or confirmation requires the use of an appropriately configured GC/MS. IMTERFERNCS5/lIMITATIONS a. Components having similar GC retention times will interfere, causing misidentification and/or faulty quantitation. b. Because of the very low sample concentrations, extreme care must be taken to insure that the sample is not degraded or contaminated by the Tedlar sampling bag, sampling apparatus, or delayed delivery to the laboratory. Exposure of the Tedlar sampling bag to temperatures greater than 25*C should be minimized. c. Only components of the sample which can be detected by PID/ECD detectors will be quantified. 4.0 APPARATUS a. Varian Model 6000 Gas Chromatograph/PID/ECD system equipped with a Varian Vista 402 dual channel data system. b. Tekmar ISC-2 Sample Concentrator equipped with Tenax trap and sampling valves as shown in Figure 1. Appendix B-2 SL 038857 c. Hatheson Model 8240 Hass Flow Controller accurately calibrated In the 5-100 cc/nin range. Laboratory timer, accurate to within 0.1 minutes. Gas tight mlcrollter syringe, 50 ul. f. GC column - 10' x 2 bid i.d. glass column packed with 1 percent SP-1000 on Carbopack B, 60/80 mesh. 5.0 REAGENTS a. Primary Gas Standard (Scott Specialty Gases - Research Triangle Institute Certified Series 1) Compound Concentration (ppb) Chloroform Carbon tetrachloride Perchloroethene Vinyl chloride Benzene 107 105 106 104 107 SL 038858 Appendix B-3 b. Primary 6as Standard (Scott Specialty Gases - Research Triangle Institute Certified Series 2) Compound Concentration (ocb) 1.2-D1chloroethane 1,1,1-Trichloroethane Trichloroethene 1.2-Dibromoethane 101 ye 100 102 c. Stock Gas Standard - Scott-Marrin Blend (assayea against primary cylinders) Compound Concentration (ppb) Dlchloromethane Chloroform 1.2- Dichloroethane 1,1,1-Trichloroethane Carbon tetrachloride Trichloroethene 1.2- Oibromoethane Perchloroethene Ylnyl chloride Benzene 4272 528 3104 424 46 336 5 43 4736 1888 Appendix B-4 d. Control Gas Standard - Scott-Harrln Blend (assayed against primary cylinder) Cocnoound Concentration (odd) 01 chioromethane Chlorofonn 1,2-D1ch1oroethane 1,1,1-Trichloroethane Carbon tetrachloride Trichloroethene 1,2-Dibromoethane Perchloroethene Vinyl chloride Benzene 6 0.2 0.2 3.6 0.3 1.8 2.5 1.2 3.3 4.6 e. Surrogate Gas Standard (Scott-Marrin Blend) Compound Concentration (ppm) Bromochloromethane 1,3-Bromochloropropane 10 33 SL 038860 Appendix B-5 6.0 PROCEDURES a. Ear,tie Traooina 1. The preconcentration system Is shown in Figure 1. 2. The high concentration Inlet Is used for high concentration calibration standards and for other samples with concentrations higher than ambient levels. The simple Is introduced through the high concentration inlet and 6 port valve into an appropriate size loop of known volume. Tine sample then passes through a 10 port valve, mass flow meter, and vacuum pump. Before an analysis, the system is leak checked by flocking the sample inlet port and observing that the mass flow meter reading drops to zero. The high concentration inlet then is connected to a Tedlar sample bag valve andi the gas bag valve opened. The loop is then flushed with sanple gas for three minutes. After three minutes of flushing,, the 6 port valve is reset so that the sample contained in the loop is carried into the trap by the helium purge gas. This continues for three minutes to ensure that all of the contents of the loop are trapped. Appendix B-6 SL 038861 3. Ambient samples are Introduced from Tedlar bags as describ d above, except that the samole loop is bypassed and the samele goes directly to the 10 port valve. After flusn*.ng the system with sample for three minutes, the 10 port valve is reset so that 200 cc's of sample is trapped (50 cc/uin. for four minutes). After sample trapping is complete, the Tenax trap is flushed with 40 cc of helium to remove water vapor and any nonadsorbed reactive gases. 4. In both ambient and high concentration cases, after the sample has been trapped, the Tekmar LSC-2 heats the Tenax trap to 180*C while the trap is swept with the 6.C.'s internal carrier gas for four minutes. The contents of the trap are thus desorbed and collected on the head of the &.C. column. The trap is baked out after the end of the desorption cycle. In the bakeout cycle, the trap is flushed with helium purge gas for eight minutes while being held at 225'C in order to prepare the trap for the next cycle. After bakeout the trap is isolated from the system and ready for the next sample. b. Analysis 1. The concentrated sample is separated under the chromatographic condition detailed below. The resulting chromatogram (see Figure II) is then integrated and quantified by reference to calibration standard gases. SL 038862 Appendix B-7 .2 Instrument Conditions: GC: Column: 1C` s l jtt. :.c. class column, pacfceo witr 1 percent SP-10QQ on C-arbopacfc B 60/80 mesn Temperatures: Injection: 200'C Detector: 35Q"C Oven: 45"C, hold for four minutes, bT/min raanp, to 210*C, hold for eight minutes Flow Rates: Carrier: He, 20 c.c/rain ECO make up: N^, 40 c.c/min Detectors: ECO: Range X 10, Attenuation X 32 P1D: Range X 1, Attenuation X 32, 10.2 Cone: Tekmar LSC-2: Purge: 4 minutes Desorb: 4 minutes at 180"C Bake: 8 minutes at 225"C Appendix B-8 3*863 J All blanks, standard.;, control samples, and ambient samples are spiked with surrogate compounds by injecting: 0 microliters cf the surrogate gas standard (S.e-.) during sample trapping. The surrogate compounds, chosen such that they simulate the characteristics of the analytes of Interest and are unlikely to occur in the environment, are added to insure that systematic errors or equipment failures will be noted and corrected promptly. 4. The first step in a calibration is to analyze a system blank. This is done by trapping and analyzing a 200 cc sample of auxiliary carrier gas. The system blank must be free of interfering peaks. A system blank must also be run after a high concentration sample is analyzed in order to detect any carry-over within the system. 5. A calibration is performed using a 1.25 cc loop of stock standard gas (5.c.). Two hundred cubic centimeters of helium gas is passed through the loop to carry the standard onto the trap. The calibration analysis is made as a normal analysis. The calculated concentration value for each component should be Inspected to insure consistency with previous analyses. The stored chromatographic information may then be used to recalculate the response factors for the subsequent analyses. The G.C. data system will not accept updated response factors which are in excess of plus or minus 15 percent of historic data. SL 038864 Appendix B-9 6. Following calibration, 200 cc of the control staple (S.d.) Is concentrated on the trap and analy2eo. The control sample cata are plotted on control charts-of the normal Shewnart type. Upper and lower warning Units are plus or sinus two times the standard deviation. Any analysis which falls outside the upper and lower warning limits is repeated and the laboratory quality control officer Is advised. Upper and lower control Units are plus or minus three times the standard deviation. If any analysis falls outside the upper or lower control Unit, the method is discontinued until the out of control situation is remedied. The laboratory quality control officer is advised and provided with written documentation of the out of control condition and how it was remedied. All data generated prior to the out of control situation must be reviewed for possible decertification by laboratory management. 7. Multipoint calibrations are conducted monthly. Each multipoint calibration includes a trap blank and three standard concentration levels to bracket the concentration ranges expected in ambient air. If subsequent data indicate that the resulting least squares analyses are consistently acceptable, less frequent multipoint calibrations may be made. Appendix B--10 SL 038865 7.0 PERFORMANCE DRAFT a. All amoient field samples are analyzed In duplicate. The relative error between analyses must be less than 20 percent. Duplicate analyses having greater than 20 percent relative error must be decertified. b. .The percent recovery of the surrogate is recorcea in the instrument laboratory workbook for each analysis. If this value is outside the 801 to 1201 range, the sample analysis must be repeated. 8.0 METHOD SENSITIVITY, PRECISION AND ACCURACY The method sensitivity, precision and accuracy are outlined in Table I. These data were produced with gaseous calibration standards, and using carrier gas as the sample matrix. The relative accuracy of the method, with the exception of dichloromethane, is based on reference to the Research Triangle Institute Certified Gas Stanoarcs (NBS traceable). Authoritative reference calibration standards for cichloromethane ar under development at NBS but are not yet available. The concentration value of the present standard was assigned by the cotnnercial manufacturer and found to be in good agreement with diluted pure dichloromethane prepared in our laboratory. The aosolute accuracy of the method has not been determined by interlaboratory testing. SL 038866 Appendix B-ll Simple Loop Figure l. Schematic of concentrator system. Sampling Conditions are: 200 cc volume, purge at 40cc/min, 1 min., desorb at ISO C for 4 min., bake for 8 min. at 225 C. Operational step Loop Fill Loop Trap * Ambient Trap Trap Desorb Trap Bake Out SYSTEM GUIDE Valve Position 6-Port 1 2 1 1 1 10-Port 1 1 2 1 1 LSC-2 1 1 1 2 1 Purge Gas Off On Off Off On Appendix B-12 SL 038867 14 q38868 1. Vinyl Chloride 2. Dlchloromethane 3. Trichlorofluoromethane 4. 1,1-01 chioroethy1 ene 5. Bromochloromethane 6. 1,1-Diehloroethane 7. t-l,2-Dichloroethylene 8. Chloroform 9. Freon 113 10. 1,2-Dichi oroethane 11. 1,1.1-Trichloroethane 12. Carbon Tetrachloride 13. Trichloroethylene 14. Benzene 15. 1,2-Dlbromoethane 16. Brooochloropropane 17. Tetrachloroethylene 18. Toluene Appendix B-13 DRAFT Table I Method Sensitivity and Precision Comoound Vinyl Chloride D1chioromethane 1,1-Dichioroethy1ene Chloroform 1,2-Dichloroethane 1,1,1-Trichloroethane Carbon Tetrachloride Trichloroethylene 8enzene 1,2-Dibromoethane Tetrachl oroethy 1ene Correlation Coefficient 0.997 0.999 0.991 0.999 0.999 0.999 U.9y9 0.999 0.998 0.974 0.994 . Slope R.S.D* (Percent! 0.946 16 0.975 5 U.S66 6 0.901 3 1.054 7 0.989 9 0.980 6 0.992 6 0.950 10 1.067 9 1.080 10 Detector PID ECO ECD ECO - ECD ECD ECO ECD PID ECD ECD LOD PP*>V 0.8 0.6 0.06 0.02 0.1 0.01 0.005 0.02 1.5 0.006 0.01 * R.S.D. - Relative Standard Deviation at 5 x LOD, n 5 Appendix B-14 sv 38869 TECHNICAL SUPPORT_DOCUMENT II PROPOSED IDENTIFICATION OF PERCHLOROETHYLENE AS A TOXIC AIR CONTAMINANT MRT B REPORT STATE OF CALIFORNIA AIR RESOURCES BOARD STATIONARY SOURCE UVI8ION SL 038870 DRAFT HEALTH EFFECTS OF TETRACHLOROETHYLENE (PCE) CALIFORNIA DEPARTMENT OF HEALTH SERVICES JANUARY, 1989 DRAFT Prepared By: California Department of Health Services Principal Editor: George V. Alexeeff, Ph.D. Douglas N. Reviewed and Revised by Lauren Zeise, Ph.D. Calvin Wilhite, Ph.D. Cox, Ph.D. (California Public Health Foundation) Hanna H. Schunk, M.S. (Contractor) Based on a Report Prepared by K. T Bogen, L. C. Hall, T. E. McKone, D. W. Layton and S. E. Patton, of che Lawrence Livermore National Laboratory, University of California Livermore, California, and titled "Health Risk Assessment of Tetrachloroethylene (PCE) in California Drinking Water." SL >Sh?3 DRAFT TABLE OF CONTENTS 1. Executive Summary................................................................................................................. 1-01 Evaluation Perspective..............................................................................................1-11 2. Pharmacokinetics and Metabolism .......................................................................... 2-01 Absorption ............................................ 2-02 Ingestion......................................................................................................................2-02 Dermal Absorption ....................................... 2-03 Pulmonary Uptake.................................................................................................. 2-03 Distribution and Bioaccumulation..................................................................... 2-06 Metabolism and Elimination....................................................................................2-15 3. Toxic Effects in Animals..............................................................................................3-01 Toxic Effects on Organs and Systems.................................................................3-02 Hepatic Toxicity....................................................................................................... 3-02 Renal Toxicity....................................................................................................... 3-05 Pancreas......................................."............................................................................3-07 Lungs, Skin, and Eyes.........................................................................................3-08 Reproductive System............................................................................................. 3-08 Cardiovascular System.........................................................................................3-09 Central Nervous System....................................................................................3-10 Teratogenicity................................................................................................................. 3.^ Mutagenic Effects............................................................................................................ 3-16 Microbial Assays.................................................................................................. 3-17 Drosophila.................................................................................................................3-20 Hi Si* 038873 Tests of DNA or Chromosomal Damage ................................................................ 3-20 Mutagenic Activity of Metabolites ........................................................... 3-22 Carcinogenicity in Animals ................................................................................... 3-22 Summary of Evidence of Carcinogenicity in Animals .................... 3-31 4. Toxic Effects in Humans...................................................................................................4-01 General Toxicity............................................................................................................ 4-01 Epidemiologic Evidence forCarcinogenicity in Humans........................4-02 Toxicity to Major Organs andSystems.............................................................. 4-11 Liver................................................................................................................................4-11 Kidneys...........................................................................................................................4-15 Lungs................................................................................................................................4-16 Skin and Eyes ....................................................................................................... 4-16 Connective Tissue.................................................................................................. 4-17 Central Nervous System . -.......................................................................... 4-1 57 * Reproductive System ......................................................................................... 4-20 Cardiovascular System ................................................................................. 4-20 Teratogenic Effects................... *...........................................................................4-20 Mutagenic Effects........................................................................................................4-21 Summary of Evidence of HumanCarcinogenicity...........................................4-21 5. Quantification of PCE's CarcinogenicPotency..................................................... 5-0L Selection of Bioassay Data Indicative of PCE Carcinogenicity........................................................................................................5-01 Pharmacokinetic Analyses......................................................................................... 5-05 u iv Methodology DRAFT ................................................................................................................5-07 Evaluation of Cancer Potency .................................................................... 5-07 Interspecies Scaling ........................................................................................ 5-09 Dose Adjustments ............................................................................................ 5-09 NCI (1977) Mouse Study: Gavage ........................................................... 5-10 Applied Dose And Tumor Incidence ............................................ 5-10 Metabolized Dose ..............................................................................5-11 NTP (1986) Rat Study: Inhalation ...................................................... 5-15 NTP (1986) Mouse Study: Inhalation ................................................. 5-20 Results of Metabolized Dose Calculations ............................. 5-22 Carcinogenic Potency Extrapolation Based on Animal Bioassay Data...................................................................................................... 5-23 Carcinogenic Potency in Terms of Human Applied Dose ......................... 5-27 Exposure Route ................................................................................................. 5-27 Interspecies Dose Equivalence ............................................................... 5-28 Human Metabolism.................................................................................................. 5-29 Excretion Half-Life In Humans ............................................................... 5-30 Urinary Metabolites as a Fraction of Total Metabolites 5-31 Continuous Versus Peak Exposures ...................................................... 5-34 Ogata and Co-workers (1971).......................................................................... 5-36 Fernandez and Co-workers (1976).................................................................5-36 Monster and Co-workers (1979).................................................................... 5-37 Bolanowska and Golacka (1972)..................................................................... 5-38 Metabolic Parameters Estimated from Different Human Data Sets ....................................................................................................... 5-39 SL 038875 v PCE Carcinogenic Potency as a Function of Human Applied Dose (TakingPCE Metabolism into Account). . . .5-43 Recommendation........................................................................................................ 5-48 Appendix A. Dose-Response Information for Acute, Subchronic, and Chronic Toxicity in Animals (Excluding Teratogenic, Mutagenic, and Carcinogenic Effects).............................................................................. A-01 References.............................................................................................................................................. R-01 vi SL 038876 DRAFT 1. EXECUTIVE SUMMARY Tetrachloroethylene, commonly referred to as perchloroethylene (PCE), is a volatile organic hydrocarbon with a chloroform-like odor used as a solvent in dry cleaning operations, a metal degreaser, a chemical intermediate in the synthesis of fluorocarbons, and in coffee decaffeination. Perchloroethylene is a lipophilic compound which readily diffuses into the blood and subsequently into adipose tissue where it accumulates due to its relative stability and slow metabolism. The main metabolic pathway for PCE in humans appears to involve its oxidation by cytochrome P-450-dependent mixed-function oxidases (referred to as P-450) to 1,1,2,2- tetrachloroethylene oxide. Following an apparently spontaneous chloride migration, the metabolites trichloroacetyl chloride and trichloroacetic acid are formed. Trichloroacetic acid accounts for about 60% of the trichlorocompounds found in urine after exposure to PCE. The chemical nature of the remaining 40% of trichlorocompounds has not been established, but is assumed to be trichloroethanol. Trichloro compounds account for most of the metabolism of PCE in rodents. However, the human PCE metabolic data are very incomplete and other unidentified products are also produced. Following P-450 metabolism and several dechlorination steps, oxalic acid and carbon dioxide (CC^) also may be produced in humans since they are common metabolic products in animals. Perchloroethylene is eliminated from the body by two major processes: At high concentrations (>50 ppm), the majority of absorbed PCE is excreted unchanged in the exhaled air while the remainder may undergo metabolism and 038877 1-1 DRAFT excretion as urinary metabolites. At lower concentrations, relatively less FCE is exhaled. Perchloroethylene has moderate acute toxicity, with the liver being its principal target organ. Acute inhalation exposure of mice to 200 ppm for four hours produced moderate fatty infiltration in the liver. Chronic exposure of laboratory animals to 100 ppm PCE caused major liver damage. The products of PCE metabolism are thought to promote liver toxicity. The extent of PCE metabolism was directly proportional to observed liver cell damage in animal experiments. The method by which PCE causes liver toxicity is not known, but metabolites of PCE may bind covalently to cell components. These bound metabolites have a slow turnover rate in the liver and accumulate upon chronic exposure. In particular, PCE appears to damage mitochondria. Perchloroethylene causes skin and eye irritation. Prolonged PCE exposure can produce erythema, burns, and blistering. It has been associated with tachycardia and sudden death from cardiac failure, and PCE may sensitize the heart to the effects of endogenous epinephrine. Massive acute exposure to PCE induces central nervous system (CNS) depression that can progress to loss of consciousness, anesthesia, and respiratory failure. Inhalation exposure of pregnant rodents to 300 ppm PCE produced maternal toxicity and fetotoxicity manifested as developmental delays and altered performance in behavioral tests in the offspring of exposed mice and rats. The No-Observed-Adverse-Effect-Level (NOAEL) for chronic inhalation in rats was reported to be 70 ppm PCE. However, rats are less susceptible than mice SL 038878 1-2 co hepatotoxicicy, the most sensitive noncancer endpoint for toxicity. Mice chronically exposed to 100 ppm PCE exhibited liver degeneration and necrosis, whereas 200 ppm PCE given to rats under the same regimen did not result in hepatic lesions. A NOAEL for mice has not been established. Humans have shown signs of liver toxicity after chronic exposure to 232-385 ppm PCE, indicating that humans may be as susceptible as mice to the hepatotoxic effects of PCE. Maximum concentrations of PCE measured in urban areas were approximately 5 ppb, a level well below the NOAEL. Consequently. California Department of Health Services (CDHS) staff do not expect noncarcinogenic adverse health effects to occur from acute or chronic exposure to PCE in ambient air. The current 8-hour threshold limit value (TLV-TWA) for PCE in the workplace is set at 50 ppm, to provide a margin of safety against eye and respiratory discomfort and subjective complaints experienced at exposure levels of 100200 ppm (ACGIH, 1986). The National Institute for Occupational Safety and Health (NIOSH) does not suggest , a safe exposure level but recommends reducing exposure to the lowest feasible limit to prevent potential carcinogenic effects (NIOSH, 1985). Perchloroethylene increased the incidence of hepatocellular tumors in laboratory mice after oral and inhalation exposure and mononuclear cell leukemia in rats after inhalation. A lifetime inhalation bioassay was conducted by the National Toxicology Program (NTP) in which mice (of both sexes) were exposed to 100 or 200 ppm of PCE and rats (of both sexes) were exposed to 200 or 400 ppm PCE (NTP, 1986). Both concentrations of PCE resulted in a statistically significant increase in hepatocellular carcinoma 1-3 and adenoma in treated mice of both sexes. The incidence of mononuclear cell leukemia was significantly increased in rats of both sexes. The NTP concluded that, under the conditions of their study, there was clear evidence of carcinogenicity of PCE for male F344/N rats, some evidence of carcinogenicity of PCE for female F344/N rats, and clear evidence of carcinogenicity of PCE for both sexes of B6C3F1 mice. In an earlier study by the National Cancer Institute (NCI, 1977) male and female mice developed hepatocellular carcinoma after oral gavage exposure to PCE 5 days/week over 78 weeks. Rats receiving PCE via gavage did not develop significant increases in tumors, but early mortality in treated animals limits the usefulness of this study in characterizing the oral carcinogenicity of PCE in rats. Epidemiological studies have provided some indication that the use of drycleaning solvents, primarily PCE, poses an in :reased risk of cnnct exposed workers. However, investigators were unable to differentiate amt:'-? exposures to various solvents, and possible confounding factors like smoking were not evaluated. Therefore, the usefulness of these reports in the assessment of the human cancer risk from PCE exposure is limited. The International Agency for Research on Cancer (IARC) reviewed the carcinogenicity data on PCE and placed it in category 2B. with sufficient evidence of carcinogenicity in animals and inadequate evidence in humans (IARC. 19871. The CDHS staff agree with IARC's conclusion. Prior to the publication of the results of the NTP (1986) inhalation bioassay the EPA concluded that the evidence for carcinogenicity of PCE in animals was limited and that the available epidemiological data for human 038880 Sl> carcinogenicity was inconclusive. Therefore PCE was placed in group C, as a possible human carcinogen (EPA, 1985a). However, in light of the NTP inhalation bioassay the classification is being reconsidered and it is likely that PCE will be placed in group 2B, as a probable human carcinogen. Perchloroethylene has generally produced negative results in genotoxicity assays using bacterial systems, but reportedly produced a dose-dependent increase in base-pair substitutions and frameshift mutations in Salmonella typhimurium TA 100 with metabolic activation (Cerna and Kypenova, 1977). An increase in mitotic recombinations and gene conversions were also reported in Saccharomvces cerevisiae strain D7 after exposure to PCE without metabolic activation (Callen et al., 1980). Perchloroethylene oxide, the first intermediate formed by microsomal oxidation, exhibited a dosedependent mutagenic response in Salmonella typhimurium TA 1535. Trichloroethanol, another metabolite of PCE, induced sister chromatid exchange in cultured human lymphocytes. These responses indicate that PCE itself and/or some of its metabolites are potentially genotoxic. The U.S. Environmental Protection Agency (EPA), after reviewing the literature on the mutagenicity of PCE, concluded that inadequate information exists to classify PCE as either a mutagen or a nonmutagen (EPA, 1985). The staff of the CDHS agrees that PCE's mutagenicity has not been clearly established, but the positive results in some of the genotoxicity assays with PCE indicate a possible interaction with DNA and the potential for PCE to be an active genotoxin. CDHS staff have found no evidence of a carcinogenic threshold level. Thus, although the mutagenic activity of PCE is unclear. the staff recommends that PCE be considered as not having a threshold for carcinogenicity. SL 038881 1-5 DRAFT R suits from the 1986 NTP inhalati n study were used as the basis for estimating the carcinogenic risk of PCE to humans. In this bioassay, PCE was 99.9% pure, and animals were exposed 6 hours/day, 5 days/week for 103 weeks. The mice in the 100 and 200 ppm dose groups were exposed to a timeweighted average (TWA) of 16 and 32 ppm, respectively (e.g., 100 ppm x 6 hours/24 hours x 5 days/7 days). Similarly, rats in the 200 and 400 ppm dose groups were exposed to a TWA of 33 and 66 ppm, respectively. The CDHS staff used the applied dose, adjusted to continuous lifetime exposure, to calculate the carcinogenic potency of PCE. A metabolized dose was not included in the suggested range of risks for several reasons. The metabolized dose adjustment did not reduce the uncertainties in the PCL risk assessement, but instead elucidated new areas of uncertainty. Uncertainties of low dose extrapolation were compounded by uncertainties in dose calculations. The current data available on the quantity of PCE metabolized by different pathways in different species is incomplete. This uncertain!./ is evident in the huge range of the metabolized doses obtained from ' different pharmacokinetic models. For example, the metabolized dose calculated by several methods, for the production of male mouse liver tumors vary by five-fold (19.5 mg/kg-day to 93.8 mg/kg-day). The metabolized dose approach, as it has been applied to PCE, assumes that oxidative metabolism leads to the production of carcinogenic metabolites; but the ultimate carcinogen(s) has not been well characterized. The extent of oxidative metabolism has generally been inferred from the production of "trichloro" metabolites. Metabolism to nontrichloro metabolites has not been quantified in humans, and a mass balance approach identifying the distribution and excretion of inhaled PCE has not b en conducted. In humans, urinary SL 038882 1-6 trichloro compounds account only for 1-2% of the inhaled PCE following exposure to concentrations of approximately 100 ppm, whereas 60-80% of the PCE taken in is exhaled unchanged through the lungs during the exposure. This implies that 20-40% of the absorbed PCE is stored (presumably in adipose tissue) and is slowly released over time. The high storage capacity and the slow release of PCE from adipose tissue prevent verification of the mass balance relationship proposed by pharmacokinetic models between PCE absorbed and PCE metabolized or excreted either as urinary metabolites or unchanged by the lungs. The residual PCE could be metabolized to trichloro compounds over time, or to nonchlorinated metabolites such as C02 and oxalic acid that could not be easily associated with PCE metabolism without use of radiolabeled PCE. Data on the amount of PCE metabolized at ambient concentrations less than 1 ppb) are not available. However, several studies indicate that PCE metabolism increases as the concentration decreases. Extrapolation of metabolism rates to ambient levels indicates that up to 100% of the PCE mav be metabolized by mice and rats. Humans may metabolize up to 60% at ambient levels compared to the 4% level estimated from experiments using concentrations 100,000-fold greater than ambient levels. Pharmacokinetic models generally do not account for individual differences in metabolism and storage. A high variability of body burden of PCE was found for different people tested (Guberan and Fernandez, 1974; Hake and Stewart, 1977; Stewart et al., 1970). The body burden depended on such factors as age, box. exercise or workload, body mass, adipose tissue mass, pulmonarv dysfunctional states, and individual differences in the intrinsic capacity to metabolize PCE. SL 038883 1-7 DRAFT The pathways of PCE metabolism are speculative at this point in time, but it is known that cytochrome P-450 is involved. The presence and basal activity of cytochrome P-450-dependent enzymes are determined genetically, but such systems may be altered by components of the diet. Pretreatment of rats with inducers of P-450 increased PCE metabolism five- to seven-fold. Epidemiologic evidence relating enzyme induction to components of human diet are not available at this time, but appears likely. Human variability in PCE metabolism could be accounted for through the incorporation of generic safety factors, although none of the available pharmacokinetic models has incorporated such a safety factor. Finally, chronic studies in mice and occupational studies in humans indicate that mice and humans appear to have similar sensitivity to PCE's induction of noncarcinogenic liver effects. Since carcinogenic and noncarcinogenic effects are thought to occur via the same metabolic pathway, the fifty-fold species difference suggested by the metabolized dose calculations does not appear justified. Thus, it appears premature to use the metabolized dose in current estimations of human ri^r to PCE. However, due to the importance of the area of study, calculations % using such methods were presented in Section 5. For the low-dose PCE risk assessment, the Crump multistage polynomial (Crump, 1981) was chosen. This model, rather than a time dependent form ol the multistage model, was chosen because most tumors were discovered only at the time of sacrifice, and survival in this study was relatively good. The amount absorbed per unit surface area was taken as the correct measure of dose and the so-called surface area correction factor was used to determine human exposure. SL 038884 1-8 DRAFT Using extra the data of both rats and mice from the NTP 1986 study, the number of 3 cancer cases expected from a lifetime exposure to 1 ppb (6.89 ^g/m ) of PCE was estimated to be between 31 to 144 cases per million persons exposed. -6 This corresponds to a range of unit risks of 5 to 21 x 10 for a 3 lifetime continuous exposure to 1 ng/m of PCE. This range represents the upper limit of upperbound risk plausible excess cancer cases. The potency for the 3 due to lifetime exposure to 1 fig/m of PCE in air (2.9 x -7 -7 10 to 9.5 x 10 ) calculated using a metabolized dose method suggested by the EPA (EPA Review Draft, 1986), is one order of magnitude lower than that derived by CDHS. However, the EPA reported that the complete range of potencies for all methods and dose-tumor incidence data evaluated for 1 3 .7 .5 Mg/m of PCE in air (2.9 x 10 to 1.1 x 10 ) includes the above range of 5 to 21 x 10 6 derived by CDHS. Consequently. the ranee of risks calculated from analysis of animal studies recommended bv CDHS for the purpose of estimating cancer risk, lie between 31 x 10 /oob (5 x 10~ /ug/m ) and 144 x 10 6/ppbI (21 x 10 6/xig/m3) . The mean environmental concentration measured by the Air Resources Board staff is 0.43 ppb in the South Coast Air Basin; that is, more than four orders of magnitude below the time-weighted average values used in the animal bioassay. Based on the CDHS potency evaluation and the annual average of 0.43 ppb PCE in the South Coast Air Basin, an upper limit of 133 to 619 additional lifetime cancer cases are estimated in the 10 million residents of the South Coast Air Basin as a result of PCE exposure. The calculations represent the upper range of plausible excess cancer risk; the actual risk, which cannot be calculated, may be insignificant. Based on --finding of carcinogenicity and the results of the risk assessment, the SL 038885 1-9 DRAFT risk assessment, the staff of CDHS concludes that. PCE is an air__ pollutant which mav cause or__ contribute to an increase in mortality or in serious illness, or which mav pose a present or potential hazard to human health. 1-10 SL 038886 Exposure Sources DRAFT EVALUATION PERSPECTIVE A. Air Levels 1. Mean ambient levels measured outdoors in the South Coast Air basin: 0.43. 2. Ambient levels measured in "hot spots": Information not available. 3. Indoor Air: PCE concentrations measured in indoor air vary depending on the quantity released from water and consumer product sources, the time since release, and the size of the room. The mean indoor air concentration measured in the U.S. ranged from 0.34 to 1.01 ppb. Maximum concentrations in homes reached levels of 14.1 ppb. B. Reported Levels in Water 1. National Data: Ambient waters: mean concentration was 1 Mg/L in 1,102 surface water measurements in 45 states. 2. California drinking water: PCE was present in 199 wells out of 2,947 sampled. The median concentration of the 199 wells was 1.9 Mg/L with the highest concentration being 166 Atg/L. 038887 DRAFT C. Reported Levels In Food 1. Highest levels were reported in butter, margarine, and olive oil. Butter contained 13 /*g/kg and olive oil and margarine contained 7 /ig/kg. II. Pharmacokinetics A. Absorption 1. Approximately 50 to 62% of the respired PCE is absorbed by humans. B. Metabolism 1. The amount of PCE metabolized by P-450 varies depending on the animal species tested and concentration applied. The metabolite most often measured in urine is tricholoracetic acid (TCA). Other metabolites found are trichloroethanol, oxalic acid, carbon dioxide (COg), chloride, dichloroacetic acid, ethylene glycol, and, possibly, thioether. Up to 47% of the PCE inspired may be metabolized by P-450. Much of the PCE inspired by humans is unaccounted for in disposition studies. 1-12 SL 038888 C. Excretion DRAFT 1, Much of the PCE Inhaled is exhaled unchanged by lungs. Some PCE has been detected in urine as trichloro metabolites. PCE has a considerably longer half-life in humans than in rodents. D. Bioaccumulation 1. PCE accumulates in tissues following chronic exposure. E. Conclusions 1. Metabolism of PCE is complex and only partially characterized in humans. Particularly, the products and rates of metabolism are poorly understood. 2. Metabolism and storage are influenced by factors such as age, sex, exercise or workload, body mass and adipose tissue mass, and pulmonary dysfunction. III. Quantitative Risk Assessment A. Shape of the Dose-Response Curve SL 038889 1-13 DRAFT 1. Animal: The data for male mice liver tumors (adenomas and carcinomas combined) were consistent with a linear doseresponse relationship (applied dose). However, only three data points were available: a control and two exposures. 2. Human: NA. B. Range of extrapolation 1. Ratio of animal experimental concentration to ambient concentrations: Approximately 10. C. Range of Risks 1. The human risks associated with a continuous, lifetime exposure to perchloroethylene have been estimated using the linearized multistage model from animal carcinogenicity bioassays. Human risks estimated from animal data range from 31 to 144 x 10 Vppb or 5 to 21 x 10 V(J*g/m^). 2. The range of risks, estimated by fitting the multistage polynomial (Crump, 1981) to the 1986 NTP rat and mouse inhalation study ranges from 8.1-to-l.4-fold. MLE to 95% UCL: male mice -- 8.1-fold female mice -- 1.4-fold male rats -- 1.65-fold female rats -- 1.6-fold 1-14 SL 038890 DRAFT XV. National and International Evaluation A. U.S. Environmental Protection Agency 1. Genotoxicity Tests: Inadequate information to classify PCE as a mutagen or nonmutagen. 2. Animal carcinogenicity tests: Prior to the NTP (1986) inhalation bioassay EPA determined there was limited evidence of animal carcinogenicity (EPA, 1985a). EPA is currently reevaluating that position. 3. Human Evidence: carcinogenicity. Insufficient data to assess human A. The EPA placed PCE in the fourth quartile of 53 carcinogens for which EPA has calculated potencies. 5. Conclusions: PCE is a possible human carcinogen. However, as a result of the NTP 1986 study, EPA is reviewing the classification of PCE. It is likely that they will conclude there is sufficient animal evidence and classify it as a probable human carcinogen. B. International Agency for Research on Cancer (IARC) Si 3SS9] 1-15 DRAFT 1. Animal carcinogenicity assays: Sufficient vidence of animal carcinogenicity by inhalation. 2. Human Evidence: carcinogenicity. Inadequate evidence of human 3. Conclusion: PCE is a possible human carcinogen, classified by IARC as a category 2B carcinogen (IARC, 1987). C. Conclusions 1. Inadequate information exists to clearly classify FCE as genotoxic or nongenotoxic. 2. Data from epidemiological studies are of very limited use for the risk assessment of PCE due to the presence of other causative factors. 3. PCE should be considered a potential human carcinogen. 4. Risk estimates should be based on "applied dose" due to the Absence of sufficient information in humans on metabolism of PCE. 5. The upper 95% Cl of human risks estimated from animal data range from 31 to 144 x 10 Vppb (5 to 21 x 10 *V(/ig/m3)). 1-16 SL 038892 2. PHARMACOKINETICS AND METABOLISM Perchloroethylene is readily absorbed through the lungs and gastrointestinal tract and, to a lesser extent, may be absorbed through the skin. Once in the body, PCE disperses into all tissues. Steady-state tissue concentrations are a function of the absorbed dose, partitioning factors, and pharmacokinetic properties, such as rate of metabolic conversion and elimination. The percentage of PCE absorbed by humans via inhalation has been reported at 50% (Ohtsuki et al., 1983) and 62% (Bolanowska and Golacka, 1972). The primary metabolic pathway of PCE is thought to involve oxidation to an epoxide as the first step, although this epoxide intermediate has never been isolated iQ vivo. The epoxide undergoes rearrangement to form trichloroacetyl chloride and, ultimately, trichloroacetic acid (TCA), which has been identified in urine (Yllner, 1961; Daniel, 1963; Moslen et al., 1977; Costa administered and Ivanetich, 1980). Studies in which radiolabeled PCE was % to animals have occasionally recovered oxalic acid as a significant urinary metabolite (Yllner, 1961; Dimitrieva, 1967; Pegg et al., 1979). Carbon dioxide is also commonly produced (Pegg et al., 1979; Schumann et al., 1980). Other metabolites detected include trichloroethanol, chloride, dichloroacetic acid, and ethylene glycol (Daniel, 1963; Dimitrieva 1967, Ikeda and Ohtsuji, 1972; Ikeda et al., 1972; Koppel et al., 1985; Monster et al., 1983; Yllner, 1961). This section presents an overview of published studies on the absorption, distribution, metabolism, and elimination of PCE. Studies that have defined SL 038893 2-1 DRAFT the rate and extent of each of these processes in humans and in rodents are emphasized. Proposed metabolic pathways are discussed in some depth because metabolism is responsible for the transformation of PCE to one or more reactive species. ABSORPTION This section reviews the relevant data on PCE uptake through ingestion, dermal absorption, and inhalation. Ingestion Absorption of PCE from the gastrointestinal tract has been measured indirectly as percent of dose recovered. The percentage of dose recovered after administration of PCE is similar in mice and rats, varying between 80 to 100%. Peak blood concentrations measured in rats one hour after a 500 mg/kg dose shows PCE is absorbed rapidly (Pegg et al., 1979). Little information exists regarding amount or rate of oral absorption of PCF in humans. Koppel and associates (1985) reported that the blood concentration of PCE following an oral dose of 400 mg was described by a two-compartment model with half-lives of 160 minutes and 33 hours (unpublished data cited in Koppel et al., 1985). The same authors (Koppel et al., 1985) measured the concentration of PCE in blood at 21.5 j*g/raL within one hour of ingestion of 12 to 16 grams. Although far from definitive, these reports suggest fairly rapid and complete oral absorption SL 038894 2-2 Dermal absorption DRAFT Jakobsen and co-workers (1982) measured the absorption of PCE through guinea pig skin. Animals were in contact with liquid PCE for 6 hours. During the exposure, blood PCE concentrations rose rapidly, and peaked within 30 minutes Tsuruta (1975) estimated the rate of absorption of PCE through mouse 2 skin to be 24 nmol/min-cm of skin. Percutaneous penetration of PCE vapor in humans exposed to ambient air concentrations of 600 ppm is approximately one percent of pulmonary absorption (Riihimaki and Pfaffli, 1978). Studies in which volunteers immersed their thumbs in liquid PCE measured a peak concentration of 0.3 ppm in expired air within 40 minutes; concentrations decreased thereafter. PCE has a relatively slow rate of elimination in breath. The relatively low concentration of PCE in exhaled air suggests that dermal absorption is limited (Hake and Stewart, 1977; Stewart and Dodd, 1964) . Pulmonary Uptake PCE is effectively absorbed through the lungs during inhalation. Yllner (1961) reported the average pulmonary absorption of mice exposed to approximately 2000 ppm (reported as 1.3 mg PCE/g of mouse in a 2.7 liter chamber) to be 70%. In these animals, absorption varied from 42 to 87%. In a study by Pegg and co-workers (1979), the peak blood concentration of PCE in rats during a 6-hour exposure to 600 ppm was approximately 10 pg/mL. During inhalation exposure, PCE diffuses across the lungs and dissolves into the bloodstream. The rate of transfer depends on the blood/gas partition SL ^^895 2-3 coefficient for PCE. One estimate placed the human blood/gas coefficient of PCE at 16 (Monster et al., 1979), while Gargas and colleagues (1986) reports this coefficient value at 10.3. Both values reflect the fact that PCE is lipophilic and readily diffuses into the blood. The uptake of PCE by the lungs is also determined by the alveolar ventilation rate, exposure concentration, exposure duration, and metabolism. Several experimental human studies have quantified PCE absorption in terms of "percent uptake" (one minus the ratio of alveolar to ambient air concentration and multiplied by 100%). However, unless a steady state absorption rate is approximated at very low exposure concentration (near 1 ppb), the net quantities of PCE retained following environmental exposure may not be accurately quantified. The percentage of PCE absorbed in humans through the lungs has been estimated at 62% (Bolanowska and Golacka, 1972). After a 6-hour exposure of five volunteers to 54 ppm PCE (reported as 0.3? mg/L), retention of PCE reportedly stabilized after 1 1/2 hours and ranged from 55 to 70%. The report of a steady state condition by Bolanowska and Golacka (1972) differs from the data of Fernandez and associates (1976) who studied humans exposed to 100 ppm PCE. The latter investigators found that alveolar air concentrations continued to steadily increase after 8 hours although the air level had reached 55 ppm. Ohtsuki and co-workers (1983) estimated that the percentage of PCE absorbed by humans was 50%, although the specific derivation of the value was not described. Monster and colleagues (1979) observed an inverse relationship between uptake and exposure duration in humans over the course of a 4-hour inhalation exposure to 72 or 144 ppm PCE. Th net uptak at the end of 4 2-4 6V DRAFT hours was appr ximately 60% of that during the first hour. This observation indicates that net uptake decreases as blood and tissue concentrations of PCE equilibrate with PCE in the air space of the lungs. Net uptake is affected by differences in ventilation rate. When volunteers were exposed to 142 ppm while under a work load (i.e., an increased ventilation rate), the uptake of PCE increased to over two times what it was at rest (Monster et al., 1979). Fernandez and co-workers (1976) exposed humans to 100 ppm PCE for 8 hours and measured the concentration in alveolar air. Alveolar concentration rose rapidly in the first half hour and then continued to increase throughout the experiment although at a slower rate, reflecting a sustained decline in percent uptake observed. At steady state, the amount of PCE taken up or retained will be equal to the amount of PCE metabolized. As pointed out by Guberan and Fernandez (1974), the alveolar concentration of PCE rises more rapidly during the first phase of uptake. PCE concentrations in rapidly perfused tissues will reach approximate steady state concentrations long before they are achieved in poorly perfused or highly lipophilic tissues. For example, during an 8-hour exposure to 100 ppm, tissues such as the liver will reach half of their maximum concentration in 34 minutes, while adipose tissue will reach half of its maximum concentration in 3 1/4 hours. Following exposure, PCE may continue to perfuse into adipose tissue and excretion from this tissue will be slow (Guberan and Fernandez, 1974). Since the uptake or retention rate varies as a function of tissue, time and a number of other variables, the results in these studies relating uptake to metabolism may overestimate SL 038897 2-5 metab lism since the quantity of PCE stored in adipose tissue has not been accounted for. Table 2-1 summarizes the absorption and recovery of experimentally administered PCE in animals. Table 2-2 lists parameters of absorption, metabolism, and disposition of PCE in humans. DISTRIBUTION AND BIOACCUMULATION PCE diffuses into the bloodstream and distributes to tissues, primarily to organs and fat. Seventy-two hours after oral or inhalation exposure to tetrachloro[^C]ethylene, measurable radioactivity was found in the liver, kidneys, fat, lungs, heart, and adrenal glands of rats. The major part of the radioactivity was concentrated in the liver, kidneys, and fat (Pegg et al., 1979). A similar distribution was observed by Frantz and Watanabe after PCE was administered to rats in a saturated drinking-water solution (containing approximately 150 ppm PCE) (Frantz and Watanabe, 1983). The binding of metabolites of PCE to hepatic macromolecules has been measured by Pegg and associates (1979), Schumann and co-workers (1980), and Mitoma and colleagues (1985). Savolainen and associates (1977) gave rats 200 ppm PCE for 6 hour/day for 4 days, and observed substantial levels of PCE in perirenal fat within 17 hours after the end of the exposure (Savolainen et al., 1977). Information on the distribution of PCE in humans comes largely from reports of accidental xposures (Stewart et al., 1961a; Stewart, 1969; Hake and 2-6 038898 T tou OCdO lOo Parameter Value TABLE 2-1. ABSORPTION AMD RECOVERY OF TETRACHLOROETHYLEHE (PCE) IK ANIMALS Species (weight) Route Concentration or dose Exposure duration Reference DRAFT 2-7 Percent absorbed (average) Percent of gose recovered , (48 h post exposure) Percent of dose recovered (72 h post exposure) Percent of dose recovered (48 h post exposure) Percent of dose recovered (72 h post exposure) Percent of dose recovered (72 h post exposure) Percent of dose recovered (72 h post exposure) 70.0 (range 42 to 87) House Inhalation 79.65 94.6 Mouse Oral Mouse (18-31 g) Oral 84.4 103.0 Rat Rat (250 g) Oral Oral 91.2 Rat (250 g) Oral 100.0 Rat (275 to 285 g) Oral, in drinking water 1.3 mg/g 900 mg/kg 500 ng/kg 1000 ng/kg 1 ng/kg 500 mg/kg 8.1 ng/kg (approx.) 2 h Yllner, 1961 5 d/wk; 4 uk Mitona et al., 1985 Single exposure Schumann et at., 1980 5 d/wk; 4 uk Mitona et al., 1985 Single exposure Pegg et at., 1979 Single exposure Pegg et at., 1979 12 h Frantz and Uatanabe, 1983 DRAFT Parameter Vatue Species (weight) TABLE 2-1. (Continued) Route Concentration or dose Exposure duration Percent of dose recovered (72 h post exposure) Percent of dose recovered (72 h post exposure) P rcent of dose recovered (72 h post exposure) P rcutaneous absorption rate Time to peak blood concentration T isas to peak blood concentration Total pulmonary uptake Total pulmonary uptake 99.9 Rat (250 g) 100.0 Rat (250 g) 80.0 Mouse 24 n/aol/min/cm^ Mouse 1h (approx.) Rat (250 g) 0.5 h 0.40 mg/enimal (16.5 ag/kg) 1.48 mg/animal (5.9 ag/kg) Guinea pig Mouse Rat Inhalation Inhalation Inhalation Dermal Oral Dermal Inhalation Inhalation 10 ppm 6h 600 ppm 6h 10 ppm 6h 500 ag/kg 10 ppm 10 ppm Single exposure 6h 6h 6h Reference Pegg et at., 1979 Pegg et al., 1979 Schumann et at., 1980 Tsuruta, 1975 Pegg et al., 1979 Jakobsen et al., 1982 Pegg et al., 1979 Pegg et al., 1979 6ftt OO Parameter Value Species (weight) TABLE 2-1. (Continued) Route Concentration or dose Exposure duration Reference Total pulmonary qptake 77.5 mg/animal (310 mg/kg) Rat Inhalation 600 ppm 6h Pegg et al., 1979 --a------------------------- Percent of dose recovered refers to the portion of adainistered dose of PCE accounted for by measurements of breath, metabolites, excreta, and tissue. The percent recovered is considered to be indicative of percent absorption; however, the two parameters are not identical (D'Souza et el., 1985; U.S. EPA, 1985a). DRAFT Parameter Percent absorbed through lungs Value SO Percent absorbed through lungs 62 to 64 TABLE 2-2. ABSORPTION, METABOLISM, AND DISPOSITION OF PCE IN HUMANS Route Concentration or dose Duration Reference Inhalation Ohtsuki et at., 1983 a Inhalation 72 or 144 ppm 4h Monster et al., 1979 Average pulmonary uptake 1350 mg Inhalation 150 ppm Sh Fernandez et al., 1976 DRAFT 038902 Average pulmonary intake Average pulmonary iptake 455 wo 945 mg Inhalation Inhalation 72 ppm 144 ppm 4h 4h Monster et al., 1979 Monster et al., 1979 Percent retention 55 to 70 Inhalation 54 ppm (390 mg/L) 4-6 h Time to peak concentration - exhaled air - lung tissue - whole body - organs - fat - blood 15 to 40 min 8h 8h 8h 12 h 0.5 h Dermal Inhalation Inhalation Inhalation Inhalation Inhalation 100 ppm 100 ppm 100 ppm 100 ppm 72 to 142 ppm 10 min 8h Bh 8h 8h 4h Bolanouska and Golacka, 1972 Stewart and Dodd, 1964 Guberan and Fernandez, 1974 Cuberan and Fernandez, 1974 Guberan and Fernandez, 1974 Guberan and Fernandez, 1974 Monster et al., 1979 DRAFT SL 038903 Parameter Percent metabolized Percent metabolized Percent eliminated unchanged via lungs Percent eliminated tnchartged via lungs Value 1 to 3 Route Inhalation TABLE 2-2. (Continued) Concentration or dose Duration 50 ppm 8h Reference Ikeda et al., 1972 Ohtsuki et al., 1983 A7 Inhalation 57 ppm 4-6 h (390 mg/L) Bolanowska and Golacka, 1972 38 Inhalation 50 ppm 8h ikeda et al., 1972 Ohtsuki et al., 1983 25 Inhalation 57 ppm 4 to 6 h (390 mg/L) / Bolanowska and Golacka, 1972 Body burden 1000 mg Inhalation 100 ppm Adipose burden 500 mg (approximate) Inhalation 100 ppm b Predicted biological t 1/2 Respiratory t 1/2 71.5 h 65 h Inhalation Inhalation 100 ppm 100 ppm Urinary metabolite t 144 h Inhalation 10 to 100 ppm 8h 8h 8h 7 h (single exposure) 8 h/d;5 d Gbfceran and Fernandez, 1974 Guberan and Fernandez, 1974 Guberan and Fernandez, 1974 Ikeda and tmamura, 1973 (based on Stewart et al., 191 Ikeda and Imamura, 1973 Parameter Time to stead/-state equilibrium, adipose tissue Saturation of metabolism Value 125 h c NA Route Inhalation TABLE 2-2. (Continued) Concentration or dose Duration 100 ppm 8 h/d Inhalation 100 ppm 8h Elimination t 160 min 33 h Oral 400 mg "8--------------------------------------------------------------- ^Values calculated fro* average estimated uptake/min. t denotes half-life. HA: not applicable. Single exposure Reference Honster et at. , 1979 Ohtsuki et al. , 1983 Koppel et al.. 1965 DRAFT 2-12 (vPO) V DRAFT nervous system (CNS) depression, cardiac arrhythmias, alteration of kidney function, and liver injury. These observations provide indirect evidence that PCE distributes to the nervous system, liver, and kidneys. Additionally, tissue concentrations in the liver and brain have been measured at levels ten to 50 times greater than those in blood following fatal intoxication (Lukaszewski, 1979). Perchloroethylene is a relatively stable molecule and is metabolized slowly. Perchloroethylene is soluble in lipids and this factor, along with its slow rate of metabolism lead to its accumulation in tissue following repeated exposure (Filser and Bolt, 1979; Loew et al., 1983). On the basis of a model of PCE uptake, distribution, and elimination, Guberan and Fernandez (1974) predicted that the solvent would disperse primarily to three body compartments: adipose tissue, muscle, and tissues rich in blood vessels. It was estimated that the body burden of a 70-kg man exposed to 100 ppm PCE for 8 hours would be 1000 mg. More than one half of the PCE would accumulate in fat; substantial amounts would also be concentrated in muscles. The rate of uptake and the time required for equilibration of tissue concentrations (for a given concentration of PCE in inspired air) depend on the solubility of the substance, the volume of each tissue, and on the rate of blood flow through the tissue. Because the blood supply to adipose tissue is less than to other tissues and because PCE is more soluble in fat than in blood, it takes a correspondingly greater time for equilibrium concentrations to be reached. Assuming a single 8-hour exposure to PCE, the 038905 2-13 DRAFT time required for fat to reach about 50% of its capacity is 25 hours (Monster et al., 1979; Guberan and Fernandez, 1974). Once whole-body, steady-state concentrations are reached (compared to concentrations in air), the amount of PCE in each tissue depends on the tissue/blood partition coefficient. Since PCE is lipophilic, the adipose/blood partition coefficient is the highest of any tissue type (at 37 C the adipose/blood partition coefficient is about 107) (Guberan and Fernandez, 1974). Guberan and Fernandez state that daily occupational exposure to PCE at 100 ppm, 8 hours/day, would lead to accumulation of PCE in fat. Savolainen and co-workers noted accumulation of PCE in the blood, liver, fat, kidneys, and brain of rats following 5 days of inhalation exposure at 200 ppm (Savolainen et al., 1977) . The PCE levels in perirenal fat, brain, and lungs rose continuously during the experiment. Concentrations in blood and liver also increased, but the rate of accumulation slowed by the third day. Evidence of PCE accumulation is also available from a study that measured the concentration of PCE in exhaled air. When human volunteers were exposed to 100 ppm, 7 hours/day for 5 days, the concentration of PCE in expired air increased with each exposure. This suggests that PCE had accumulated in the tissues on each consecutive day of exposure and had not been eliminated between exposures (Hake and Stewart, 1977). 2-14 SL 038906 METABOLISM AND ELIMINATION DRAFT Among the most important enzyme systems for metabolism of toxic substances are the mixed-function oxygenases (MFO). These enzymes are concentrated in the liver, kidneys, lungs, and skin, and are present in other tissues as well. Mixed-function oxygenases catalyze the addition of oxygen to compounds, which facilitates their excretion from the body. Oxidation of a compound can function as a mechanism of detoxification or can transform it to a reactive (toxic) substance. Evidence that MFO's are directly involved in the metabolism of PCE comes from the work of several investigators. Moslen and colleagues (1977) demonstrated that a number of substances, including phenobarbital (PBT) and Aroclor 1254, induced hepatic MFO. Pretreatment of rats with either of these compounds, followed by administration of PCE, increased the metabolism of PCE five to seven times over controls. Costa and Ivanetich (1980) showed that substances that inhibit cytochrome P-450, a component of MFO's, also inhibit metabolism of PCE in rats. Induction of cytochrome P-450 by PBT or pregnenolone-16a-carbonitrile increased the metabolism of PCE. The solvent was also shown to bind to the active site of P-450 in rat hepatic microsomes. The first step in the general metabolism of PCE is thought to be transformation to an epoxide by the MFO, although this epoxide has never been isolated in vivo (Bonse et al., 1975; Greim et al., 1975). The epoxide of PCE apparently rearranges spontaneously with migration of a chlorine to 2-15 DRAFT form triehloroacetyl chloride and, ultimately, trichloroacetic acid (Moslen et al., 1977; Leibman and Ortiz, 1977; Reichert, 1983). Trichloroethanol has also been reported as a metabolite; however, a pathway has not been proposed that explains its formation (Ikeda and Ohtsuji, 1972; Ikeda et al., 1972; Monster et al., 1983; Koppel et al., 1985). As yet, it is not clear if conjugation (phase XI) reactions are involved in the metabolism of PCE (Pegg et al., 1979; Lafuente and Mallol, 1986). Although not consistently reported, the production of oxalic acid and CC^ as metabolites of PCE has been documented in rodents (Yllner, 1961; Pegg et al., 1979; Schumann et al., 1980). Their formation argues for a second and possibly minor pathway of oxidative metabolism for rats and mice. Xt has been proposed that oxalic acid and CO^ are formed as end-products of a metabolic pathway that also includes epoxide formation as the first step (Pegg et al., 1979). In this scheme, chloroethylene glycol is formed from the epoxide by the action of epoxide hydrolase. This reaction is followed by hydrolysis to oxalic acid and/or decarboxylation to CO^ and possibly formic acid. Chloride, dichloroacetic acid,and ethylene glycol have also been reported as urinary metabolites of PCE in rodents (Daniel, 1963; Yllner, 1961; Dimitrieva, 1967). Xt appears that these substances are minor metabolites, since each has been reported by only one investigator. Little is known about their formation. Figure 2-1 shows the structure of metabolites as well as the proposed metabolic pathways. 2-16 SL 038908 DRAFT There are some similarities in the urinary metabolites produced by humans and rodents following exposure to PCE. Trichloroacetic acid has been identified in all species. Tables 2-3 and 2-4 list the metabolites identified in humans, mice, and rats, as well as the conditions of exposure. Yllner was the first to analyze urinary metabolites of PCE in mice. Although 18% of the radiolabeled compound was not accounted for, 52% of the portion metabolized was recovered as trichloroacetic acid (TCA), 11% as oxalic acid, and a trace amount 'as dichloroacetic acid (Yllner, 1961). Daniel fed labeled PCE to rats and found TCA (0.6%) and inorganic chloride as the only metabolites (Daniel, 1963). Total urinary metabolite production in rats was measured by Moslen and co-workers (1977). They did not identify the actual metabolites, with the exception of TCA (which was the major metabolite produced). Pegg and associates and Dimitrieva identified oxalic acid as the primary product of PCE metabolism in rats (Pegg et al., 1979; Dimitrieva, 1967). Carbon dioxide has been recovered as a metabolite of PCE in mice (Schumann et al., 1980) and in rats (Pegg et al., 1979). Trichloracetic acid, trichloroethanol, and an unidentified organic chloride have been measured in urine from humans exposed to PCE. Based on the Yllner (1961) study in mice, Ogata and colleagues measured only trichloro compounds in the urine of human volunteers (Ogata et al., 1971). In this study the major metabolite was an unidentified organic chloride. In some instances, identification of trichloroethanol has been by the Fujiwara reaction (Ikeda et al., 1972; Ikeda and Ohtsuji, 1972). The results of this test are qualitative and the accuracy is questionable. However, Monster and co workers (1983) and Koppel and associates (1985) determined the presence of SL 038909 2-17 Cl ^ /C' _ C=C a-" Cl Ttrachlootitiyl*n* Mta<MtfnciiOn saidlKi (Oj.NAOPH) 'S*^' Cl ^ Cl Tttrachleaottxvkn* o*da Chlototthylm* 9iyto* Hydrolywt [-<1CO, + OH] Carbon d'Oaida Forme acid Oxalic add Cl----- C----- C, I Cl Cl Trichloroaeotyl eMondo Myd*oly*< Cl ICl- -- C N OH Cl Trichleroc*t>e acid Figure 2*1. Metabolic pathways of PCE (brackets denote compounds that have not been isolated in vivo) (Daniel, 1963; Pegg, 1979; Costa and Ivanetich, 1980). Although trichloroethanol, dichloroacetic acid, ethylene glycol, and a thioether have been identified as urinary metabolites, the route(s) by which each formed has not been characterized. 2-18 SL 038910 DRAFT SL 038911 Metabolite Trichloroacetic acid TABLE 2-3. PCE METABOLITE PRODUCTION IN RODENTS Species (no. of animals) Concentration Route of or dose exposure Duration of exposure Mouse (5) 1-3 mg/g Inhalation 2h Rat (6) 1000 mg/kg Oral Single exposure Rat (NA) 5 mg/L Inhalation 5 h/d for 3 d Rat (8) 200 ppm Inhalation 8h Rat (7) 2.78 mmol/kg Intraperitoneal Single exposure Mouse (1) 200 ppm Inhalation 8h Mouse (2) 2.78 mmol/kg Intraperitoneal Single exposure Rat (M 0.75 raL/kg Gavage Single exposure Mouse (4 to 24 per dose level) 20 to 2000 rag/kg Gavage 5 d/wk for 6 wk Reference Yllner, 1961 Daniel, 1963 Dimitrieva, 1967 Ikeda and Ohtsuji, 1972 Ikeda and Ohtsuji, 1972 Ikeda and Ohtsuji, 1972 Ikeda and Ohtsuji, 1972 Moslen et al., 1977 Buben and 0'Flaherty, 1985 Trichloroethanol Rat (8) Mouse (1) Rat (?) Mouse (2) 200 ppm 200 ppm 2.78 mmol/kg 2.78 mmol/kg Inhalation Inhalation Intraperitoneal Intraperitoneal 8h 8h Single exposure Single exposure Ikeda and Ohtsuji, 1972 Ikeda and Ohtsuji, 1972 Ikeda and Ohtsuji, 1972 Ikeda and Ohtsuji. 1972 DRAFT 038912 cn Metabolite Oxalic acid Species {no. of animals) Mouse (5) Rat (NA) Rat (3) Rat (3) TABLE 2-3. (Continued) Concentration or dose Route of exposure 1.3 mg/g Inhalation 5 mg/L Inhalation 10 or 600 ppm Inhalation 1 or 500 mg/kg Gavage Duration of exposure 2h 5 h/d for 3 d 6h Single exposure Reference Yllner, 1961 Dimitrieva, 1967 Pegg et al., 1979 Pegg et al., 1979 Carbon dioxide Rat (3) RaL (3) Mouse (31 Mouse (3) 10 or 600 ppm Inhalation 1 or 500 mg/kg Gavage 10 or 600 ppm Inhalation 1 or 500 rag/kg Gavage 6h Single exposure 6h Single exposure Pegg et al., 1979 Pegg et al., 1979 Schumann et al., 1980 Schumann et al., 1980 Dichloroacetic acid Mouse (5) 1.3 mg/g Inhalation 2h Yllner, 1961 Ethylene glycol Rat (NA) 5 mg/L Inhalation 5 h/d for 3 d Dimitrieva, 1967 Chloride Rat (6) 1000 rag/kg Oral Single exposure Daniel, 1963 DRAFT SL 038913 Metabolite Trichloroacetic acid TABLE 2-4. PCE METABOLITE PRODUCTION IN HUMANS No. of individuals Concentration or dose Route of exposure Duration of exposure 85 10 to 400 ppm Inhalation 8 h/d, 6 d/wk 5 390 mg/L Inhalation 6 h (single exposure) 4 20 to 70 ppm Inhalation Daily (Intermittent) 66 200 to 400 ppm Inhalation Daily (intermittent) 24 100 to 200 ppm Inhalation 1 to 8 h 6 70 or 140 ppm Inhalation 4 h (single exposure) 5 7 ppm3 Inhalation 37 to 54 h/wk 9 47 ppma Inhalation 37 to 54 h/wk 9 53 ppm3 Inhalation 37 to 54 h/wk 1 8 to 10 mL Oral Single exposure Reference Ikeda et al., 1972 Bolanowska and Golacka, 1972 Ikeda and Ohtsuji, 1972 Ikeda and Ohtsuji, 1972 Fernandez et al., 1976 Monster and Houtkooper, 1979 Monster et al., 1983 Monster et al., 1983 Monster et al., 1983 Koppel et al., 1985 Trichio roe thano1 85 4 66 10 to 400 ppm 20 to 70 ppm 200 to 400 ppm Inhalation Inhalation Inhalation 8 h/d, 6 d/wk Daily (intermittent) Daily (intermittent) Ikeda et al., 1972 Ikeda and Ohtsuji, 1972 Ikeda and Ohtsuji, 1972 Metabolite TABLE 2-4. (Continued) No. of individuals Concentration or dose Route of exposure Duration of exposure 24 100 to 200 ppm Inhalation Daily (intermittent) 6 70 or 140 ppm Inhalation 4 h (single exposure) 5 17 ppma Inhalation 37 to 54 h/wk 9 4> 7-i ppma Inhalation 37 to 54 h/wk 9 5c3, ppma Inhalation 37 to 54 h/wk Reference Ikeda and Ohtsuji, 1972 Monster and Houtkooper, 1979 Monster et al., 1983 Monster et al., 1983 Monster et al., 1983 Thioether derivative 6 15 t^ 50 ppm Inhalation 8 h/d, 5 d/wk Lafuente and Mallol, 1986 "Values represent the median of the time-weighted-average exposures for individuals. DRAFT cn r* vteseo trichloroethanol by gas chromatography. Still other studies have failed to detect trichloroethanol; therefore, it is far from clear under what circumstances this substance is formed (Monster et al., 1979; Fernandez et al. , 1976). Measurements of human metabolite production are confounded by the fact that CC^, oxali-c acid, and chlorine are normal products of metabolism. Their presence as products of PCE metabolism could be quantified only by the administration of radiolabeled PCE--a procedure that has not been undertaken in humans. Furthermore, studies in humans have primarily focused on quantifying the amount of trichloro compounds detectable in urine and not on accounting for all the metabolic products of PCE. Lafuente and Mallol have reported the presence of a thioether derivative of PCE in the urine of women occupationally exposed to PCE (Lafuente and Mallol, 1986). A. gradual increase in thioether production was observed over the course of a week. This increase appears to have been associated with continued exposure to PCE (with concomitant accumulation) over the work week. However, measured amounts of thioether in exposed women were reported not to be statistically different from levels found in nonexposed individuals. The recent identification of thioether derivatives in humans exposed to PCE, coupled with problems associated with accurate and complete identification of metabolites, and the fact that the bulk of an absorbed human dose has not been accounted for indicates that not all human metabolites of PCE have been identified. Characterization of PCE metabolites in rodents may be incomplete as well, since no complete mass-balance studies have been SL 038915 2-23 conducted. Although the major metabolites recovered in all species are presumably produced by the similar enzymes, many uncertainties remain. Based on present information, humans and rodents appear to metabolize PCE in qualitatively similar ways. Perchloroethylene is eliminated from the body by two major processes: metabolism followed by excretion of urinary metabolites and pulmonary elimination of unchanged PCE. Although some PCE may be eliminated through the skin, preliminary measurements indicate that this is a minor route in humans (Bolanowska and Golacka, 1972). Except for the Schumann and associates (1980) study, most of the PCE systemically absorbed under experimental conditions (e.g., in mice, rats, and humans) was eliminated unchanged in expired air, so that metabolic degradation and subsequent elimination appear to account for less than 50% of absorbed PCE. Experimental data reviewed below Indicate that PCE metabolism is dosedependent and saturable, and that the amount of PCE metabolized appears to be species-dependent as well. When the production of urinary metabolites was measured in mice exposed for two hours to approximately 2000 ppm (reported as 1.3 mg/g of PCE in a 2.7 liter chamber), Yllner found that only 2% of an inhaled dose was excreted by this route (Yllner, 1961). Seventy percent of the parent compound was recovered in expired air. Pegg and colleagues (1979) compared the metabolism of PCE in Sprague-Dawley rats at different doses and routes of exposure. The PCE was administered by gavage (1 or 500 mg/kg) or by inhalation (10 or 600 ppm). The primary route of elimination of PCE occurred through the lungs as the unmetabolized parent compound. Urinary 2-24 SL 038916 DKAH excretion of metabolites accounted for the majority of the remaining PCE. The percentage of PCE metabolized was dose dependent: elimination of unmetabolized PCE in expired air following oral exposure increased from 72% after 1 mg/kg to 90% after a dose of 500 mg/kg. There was a corresponding decrease in metabolites from 28 to 10%, An analogous pattern was seen after inhalation exposure. At 10 ppm, 68% of the dose was eliminated unchanged through the lungs and 32% was recovered as metabolites. Treatment at 600 ppm caused an increase in pulmonary elimination to 88% of the dose with a concomitant decrease in metabolites to 12%. Thus, it is likely that exposure to concentrations below 10 ppm would result in a decrease in the pulmonary elimination of PCE and in an increase in relative metabolism of PCE. Pulmonary elimination of PCE was linear and had a half-life of approximately 7 hours. The half-life was Independent of dose or route of administration. The half-life of PCE in blood was 6 and 7 hours, after oral and inhalation exposure, respectively. The pharmacokinetics of PCE in Sprague-Dawley rats and B6C3F1 mice were studied by Schumann and co-workers (1980). Rats were given a single oral dose of 500 mg/kg PCE or were exposed to 10 ppm 14C-PCE by inhalation. The authors reported that at 10 ppm the major route of elimination was excretion of unmetabolized PCE in expired air, although supporting data were not provided. Some PCE was metabolized, and metabolites were recovered in the urine. After an oral dose of 500 mg/kg of labeled PCE, radioactivity was measured in the expired air, urine, feces, and carcass. The relative importance of each route in the elimination of PCE was not discussed. SL 038917 2-25 DRAFT The elimination of PCE by B6C3F1 mice differed, depending on the dose and possibly on the route of administration from that found in Sprague-Dawley rats. Metabolism, with urinary excretion of PCE, was the primary route of elimination after inhalation exposure at 10 ppm; 62.5% of the dose was recovered as urinary metabolites and only 12% was excreted through the lungs. Eighty-three percent of a single oral dose (500 mg/kg) was eliminated through the lungs, while 10.3% appeared as urinary metabolites (Schumann et al., 1980). The fate of PCE in Sprague-Dawley rats fed PCE in their drinking water was reported by Frantz and Watanabe (1983). Animals were given PCE in a saturated solution over a 12-hour period, resulting in an average dos,- of 8.1 mg/kg. Treatment was followed by a 72-hour observation period prior to sacrifice. Most of the PCE (87.9%) was eliminated unmetabolized via the lungs. Although urinary excretion was the second largest route of elimination, the amount metabolized was relatively small (7.2%). The half- life of pulmonary elimination was 7.1 hours. This is nearly identicai to the value determined by Pegg and colleagues (1979) after oral doses of PCE were given to rats. Mitoma and co-workers studied the metabolic disposition of PCE in OsborneMendel rats and B6C3F1 mice (Mitoma et al., 1985). However, the relatively short follow-up period of 48 hours used in this study tends to underestimate the quantity of metabolites excreted. Substantial differences between species in the amount of PCE eliminated unchanged in expired air were noted. There were also differences in the percentage of dose metabolized. Rats eliminated 79% through their lungs and metabolized 5% of a 1000 mg/kg dose. 2-26 SL 038918 DRAFT Pulmonary elimination by mice (of a 900 mg/kg dose) was 57.5%, while 22% was metabolized. The actual amount of PCE metabolized (measured as mmol/kg) also differed by species. Mice metabolized approximately four times as much of the total dose of PCE as rats. However, in both species as the dose was quadrupled, the amount metabolized only increased about 2.5-fold. These data suggest that metabolism approaches saturation at high doses in both species. Conversely, at low concentrations the proportion of PCE metabolized increases. Buben and 0'Flaherty also demonstrated that oxidative metabolism of PCE in Swiss mice decreases with increasing dose and appears to be a saturable process (Buben and O'Flaherty, 1985). Animals received 0, 20, 100, 200, 500, 1000, 1500, or 2000 mg/kg-day of PCE by gavage and were followed for 72 hours post exposure. Trichloroacetic acid was measured as an index of PCE metabolite production; the estimated maximum rate of urinary metabolite excretion was 136 mg/kg-day. As the dose increased, the percentage of the dose metabolized decreased. At the lowest doses, approximately 25% of the PCE was metabolized. This amount decreased to 5% at the highest doses. Tables 2-5 and 2-6 summarize data on the metabolism of PCE in rats and mice, respectively. Table 2-7 lists some pharmacologic constants of PCE in rodents. The data reviewed above indicate that metabolism of PCE in both rats and mice displays saturable kinetics. The percentage of dose metabolized decreases as the dose is increased until the amount metabolized is no longer a function of the dose (a zero-order reaction). The extent of metabolism of PCE appears species-dependent. Rats consistently metabolize a relatively 038919 2-27 DRAFT smaller amount of PCE regardless of the route of administration. Mice typically metabolize a greater percentage of a dose than rats. Measurements of PCE metabolism in humans have many uncertainties associated with them. Total trichlorinated metabolites in urine have usually been determined colorimetrically and therefore are difficult to evaluate quantitatively. By this method, production of chlorinated metabolites is equated with the amount metabolized. This approach does not account for the possibility that some metabolites may be produced that are not chlorinated, such as CC^, the thioether, and oxalic acid. Studies of PCE metabolism in humans have consistently considered TCA to be a principal metabolite. However, a major purpose of the studies has been to ascertain if PCE exposure in the workplace can be quantified by measurements of trichloro compounds in urine. ' Thus, the studies do not reflect an attempt to measure all possible metabolites of PCE. Trichloroacetic acid production has also been measured to estimate the extent of metabolism of PCE and to characterize the kinetics of urinarv elimination. Volunteers exposed to 87 ppm of PCE for 3 hours excreted about 0.40 mg/h of TCA (Ogata et al., 1971). Monster and colleagues analyzed the TCA content of blood and urine from volunteers exposed to 72 or 144 ppm of PCE for 4 hours (Monster et al., 1979). The mean production of TCA (over a 70-hour period) was 6.0 and 11.0 mg, respectively. Blood levels of TCA increased over the course of the experiment and continued to rise until about 20 hours after the end of exposure. The TCA was eliminated from blood by first-order processes; the half-life of elimination was 65 to 90 hours. 2-28 SL 038920 DRAFT SL 038921 Route (n) Oral (3) TABLE 2-5. METABOLISM AND DISPOSITION OF TETRACHLOROETHYLENE IN RATS Post a-----------Percent metabolized Concentration Exposure exposure or dose duration period Urinary Total d metabolites metabolites Percent eliminated unchanged via lungs Reference 1 mg/kg Single exposure 72 h 16.5 28.5 71.5 Pegg et al., 1979 Inhalation (3) 10 ppm (12.2 mg/kg) 6h 72 h 18.7 32 68 Pegg et at., 1979 Oral (in drinking water) (4) 6 mg/kg 12 h 72 h 7.2 12 88 Frantz and Watanabe, 1983 Inhalation (3) 600 ppm (232 rng/kg) 6h 72 h 6 12 88 Pegg et al., 1979 Oral (3) 500 mg/kg Single exposure 72 h 4.6 10 90 Pegg et al., 1979 Oral C4) 1000 mg/kg 5 d/wk; 4 wk 42 h 2.4 5 79 Nitoma et al., 1985 a These values represent the percent of recovered dose that was metabolized, b n = number of treated animats, c ^For inhalation exposures, the dose in mg/kg was calculated assuming a breathing rate of 0.18 m3/d, and an average weight of 250 g per animal. These values represent metabolite recovery from all routes of excretion. DRAFT 2-30 SL 0 38922 Route a (n) Oral (12 15) Inhalation (3) Inhalation (N/A) Oral (12-15) Oral 02-15) Oral (3) Oral (4) Oral (12-15) TABLE 2-6. METABOLISM AMD DISPOSITION OF TETRACHLOROETHTLENE IN MICE Post Percent metabolized Concentration Exposure or dose duration exposure period Urinary Total cd metabolites metabolites Percent eliminated unchanged via lungs 20 mg/kg 5 d/wk; 6 wk 0h 25 e NA HA Buben and O'Flaherty, 1985 10 ppm 6 hr 72 h 62.5 88.0 12 Schumann et al., 1980 (29 mg/kg) 1.3 mg/g 2h 96 h 20 NA 70 Yllner, 1961 200 mg/kg 5 d/wk; 6 wk 0h 15.5 NA HA Buben and O'Flaherty, 1985 500 mg/kg 5 d/wk; 6 wk 0h 12.6 NA NA Buben and O'Flaherty, 1985 500 mg/kg Single exposure 72 h 10.3 17.4 82.6 Schimam et al., 1980 900 mg/kg 5 d/wk; 4 wk 48 h 14.4 22.2 57.5 Mitoma et al., 1985 1000 mg/kg 5 d/wk; 6 wk 0h 8.1 HA NA Buben and O'Flaherty, 1985 SI* 038923 Route a <n) TABLE 2-6. (Continued) Post Percent metabolized Concentrationi Exposure or dose duration exposure Urinary c Total d Percent eliminated period metabolites metabolites unchanged via lungs Reference Oral (4-6> 2000 mg/kg 5 d/wk; 6 uk 0h 5.1 NA NA Buben and 0`Flaherty, 1985 --a" n = number of treated animals. b3 ^for inhalation exposure(s), the dose in mg/kg was calculated assuming a breathing rate of 0.0345 m id, and an average weight of 20 g per animal. These values represent recovery of urinary metabolites only. Cl These values represent recovery from all routes of exposure, with the exception of expired tetrachioroethylene. e NA; not available. DRAFT 2-31 t Parameter Value TABLE 2-7. PHARMACOLOGIC CONSTANTS OF TETRACttLOROETHTLENE IN RODENTS Species (Height) Route Concentration or dose Exposure duration Reference DRAFT SL 038924 Blood elimination t * 1/2 a Blood elimination t 1/2 6h 7h Rat (250 g) Rat (250 g) Oral 500 mg/kg Inhalation 600 ppm Single exposure 6h Pegg et al.. 1979 Pegg et al.. 1979 Pulmonary elimination a *1/2 7h (approx.) Rat (250 g) Oral and inhalation 1 or 500 mg/kg and 10 or 600 ppm Single exposure 6h Pegg et al.. 1979 Blood k (elimination constant) 0.12 h-1 Rat (250 g) Oral 500 mg/kg Single exposure Pegg et al.. 1979 rio OJ Blood k 0.10 h-1 r Rat Inhalation 600 ppm ro (elimination constant) (250 g) 6h Pegg et al.. 1979 Pulmonary k eq (elimination constant) 0.10 h-1 Rat (250 g) Oral and inhalation 1 or 500 mg/kg and 10 or 600 ppm Single exposure Pegg et al.. 1979 V (maxi mu* amouit of mex urinary metabolites formed and excreted in 24 h) V (maximal velocity metabolic elimination) 136 mg/kg-d /7 mmol/h-kg Mouse Oral Rat (200 to 250 g) Inhalation 20 to 2000 mg/kg 5 d/uk; 6 uk 1000 ppm (initial concentration) 12 h Buben and O'Flaherty, 1985 Filser and Bolt, 197 K (equilibrium eq constant between gas phase and an nae 1 tissue) 105 Rat 200 to 250 gi Inhalation 1000 ppm 12 h Filser and Bolt, 1979 2-33 SL 038925 Parameter Value Specles (weight) Body burden 5.9 mg/kg Body burden 310 mg/kg Body burden 16.5 mg/kg "aa ;-------------------------------------------- ,s ***e hlf-time of elimination. Rat (250 g) Rat <250 g) House TABLE 2-7. (Continued) Route Concentration or dose Exposure duration Inhalation 10 ppm 6h Inhalation 600 ppm 6h Inhalation 10 ppm 6h Reference Pegg et al., 1979 Pegg et at., 1979 Schumann et al., I960 50 -n DRAFT Urinary eliminati n of TCA followed the disappearance of TCA from blood. Fernandez and associates measured excretion of TCA from individuals exposed to 150 ppm for 8 hours (Fernandez et al., 1976). Over a 72-hour collection period, the average amount of TCA produced was 25 mg. This is equivalent to about 0.34 mg of TCA per hour. Ikeda (1977) and Ikeda and Imamura (1973) qualitatively measured the half-life of urinary trichloro compounds and calculated the mean half-life to be 144 hours (range of 123 to 190 hours). The lengthy half-life may result from the continued formation of metabolites from PCE mobilized from tissues where it has accumulated. Measurements of TCA production suggest that human metabolism of PCE via this pathway represents only a small portion of the absorbed dose. For example, the values published by Monster and associates (6.0 and 11.0 mg TCA after a 4-hour exposure to 72 or 144 ppm PCE) respectively, represent only about 1 to 2% of the estimated absorbed dose (Monster et al., 1979). Ogata and ccworkers reported that TCA excretion was equivalent to 1.8% of the retained PCE with total excretion of organic chloride accounting for only 2.8% of the retained dose (Ogata et al., 1971). These data agree with those of Fernandez and colleagues, who estimated that 1350 mg of PCE would be absorbed after exposure to 150 ppm PCE for 8 hours (Fernandez et al., (1976). Ikeda and co-workers and Ohtsuki and associates have estimated that only about 2% of an 8-hour exposure to 50 ppm PCE would be metabolized and that 38% would be eliminated through the lungs unchanged by the end of the exposure period; the remaining 60% of the inhaled dose was hypothesized to be stored in the body and available for subsequent metabolism and/or pulmonary elimination (Ikeda et al., 1972; Ohtsuki et al., 1983). Ohtsuki and colleagues found that human urinary metabolite production did not appear 2-34 SL 038926 DRAFT to be linearly related to exposure concentration. A graph of total trichloro compounds (from urine) plotted against PCE concentrations in air showed that metabolite production appeared to be dose - dependent, leveling off at approximately 400 ppm PCE (8-hour exposure). This suggested metabolic saturation; however, no statistical test of departure from linearity was performed in this study and a questionable nonlinear model was assumed. Bolanowska and Golacka proposed that at steady state approximately 62% of respired PCE remains in the body, (based on their measurements of inspired and exhaled PCE) (Bolanowska and Golacka, 1972). Of the retained dose an estimated 25% is later exhaled. Thus, 75% of the retained dose, or 47% of the respired dose, is available for metabolism. At steady state, absorbed dose equals metabolized dose. However, since the steady state probably is not been attained, less than 47% of the respired dose is metabolized, with the balance representing the quantity of PCE accumulating in adipose tissue. These calculations are much higher than those obtained by other studies of PCE uptake and metabolism in humans (Ogata et al., 1971; Ikeda et al., 1972; Fernandez et al., 1976; Ikeda, 1977; Monster et al., 1979; Ohtsuki et al., 1983). However, the difference in large part may be due to Bolanowska and Golacka's consideration of total metabolism and not focusing solely on trichloro compounds as other investigators have done. Ogata and co-workers only analyzed for the excretion of trichloroacetic acid and an unidentified organic chlorine compound (Ogata et al., 1971). Ikeda and associates only measured the production of trichloro compounds and creatinine (Ikeda et al., 1972) Fernandez and co-workers measured only the excretion of TCA (Fernandez et al., 1976). Monster and fellow researchers analyzed urine for SL 038927 2-35 draft TCA and trichloroethanol (Monster et al., 1976). Ohtsuki and colleagues measured only total trichloro compounds in the urine (Ohtsuki et al., 1983). In fact, Bolanowska and Golacka reported that trichloro compound excretion in urine accounted for a few percent of the PCE retained, which is similar to results reported in the other investigations. Consequently, it appears that up to approximately 47% of the respired PCE may be metabolized based on the 1972 Bolanowska and Golacka study, which examined inhaled, retained, and exhaled PCE in humans. A long period of time is necessary for pulmonary elimination of unmetabolized PCE. Stewart and co-workers have analyzed the pulmonary excretion of PCE following experimental human exposures and found that it is biphasic (Stewart et al., 1970; Hake and Stewart, 1977). Initially, elimination is rapid but the second phase is prolonged, with a half-life of approximately 65 hours. Monster and co-workers determined that human pulmonary elimination of PCE has three different phases, with half-lives of 12 to 16 hours, 30 to 40 hours, and 55 to 50 hours, respectively (Monster et al., 1979). Bolanowska and Golacka identified four phases for pulmonary excretion of PCE, with half-lives of 0.025 hours, 0.6 hours, 4.8 hours, and 34 hours, respectively (Bolanowska and Golacka, 1978). Fernandez and associates note that humans exposed to 100 ppm for 8 hours required about 2 weeks to eliminate PCE (Fernandez et al., 1976). 2-36 038928 Sh DRAFT 3. TOXIC EFFECTS IN ANIMALS Estimates of human health risks resulting from exposure to a toxic substance are frequently based on an assessment of animal dose'response data because specific human data are often inadequate for this purpose. In this section, animal PCE toxicity studies are reviewed, including data from bioassays conducted to evaluate the carcinogenicity of PCE. Bioassay results are also used as the basis of the quantitative assessment of carcinogenic potency in Section 5. The toxicity of PCE has also been reviewed by IARC (1979), Reichert (1983), WHO (1984), and the U.S. EPA (1980, 1982, 1984b, 1985a, and 1985b) . The discussion of PCE toxicity begins with analyses of toxic effects to major body organs and systems. For these effects, Appendix A presents a review and summary of dose-response information for different routes and periods of exposure. The information in Appendix A would be relevant to the development of safety limits for PCE exposure in terms of preventing acute, subchronic, and noncarcinogenic chronic toxicological end points in the absence of adequate human-toxicity data. Studies dealing with the teratogenicity of PCE are examined. The section reviews the mutagenic potential of PCE and its metabolites and summarizes the results of animal carcinogenicity bioassays. SL 038929 3-1 TOXIC EFFECTS ON ORGANS AND SYSTEMS DRAFT Hepatic Toxicity Cornish and fellow researchers administered 0.3 to 2.0 mL/kg (0.33 to 4.95 mg/kg) of PCE intraperitoneally (IP) to rats. Liver damage was measured by an increase in serum glutamic oxalacetic transaminase levels (SGOT) and was observed at all doses (Cornish et al., 1973). Ogata and associates observed a decrease in the adenosine triphosphate (ATP) content of liver as well as an increase in the content of lipids and triglycerides after mice were exposed to 800 ppm for 3 hours (Ogata et al., 1968). Elevation of serum glutamic pyruvate transaminase (SGPT) levels in mice was elicited by exposure to 3700 ppm for 9 to 12 hours, as well as by IP administration of 3900 mg/kg (Gehring, 1968). Klaassen and Plaa also measured increased levels of SGPT in mice that received single intraperitoneal doses of l 9 mL/kg (Klaassen and Plaa et al., 1966). These animals had enlarged hepatocytes and slight liver necrosis. A single intraperitoneal dose ot 1,23 mL/kg elevated SGPT levels in dogs. Cornish and Adefuin studied the effects of PCE administered in combination with ethanol (Cornish and Adefuin et al., 1966). A single dose of ethanol was administered by stomach tube (5 g/kg). Rats were then exposed to 4000, 5000, or 10,000 ppm PCE for 6, 4, or 2 hours, respectively. None of these treatments had a statistically significant effect on levels of SGOT, SGPT, or SICD (serum isocitric dehydrogenase). 3-2 DRAFT Carpenter studied the subchronic and chronic inhalation toxicity of PCE in rats (Carpenter et al., 1937). Although no effects were observed in animals treated with 70 ppm PCE (8 hours/day, 5 days/week for 7 months), rats that received 150 exposures of 230 ppm had less glycogen storage than unexposed animals. Exposure to 470 ppm PCE (150 days) caused liver congestion and swelling. Rowe and co-workers exposed guinea pigs to 100 ppm 7 h/d over a period of 17 to 185 days (Rowe et al., 1952). No effects were apparent in animals that received 13 exposures in 17 days. However, when the number of exposures was increased to 132 over 185 days, females had a significant increase in liver weight (p - 0.01) and animals of both sexes exhibited lipid accumulation in the liver. Schumann and associates dosed mice and rats orally with 100, 200, 500, or 1000 mg/kg of PCE daily for 11 days (Schumann et al., 1980). In mice, all dose levels produced hepatocellular swelling, a significant increase in absolute liver weight (p<0.05), and a significant decrease in hepatic DNA content (p<0.05). All of these changes are indicative of hypertrophy (the enlargement of an organ due to an increase in size of its constituent cells). Mice that received 100 mg/kg of PCE displayed an increase in hepatic DNA synthesis. In contrast to the pathologies that developed in mice, only the highest PCE dose, 1000 mg/kg, induced hepatic toxicity in rats. These animals had a statistically significant increase in relative liver weight (p<0.05). A change in the staining affinity of hepatocytes (for hematoxylin and eosin) was also observed. The significance of the latter observation is not known. SL 038931 3-3 DRAFT Buben and 0'Flaherty treated mice by gavage with FCE dosages ranging from 20 to 2000 mg/kg PCE, 5 days/week for 6 weeks (Buben and 0'Flaherty et al., 1985). Liver weights and liver triglycerides were significantly greater than those of controls at doses of 100 mg/kg of PCE and above (p<0.001). A dose-dependent increase in liver degeneration and karyorrhexis was also observed at PCE doses of 100 mg/kg and greater. Activity of glucose-6phosphatase (G6P) was inhibited and a significant increase in SGPT occurred at 500, 1000, 1500, and 2000 mg/kg PCE (p<0.001). Hepatic DNA content was measured in mice treated with 200 or 1000 mg/kg PCE; animals that received 1000 mg/kg had significantly lower levels of DNA (p<0.01). In an NTP-sponsored study of the effects of PCE on mice and rats, animals were exposed to PCE by inhalation 6 hours/day, 5 days/week for 103 weeks (NTP, 1986). Mice were exposed to 100 or 200 ppm PCE and rats to 200 or 400 ppm. Male and female mice of both exposure groups developed liver degeneration and necrosis. Development of these pathologies appeared to be dose related. The incidence of liver degeneration in male mice was as follows: controls, 2/49 animals; low dose, 8/49; and high dose, 14/50. The observed incidence of liver degeneration in female mice in the control group was 1/49 animals; in the low dose group, 2/50; and in the high dose group, 13/50. The number of male mice exhibiting necrosis in the treatment groups increased with Increasing exposure concentrations (i.e., controls, 1/49 animals; low dose, 6/49; and high dose, 15/50). For female mice the incidence of necrosis was: controls, 3/48 animals; low dose, 5/50; and high dose, 9/50. Treated male mice (but not females) had a greater incidence of hepatic nuclear inclusion than, controls (i.e., for controls, 2/49 animals; low dose, 5/49; and high dose, 9/50), The statistical significance of these 3-4 SL 038932 DRAFT data was not evaluated. Under the conditions of this study, rats did not develop hepatic lesions in response to exposure to PCE. Renal Toxicity Klaassen and Plaa administered PCE to dogs intraperitoneally and measured excretion of phenolsulfonephthalein (PSP), a substance used to test for renal function (Klaassen and Plaa et al., 1967). Control dogs excreted 56% of the PSP within 30 minutes; excretion of less than 39% was considered to be an indicator of kidney dysfunction. Kidney function was significantly affected after a single IP dose of 1.4 mL/kg PCE (statistical significance was not given). Plaa and Larson reported that all mice given a single IP dose of 2.5 raL/kg PCE exhibited swelling of the proximal convoluted tubule, and one animal (of six treated) developed necrosis of the proximal convoluted tubule (Plaa and Larson et al., 1965). Mice that received a single dose of 2.5 or 5.0 mL/kg PCE excreted protein in their urine. The statistical significance of these responses was not reported. Carpenter found that rats given 230 ppm PCE (4t>8 hours/day, 5 days/week) for 21 days developed swelling and congestion of the kidneys (Carpenter et al., 1937). This response was exacerbated when the concentration was increased to 470 ppm PCE, Subchronic exposure of mice and guinea pigs to 400 ppm PCE, 7 hours/day for 169 times in 236 days caused swelling of the tubular epithelium along with an increase in kidney weight (Rowe et al., 1952). The NCI cancer bioassay of PCE documented a high incidence of toxic nephropathy in both species of rodents and in all dose groups (toxic nephropathy was defined as degenerative changes in the proximal convoluted SL 038933 3-5 DRAFT tubule, fatty d generation, and necrosis of the tubular epithelium) (NCI, 1977). PCE was administered by gavage. In this study, mice received timeweighted-average (TWA) daily doses of 386 to 1972 mg/kg PCE; toxic nephropathy was observed in 82 to 100% of the animals. Rats received TWA daily doses of 471 to 949 mg/kg PCE; 58 to 94% of these animals developed toxic nephropathy (see Table A-7 for specific data). A bioassay sponsored by the NTP documented kidney casts, nephrosis, and tubular cell karyomegaly in mice (animals received 100 or 200 ppm of PCE 6 hours/day, 5 days/week for 103 weeks) (NTP, 1986). Casts occurred more frequently in treated male mice than in controls (incidence in controls, 3/49 animals; low dose, 9/49; and high dose, 15/50). The trend in female mice was not clearly dose related (incidence in controls, 4/48 animals; low dose, 4/49; and high dose, 15/50). Nephrosis developed at a greater incidence in treated female mice (control, 5/48 animals; low dose, 14/49, high dose, 25/50). For male mice, the incidence of nephrosis in thi.controls was 22/49 animals; low dose, 24/49; and high dose, 28/30. Karyomegaly of tubular cells was treatment-related. The incidence of this pathology in male mice was control, 4/49 animals; low dose, 17/49; high dose, 46/50. In female mice, the incidence of nephrosis was 0/48 animals, 16/49, and 38/50 in the controls, low-dose, and high-dose groups, respectively. The same study reported a dose-related increase of renal tubular cell karyomegaly in rats of both sexes (NTP, 1986). Low-dose animals received 200 ppm of PCE; high-dose animals received 400 ppm (the exposure regime was the same as listed above for mice). The incidence of karyomegaly in male 3-6 SL 038934 DRAFT rats for the corresponding control group was 1/49 animals; low dose, 37/49; and high dose, 47/50. In female rats, the incidence was 0/50 animals, 8/49, and 20/50, respectively. Male rats exhibited a dose-related increase in renal tubular cell hyperplasia (controls, 0/49 animals; low dose, 3/49; and high dose, 5/50). Only one high-dose female rat had renal tubular cell hyperplasia. Pancreas Hamada and Peterson studied the effects of PCE on the electrolyte and protein concentration in bile duct-pancreatic fluid (BDPF) (Hamada and Peterson et al., 1977). The actual source of this fluid (bile duct and/or pancreas) is not known. Rats were given a single IP dose of PCE (1.3 mL/kg in corn oil). Animals were then fasted for 24 hours at which time BDPF was collected and analyzed. The PCE caused a significant increase in BDPF flow, a decrease in concentration of protein in the BDPF, and an increase in the concentration of chloride and potassium (p<0.05 for all parameters). The mechanism of enhanced BDPF is not known. Although Hamada and Peterson discussed possible mechanisms that may be analogous to secretion or cholinergic stimulation, they concluded that PCE (and other chlorinated aliphatic hydrocarbons) altered BDPF by an unknown mechanism, and that the toxicological significance of the reported effects is not known (Hamada and Peterson et al., 1977). SL 038935 3-7 Lungs. Skin^and Evas DRAFT Perchloroethylene is an eye and skin irritant. Application of PCE to the eye of rabbits caused abrasion of the epithelium and conjunctivitis. PCE was also extremely irritating when applied topically to the skin of rabbits (Duprat et al., 1976). However, Jakobsen and co-workers saw no visible sign of skin irritation when guinea pigs were exposed to liquid PCE (Jakobsen et al. , 1982). The NCI reported a high incidence of pneumonia in animals used in the bioassay of PCE (NCI, 1977). Sixty-two to 79% of treated rats and 29 to 66% of treated mice developed pneumonia. However, 95 to 100% of control rats and 28 to 35% of control mice also developed pneumonia. Because of the relatively high incidence of pneumonia observed in control animals, PCE probably did not contribute directly to the infection. In a separate study, chronic inhalation of PCE caused a dose-related incidence of passive congestion of the lungs in mice (NTP, 1986). Reproductive System The only indication that PCE has any effect on the reproductive system comes from the work of Rowe and fellow researchers (Rowe et al., 1952). Seven male guinea pigs were exposed to 1600 ppm, 7 hours/day for 8 exposures within 10 days. Microscopic examination of tissues revealed slight degenerative changes in the germinal epithelium of the testes. The implication of this observation was not discussed and subsequent studies 3-8 SL 038936 have not confirmed the finding. draft Therefore, it is difficult to evaluate the significance of this report. Cardiovascular System Perchloroethylene has been associated with sudden death from cardiac failure (Rowe et al., 1952; Reinhardt et al., 1973). It has been suggested that PCE (as well as a number of other solvents) may sensitize the heart to the effects of endogenously produced epinephrine (Price and Dripps, 1970; Reinhardt et al., 1973). If sensitization occurs, epinephrine-induced stimulation can lead to tachycardia and cardiac failure. This response is apparently precipitated by physical exertion and exposure to high concentrations of this agent. Kobayashi and associates investigated the action of intravenously administered PCE on cardiac rhythm (Kobayashi et al., 1982). Rabbits were anesthetized with urethane, while cats and dogs were anesthetized with pentobarbital. A mean dose of 10 mg/kg of PCE administered with 0.7 A*g/kg of epinephrine produced tachycardia in rabbits (although the most sensitive animals were effected by 5 mg/kg of PCE). Tachycardia also occurred in dogs given a mean dose of 13 mg/kg PCE with 4.2 mg/kg of epinephrine, while doses of 30 to 40 mg/kg PCE decreased left intraventricular pressure. Cats exhibited ventricular arrhythmias after 24 mg/kg of PCE was administered in conjunction with 13 to 14 mg/kg of epinephrine. Rowe and co-workers speculated that death In some rats exposed to concentrations of 3000 ppm PCE or more in air was caused by cardiac failure; SL 38937 3-9 DRAFT however, it is p ssible that cardiac failure occurred as a result of extreme CNS depression (Rowe et al., 1952), Reinhardt and colleagues studied the cardiotoxicity of PCE by exposing unanesthetized dogs to 5000 or 10,000 ppm (R inhardt et al., 1973), Arrhythmias and cardiac failure were not observed and was no evidence of sensitization was seen. The PCE dose levels used in the aforementioned studies are not representative of typical human exposure levels. Furthermore, the work of Kobayashi and associates utilized anesthetized animals, administered PCE intravenously, and used relatively large amounts of epinephrine, none of which readily facilitates extrapolation of results to humans (Kobayashi et al., 1982). The cardiotoxic potential of FCE requires additional experimental work before any conclusions can be drawn. Central Nervous System Acute exposure to PCE typically induces CNS depression. Initial depression can progress to loss of consciousness, anesthesia, and respiratory failure with prolonged or massive exposure. Single oral doses of 1623 mg/kg PCE produced reversible CNS effects in cats (Maplestone and Chopra, 1933), while a single dose of 6492 mg/kg caused lethal CNS depression (Lamson et al., 1929). Death from CNS depression resulted from single PCE doses of 4700 mg/kg (rat) and 6492 mg/kg (dog) (Smyth et al., 1969; Lamson et al., 1929). Over a 4-hour period, 2300 ppm PCE caused an impairment of muscular coordination in female rats, which contributed to a loss of 80 percent of avoidance and escape responses. Animals apparently developed some tolerance 3-10 SL 038938 DRAFT to PCE, because this effect did not persist when dosing was continued over a 2-week period (Goldberg et al., 1964). Rats exposed to 6000 ppm PCE lost consciousness within a few minutes; decreasing the concentration to 3000 ppm required several hours to elicit the same effect (Rowe et al., 1952). The NTP study found that exposure of mice to 2917 ppm for 4 hours was lethal to all animals. Rats appear to be less sensitive to PCE, because a 4-hour exposure to 5163 ppm was required to produce 100 percent mortality (NTP, 1986). Carpenter studied the effects on rats of chronic exposure to 70, 230, 470, or 7000 ppm of PCE (Carpenter et al., 1937). Although various pathological changes were observed at 230 ppm PCE and above, no CNS effects were reported. Savolainen and co-workers observed a slight decrease in brain RNA content and an increase in nonspecific cholinesterase in rats exposed to 200 ppm PCE (6 hours/day for 4 days) (Savolainen et al., 1977). A one-month study conducted by Honma and co-workers documented a dose-dependent decrease in dopamine content of the striatum ip rats exposed to 200, 400, or 800 ppm PCE 12 hours/day (Honma et al., 1980). The decrease was not statistically significant. Norepinephrine content of the hypothalamus and serotonin levels of the cortex and hippocampus increased after exposure to PCE (all three concentrations). None of the increases were statistically significant. A significant decrease in acetylcholine (ACh) levels in the striatum was m asured after exposure to 800 ppm (p<0.05). Drowsiness and other symptoms indicative of CNS depression were reported by Rowe and associates after rats were exposed to 1600 ppm PCE 7 hours/day, 5 days/week over a 25 day period (Rowe et al., 1952). At 2500 ppm PCE, 13 exposures within 18 days caused SL 038939 3-11 Species Concentration or dose TABLE 3-1. TERATOLOGY OF TETRACHLOROETHYLENE TO ANIMALS Exposure duration Exposure period Effect Reference Rat (Sprague Dawley) 100 ppn 7 h/d Days 14 to 20 of gestation No observed fetal or maternal toxicity Nelson et al., 1980 Rat {Sprague Dawley) 300 ppm 7 h/d Days 6 to 15 of gestation Maternal toxicity: reduction in mean body weight (4 to 5%) Fetal toxicity: slight but significant increase in fetal resorptions (9 of 17 litters) Schwetz et al. , 1975 Mouse (SwissWebster) 300 ppo 7 h/d Days 6 to 15 of gestation Maternal toxicity: increase in relative liver weight Fetal toxicity: decrease In body weight, delayed ossification of skull bones, split sternebrae, increase in subcutaneous edema Schwetz et al. , 1975 Rat (SpragueDawley) 900 ppm 7 h/d Days 7 to 13 of gestation Maternal toxicity: decrease in feed consumption and weight Pup toxicity: diminished performance in some behavioral tests Nelson et al. , 1980 SL 038940 Rat (SpragueDawley) 900 ppm 7 h/d Days 14 to 20 of gestation Maternal toxicity: decrease in feed consumption and weight Pup toxicity: diminished performance on ascent test Nelson et al., 1980 DRAFT SL 038941 Species Chicken embryo Concentration or dose Exposure duration TABLE 3-1, (Continued) Exposure period Effect 5 to 10 mmole/egg Injected Examined 14 d into air after incubation space on days 2, 3, and 6 of incubation Malformed embryos at 10 mmole; Estimated LD50 over 100 mmole Reference Elovaara et al., 1979 U1 CO the death of most rats and guinea pigs from CNS depression. Rabbits that received the same treatment displayed signs of CNS depression but did not lose consciousness. TERATOGENICITY The teratogenic activity of PCE has been studied in rats (Nelson et al., 1980; Schwetz et al., 1974; Schwetzet al., 1975) and mice (Schwetz et al., 1975). Maternal exposure levels ranged from 100 to 1800 ppm PCE. Although some minor effects were seen in the progeny, PCE is not considered to be a teratogen. A summary of these studies is provided in Table 3*1. Investigations of the teratogenic and/or developmental effects of PCE have most commonly shown evidence of maternal toxicity, rather than adverse effects on the progeny. Toxicity was evident at 300 ppm (Schwetz et al. 1974; Schwetz et al., 1975) and 900 ppm (Nelson et al., 1980), while maternal death occurred at 1800 ppm (Nelson et al., 1980). Dams exposed to 300 ppm PCE 7 hours/day on days 6 to 15 of gestation had reduced body weights (rats) or an increase in liver weight (mice). The pups of these mice had lower body weights, and there was a slight increase in the number of runts. Some fetuses had subcutaneous edema or delayed ossification of the skull and stemebrae, as well as splits in the sternebrae. These pathologies probably reflect developmental delays and, as such, are considered to be reversible. Developmental delays are believed to result from maternal toxicity rather than from any direct teratogenic activity of PCE (Schwetz et al., 1975). The mechanism of maternal toxicity is unknown, but is thought to involve CNS depression (weight loss) and cytotoxici ty 3-14 DRAFT (change in liver morphology). Because fetal health is often a reflection of the health of the mother, maternal toxicity is a significant concern. The loss of maternal weight, most probably due to decreased feed consumption from subclinical effects (ataxia and anesthesia), can have a great impact on the growth and maturation of the fetus. Maternal malnutrition can cause developmental retardation of the fetus (Doull et al., 1980). The hepatotoxicity observed in mice could also have a profound effect on the growth of the fetus. Maternal toxicity is also suspected of causing a small but significant increase in the number of resorptions in treated rats (p<0.05) (300 ppm PCE, 7 hours/day) (Schwetz et al., 1975). Nelson and associates performed a series of behavioral and biochemical tests on the offspring of exposed rats (Nelson et al., 1980). There were no adverse effects to mothers or their pups following exposure of the mothers to 100 ppm of PCE (7 hours/day) on days 14 to 20 of gestation. Exposure at 900 ppm PCE for 7 hours/day during days 7 to 13 of gestation produced significant differences in neuromuscular coordination (p<0.02) and wire mesh ascent (p<0.05). Exposure at 900 ppm PCE for 7 hours/day during days 14 to hi 20 of gestation caused diminished performance in the wire-mesh ascent test, but increased the performance in the neuromuscular coordination test. A neurochemical analysis of whole brain (minus cerebellum) was performed on newborn and 21-day old pups. Twenty-one-day-old pups from dams exposed during either period had a significant decrease in acetylcholine (p<0.05). A significant decrease in dopamine (p<0.05) was measured in 21-day old pups from dams exposed during days 7 to 13 of gestation. SL 038943 3-15 iJ E Elovaara and co-workers injected 5 to 100 /imol of FCE into the air space of chicken eggs and studied the gross effects on the embryo (Elovaara et al., 1979). Malformations observed were exteriorization of viscera, as well as skeletal and eye abnormalities. These deformities occurred in six embryos of 61 examined. Tests conducted in rodents have not clearly demonstrated that PCE is a teratogen. However, there is some evidence that inhalation exposure of pregnant rodents to FCE can induce developmental delays and altered performance in behavioral tests of the offspring. MUTAGENIC EFFECTS Short-term assays have been conducted to evaluate the ability of PCE to permanently alter genetic information. Most of the tests of genetic activity have been microbial assays that measured forward or reverse mutations. Chromosomal effects have been studied in cultured mammalian cells. * The ability of short-term assays to detect mutagens is compromised by lack of knowledge of the mechanisms involved, by different sensitivity and predictive ability of each test, and by variations in protocols used by separate labs. Despite these problems, short-term assays provide supportive evidence in the evaluation of a compound's carcinogenic potential. 3-16 Si 3a Microbial Assays DRAFT Greim and fellow researchers evaluated the mutagenic activity of PCE (purity >99.9%) in Escherichia coli K12 with and without metabolic activation (S-9) (Greim et al., 1975). The results were negative at four loci tested. Three of these loci are back mutation systems (gal+, arg+, and nad+), while one measures a forward mutation that produces resistance to 5-methyl-DLtryptophan. Only one PCE concentration was used (0.9 mM), and detailed data were not reported. The data of Greim and colleagues was the only study cited by Fishbein in his review of the mutagenicity of halogenated aliphatics (Greim et al., 1976; Fishbein et al., 1976). Cerna and Kypenova reported in an abstract that PCE of unspecified purity produced base-pair substitutions and frameshift mutations in Salmonella tvphimuriujn without S-9 (Ames test) (Cerna and Kypenova, 1977). Concentrations of 0.01, 0.1, and 1.0 mg/mL of PCE produced a dose-dependent increase in the number of revertants. This response was significant only in TA100 (5% level of significance), a strain sensitive to base-pair substitutions. In a host mediated assay that used ICR mice and S. typhimurium strains TA1950, TA1951, and TA1952, PCE reportedly caused a significant increase in the number of revertants (level of significance was not given). The doses used were listed as LD5Q and 1/2LD5Q, but exact quantities were not specified. Because no information was provided on the purity of PCE used, revertant counts, or controls, the significance of these data cannot be evaluated. SL 038945 3-17 DRAFT The results of Bartsch and associates conflict with those of Cerna and Kypenova (Bartsch et al., 1979; Cerna and Kypenova, 1977). Perchloroethylene (99.7% pure) was studied in the Ames test with strain . ryphlmnrlum TAlOO. Concentrations up to 4 x 10 *3 M were not mutagenic in the presence of an S-9 liver fraction from mice pretreated with phenobarbltal. 5 x 10'4 M. Toxicity was observed at concentrations greater than Kringstad and associates evaluated the mutagenic activity of PCE in the Ames test (Kringstad et al., 1981). The PCE (99.0% pure) was tested at a single concentration (0.1 mg/plate) in . tvphimurium. strain TA1535. No source of exogenous metabolic activation (S-9) was used. A slight increase in the number of revertants was observed (31/plate after PCE compared to 19/plate in the ether controls). However, this response was considered negative. The NTP reported the results of a series of Ames tests on PCE conducted by the Environmental Mutagen Test Development Program (NTP, 1986). Salmonella tvphimurium (strains TA98, TA100, TA1535, and TA1537) was incubated with technical-grade PCE in covered test tubes for 20 minutes. The test was conducted both with and without S-9 (S-9 fractions were prepared from the livers of male Sprague-Dawley rats and Syrian hamsters pretreated with Aroclor 1254). The greatest number of revertants was observed in strain TAlOO (all doses, both with and without S-9); high doses (333 /jg/plate) were t xic to TA1535 and TA1537 in the absence of S-9. However, PCE was judged not to be mutagenic in any of the strains, regardless of the concentration tested. 3-18 SL 0389 DRAFT Callen and associates used Saccharomvces cerevisiae. strain D7, to study the frequency of gene conversion (trp5 and ilvl loci) and mitotic recombination (ade2 locus) (Callen et al., 1980). The PCE (purity not given) was added directly to a cell suspension (log phase) and incubated in a closed vial for one hour. Samples were centrifuged, resuspended in buffer, and plated on a medium with glucose. Survival decreased as PCE dose increased from 0 to 4.9, 6.6, or 8.2 mM, At 8.2 mM, survival was greatly reduced and assessment of genetic activity was precluded. Exposure to 6.6 mM PCE elicited substantial increases survivors versus 3.3 in mitotic recombination (52.6 recombinants per 10 4 recombinants per 10 survivors in controls). The number of gene conversions (trp5) also increased at 6.6 mM (8.3 convertants per 10^ survivors versus 1.4 x 10^ convertants per 10^ survivors in controls). The number of revertants at the ilvl locus was not measured at this exposure concentration. Fabre has proposed that mitotic recombination and gene conversion may be induced by the same mechanism (Fabre, 1978). If so, the response may have been inaccurately evaluated, because ade2 recombinants were estimated from plates that had been previously used to determine the number of trp5 convertants. Consideration of this possibility as well as a lack of statistical analysis of results limit the strength of evidence presented by Callen and co-workers (Callen et al., 1980). However, mitotic recombination and gene conversion are indicative of interaction of a substance with DNA, Since the increase in frequency of mitotic recombination was pronounced, this response may warrant additional study. Bronzetti and associates also studied the effect of PCE on the trp5, ade2, and ilvl loci in . cerevisiae (Bronzetti et al., 1983). Cells in the stationary phase were exposed for 2 hours to 5, 10, 20, 60, or 85 mM PCE. Si 38947 3-19 All results were negative. The PCE used was 99.5% pure, whereas the purity of that used by Callen and colleagues was not reported (Callen et al., 1980). When comparing the results of these two studies, consideration must be given to the possibility that cells in the stationary phase may be resistant to the mutagenic or toxic action of xenobiotics. In addition, Bronzetti and fellow researchers conducted an intrasanguineous host-mediated assay that tested the ability of PCE to induce genetic effects in S. cerevisiae (Bronzetti et al., 1983). Yeast were exposed to PCE and its metabolites in the liver, lungs, and kidneys of mice. The PCE did not induce point mutations, mitotic recombination, or mitotic gene conversion. Drosophila The NTP reported the results of an assay that Rested the ability of PCE to induce sex-linked recessive lethal mutations in Drosophila (NTP, 1986' Males were exposed by injection or by feeding and were mated to a series, of untreated females. Neither route of administration produced a statistically significant increase in sex-linked recessive lethal mutations. TESTS OF DNA OR CHROMOSOMAL DAMAGE Cerna and Kypenova did not observe any cytogenetic effects in mouse bone marrow cells following single or repeated IP injections of PCE (daily injections for five days) (Cerna and Kypenova et al., 1977). Cells were recovered for analysis 6, 24, or 48 hours after the last injection of PCE, No details were provided regarding doses used or the specific end points 3-20 0389^8 DRAFT that were studied (the study was published only as an abstract). The NTP published the results of assays for PCE induced chromosomal aberrations and sister chromatid exchanges in Chinese Hamster Ovary (CHO) cells, both with and without S-9 (NTP, 1986). The S-9 fractions were obtained from livers of male rats pretreated with Aroclor 1254. Data were reported in tables only, with no supportive discussion (such as the number of cells scored for each dose level). However, at least three dose levels were used (with and without S-9), as were both positive and negative controls. The PCE had little, if any, cytogenetic effect in either assay. Perchloroethylene was not mutagenic in L5178Y/TK+'/- mouse lymphoma cells, with or without metabolic activation. Cells were treated for 4 hours with 6.25, 12.5, 25.0, 50.0, and 100.0 nL/mL of PCE. Following incubation for 48 hours, cells were plated in medium supplemented with trifluorothymidine for selection of cells mutant at the thymidine kinase (TK) locus. No statistically significant increase in mutation frequency was observed at any dose level (NTP, 1986). Somatic mutations are thought to occur when a substance or one of its metabolites interacts with DNA. Alkylation of DNA is therefore one possible indicator of genotoxicity. Schumann and co-workers measured binding to hepatic macromolecules of mice after administration of radiolabeled PCE (Schumann et al., 1980). No radioactivity was detected bound to hepatic DNA. The specific activity of the tetrachloro[^C]ethylene used in this study was too low to permit detection of low levels of DNA binding. Therefore, these results do not rule out the possibility of DNA alkylation following exposure to PCE. SL 038949 3-21 DRAFT Mutagenic Activity of Metabolites Tetrachloroethylene oxide (PCE oxide) is believed to be the first intermediate formed by microsomal oxidation of PCE. A concentrationdependent mutagenic response was produced by PCE oxide (0.5, 1.3, 2.5, 5.0, and 25.0 mM) in S. tvphlmurlum TA1535 (without metabolic activation) but PCE oxide was not mutagenic to . coli WP2 uvrA. The mutagenicity of this epoxide was also evaluated in a DNA-repair-deficient strain of . coll (. coli pol A 1*), In the latter assay, genotoxicity is measured by comparing differential growth inhibition of the DNA-polymerase-deficient strain, pol A 1 , with a polymerase-proficient strain, pol A 1+. The PCE oxide gave positive results in this test; growth of the pol A 1 strain was inhibited at all dose levels used (Kline et al., 1982). Trichloroacetic acid (TCA) is the principal metabolite of PCE excreted in the urine of rodents and humans; trichloroethanol has also been identified as a metabolite of PCE. In an Ames test conducted with metabolic activation, TCA (0.45 mg/plate) and trichloroethanol (7,5 mg/plate) were not mutagenic to S. tvphimurlum strains TA98 and TA100. There is some indication, however, that trichloroethanol can induce sister-chromatid exchange in cultured human lymphocytes (Gu et al., 1981). CARCINOGENICITY IN ANIMALS Two lifetime bioassays have been completed on PCE (NCI, 1977; NTP, 1986). Additionally, three other studies have addressed the question of PCE carcinogenicity (Rampy et al., 1978; Theiss et al., 1977; Van Duuren et al., 3-22 SL 038950 DRAFT SL 038951 Study NCI, 1977 TABLE 3-2. SUMMARY OF TUMOR-INCIDENCE DATA FROM ANIMAL BIOASSAY STUDIES Species (strain) Concentration Sex or dose Tumor responses Type3 Incidence p-value Mice (B6C3F1) M 0 mg/kg-d HC 2/17 NA 536 mg/kg-d HC 32/49 p<0.00l" 1072 mg/kg-d HC 27/48 p<0.001D F0 386 mg/kg-d 772 mg/kg-d HC 2/20 NA HC 19/48 p<0.00l" HC 19/48 p<0.001 NTP, 1986 Mice (B6C3F1) M M F 0 ppm 100 ppm 200 ppm 0 ppm 100 ppm 200 ppm 0 ppm 100 ppm 200 ppm 0 ppm 100 ppm 200 ppm HC HC HC HAC HAC HAC HC HC HC HAC HAC HAC 7/49 25/49 26/50 16/49 8/49 18/50 1/48 13/50 36/50 3/48 6/50 2/50 NA p<0.00l" p<0.001 NA NS NS NA pCO-OO^ p<0.001 NA NS NS Study NTP, 1986 Species (strain) Rats (F344/N) Sex H F TABLE 3-2. (Continued) Concentration or dose Tumor responses Type Incidence 0 ppm 200 ppm 400 ppm MLK MLK MLK 28/50 37/50 37/50 0 ppm 200 ppm 400 ppm MLK 18/50 MLK 30/50 MLK 29/50 p-value NA p - 0.046 p - 0.004C NA p - 0.023 p - 0.053C DRAFT SL 038952 HC -- hepatocellular carcinoma; HAC - hepatocellular adenoma; and MLK - mononuclear cell leukemia. ^Probability level, Fisher Exact Test (NCI, 1977; NTP, 1986). V1 CProbability level, Life Table Analysis (NTP, 1986). r4o* NA -- not available. NS - not statistically significant. DRAFT resulted in significant increases in malignant neoplasms among the exposed animals. The National Cancer Institute conducted a study in which B6C3F1 mice and Osborne Mendel rats were administered PCE in corn oil by gavage, 5 days/week for 78 weeks (NCI, 1977). Animals were then observed for 32 weeks (rats) or 12 weeks (mice). Mice were 25 days old at initial treatment; rats were 35 days of age. The time-weighted average daily doses of PCE were 536 and 1072 mS/kg for male mice, 386 and 722 mg/kg for female mice, 471 and 941 mg/kg for male rats, and 474 and 949 mg/kg for female rats. Perchloroethylene caused a statistically significant increase in the incidence of hepatocellular carcinoma in mice of both sexes and both dosage groups (p<0.001) (Table 3.2). The time to first tumor development was considerably shorter in treated mice than in controls. Hepatocellular carcinomas were first detected at weeks 91 and 90 in untreated and vehicle controls respectively. However, in male mice, hepatocellular carcinomas were detected after 27 weeks (low-dose) and 40 weeks (high-dose). The first hepatocellular carcinomas were observed in female mice at week 41 (low-dose) and week 50 (high-dose). Exposed mice also exhibited a high incidence of toxic nephropathy, a condition that was not observed in controls. Median survival times of mice were inversely related to dose. Control males had median survival times of more than 90 weeks; this decreased to 78 weeks in low-dose males and 63 weeks in high-dose males. The median survival time of control females was SL 038953 3-25 DRAFT also greater than 90 weeks. Median survival times of low- and high-dose females were 62 and 50 weeks, respectively (NCI, 1977). Early mortality occurred in all groups of rats dosed with PCE. Half of the high-dose males had died by week 44; half of the high-dose females died by week 66. The median survival time of control animals was 88 to 102 weeks, depending on sex. The National Cancer Institute determined that there was a statistically significant association (p<0.001) between increased dosage of PCE and increased mortality (NCI, 1977). The early mortality observed in rats and its statistical association with dose of PCE indicate that the doses given to rats in this bioassay were inappropriately high. Because the optimum dosage was not used and because significant early mortality occurred, these results preclude any conclusions regarding the carcinogenicity of PCE in rats. Questions have been raised about the purity of PCE used in the NCI mouse ami rat bioassays. The PCE was produced by Aldrich Chemical Co. and had u purity of 99%. However, epichlorohydrin (ECH) was apparently used as a stabilizer. It has been suggested that the presence of this contaminant may have directly contributed to tumor induction. The ECH is a direct-acting alkylating agent and is mutagenic (Kucerova et al., 1977; Bridges, 1978). Van Duuren and associates demonstrated that ECH was carcinogenic in mice when injected subcutaneously (Van Duuren et al., 1974). A subsequent study by Laskin and colleagues showed that ECH induced neoplastic lesions of the nasal cavity of rats (Laskin et al., 1980). Most of these tumors were carcinomas of the squamous epithelium. Interestingly, 30-day exposures to 100 ppm ECH produced a much greater incidence of cancer than lifetime 3-26 03895* DRAFT exposures of 30 ppm (exposures were 6 hours/day, 5 days/week). A study by Konishi and co-workers and Kawabata also showed that ECH-fed discontinuously to rats in drinking water at a concentration of 1500 ppm (and at a lifetime TWA dose of approximately 40.2 mg/kg-d) induced a significantly increased incidence of papillomas and squamous cell carcinomas of the forestomach above that of control animals (Konishi et al., 1980; Kawabata, 1981). The exact quantity of ECH present in the PCE used in the NCI study is not known, but it has been estimated that high-dose male mice received 0.42 mg/kg-d (EPA, 1985a). This represents a small fraction of the dose that elicited squamous cell carcinomas in rats. Furthermore, ECH appears to initiate tumors by a localized tumorigenic reaction at sites where it is in direct contact with tissue, such as nasal or forestomach squamous-cell epithelium (EPA, 1984c). No animal in the NCI bioassay developed tumors at these sites. The ECH is among the weakest of the more than 50 suspect carcinogens evaluated by the EPA Carcinogen Assessment Group, having an estimated upper-bound carcinogenic potency, or effect per unit dose at low doses, to humans of 9.9 x 10 *3 (mg/kg-d) * 1 , based on data indicating increased nasal cavity tumor incidence in rats exposed to ECH via drinking water (EPA, 1984c). Using the methodology of the EPA, the equivalent .3 potency to mice would be 9.9 x 10 x (f /f, ), where f and fv are the fractions of body weight consumed as water respectively (EPA, 1984c). The potency for ECH to mice is therefore estimated to be 0.058 (mg/kg-d) \ Using this potency estimate, the highest dosed animals (high-dose male mice) in the NCI bioassays would be expected to incur an increased cancer risk of (0.42) x (0.058) - 0.024, or less than 2.5% (NCI, 1977). Therefore, it is SL 038955 3-27 DRAFT unlikely that ECH contributed significantly to the observed increased tumor incidence in PCE-exposed mice in the NCI bioassay (NCI, 1977). Rampy and fellow researchers exposed male and female Sprague-Dawley rats to PCE by inhalation (300 or 600 ppm) 6 hours/day, 5 days/week for 12 months (Rampy et al., 1978). Animals were subsequently observed for 18 months. High-dose males had slightly greater mortality than did controls, but neither sex exhibited an increased incidence of tumors, regardless of dose. Interpretation of this study is limited by the duration of the exposure and by the fact that it was reported only as an abstract. Theiss and co-workers studied the ability of PCE to induce lung adenomas in A/St male mice (Theiss et al,, 1977). Animals 6 to 8 weeks old were given 80, 200, or 400 mg/kg of PCE in tricaprylin (intraperitoneally) three times a week. Each group received 14, 24, or 48 injections. Animals were sacrificed 24 weeks after the first injection and were examined histologically for the presence of pulmonary tumors. Treated animals did not exhibit a significant increase in the average number of lung tumors when compared to controls. The relevance and validity of these test results are of questionable significance, though, as this test has not produced positive results with several known animal carcinogens. The ability of PCE to initiate and/or promote skin tumors in ICR/Ha Swiss mice was investigated by Van Duuren and associates (Van Duuren et al., 1979). In one group, 163 mg of PCE was applied once to surface skin. Fourteen days after this, phorbol myristate acetate, a promoter, was applied to the same area three times a week for 428 to 576 days. A second group 3-28 038956 DRAFT received 54 mg of PCE by topical application three times a week for 440 to 594 days. The PCE did not show any initiating activity. It also gave negative results in the portion of the experiment that tested its ability to act as a complete carcinogen. It is difficult to interpret these data in relation to the carcinogenic action of PCE because the significance and sensitivity of skin application tests are not thoroughly understood. The most definitive study of the carcinogenic potential of PCE was conducted by Battelle Pacific Northwest Laboratories for the National Toxicology Program (NTP, 1986). In this experiment, B6C3F1 mice and F344/N rats were exposed to 99.9% pure PCE by inhalation, 6 hours/day, 5 days/week for 103 weeks. Mice were exposed to concentrations of 0, 100, or 200 ppm; rats were exposed to concentrations of 0, 200, or 400 ppm. Treated male rats had lower survival rates than control animals (controls, 23/50 animals; 200 ppm, 20/50; 400 ppm, 12/50). Survival rates among female rats showed little variation (controls, 23/50 animals; 200 ppm, 21/50; 400 ppm, 24/50). Both exposure concentrations produced significant increases in mononuclear cell leukemia in female rats (incidence in controls, 18/50 animals; in rats receiving 200 ppm, 30/50; and in rats receiving 400 ppm, 29/50). Life Table analysis showed the significance of these increases to be p - 0.023 (200 ppm PCE) and p -- 0.053 (400 ppm PCE). Treated male rats also developed mononuclear cell leukemia in greater numbers than controls (controls, 28/50 animals; 200 ppm, 37/50; 400 ppm, 37/50). Levels of significance (evaluated by Life Table analysis) are p - 0.046 (200 ppm PCE) and p - 0.004 (400 ppm PCE) (Table 3-2). SI 038957 3-29 Renal DRAFT,tubular-cell adenomas are rare lniee$oppTTaassmmffrrwwiith a historical occurrence at Battelle Laboratories of less than one percent (Appendix F, NTP, 1986). Renal tubular-cell adenocarcinomas are even less common, and have not been documented in any NTP studies (NTP, 1986). Male rats (at the 200 and 400 ppm PCE exposure levels) exhibited an increased incidence of both of these neoplasms (see Table 3*3). Although the increases were not statistically significant, they appeared to be dose-related. Tubular-cell hyperplasia was observed in eight treated males but only in one treated female rat. Tubular-cell karyomegaly developed in a majority of male rats but was less common in females. TABLE 3-3. INCIDENCE OF RENAL TUBULAR CELL ADENOMAS AND ADENOCARCINOMAS IN MALE RATS EXPOSED TO PCE BY INHALATION3 Treatment Neoplasm Control 200 ppm 400 ppm Tubular cell adenoma Tubular cell adenocarcinoma Tubular cell adenoma or adenocarcinoma 1/49 0/49 1/49 3/49 0/49 3/49 (p - 0.259 ) 2/50 2/50 4/50 (p - 0.070 ; aData are from NTP, 1986. V P-values are based on Life Table Tests (Appendix E, NTP, 1986). Brain glioma is a rare tumor of neuroglial cells (the cells that compose the supporting structure of nervous tissue). Brain gliomas were observed in one male control rat and in four male rats that were exposed to 400 ppm PCE (NTP, 1986). This increase was not statistically significant. However, because the historical incidence of these tumors is quite low (0.2% at 3-30 SL 038958 DRAFT Battelle Laboratories), the increased incidence in treated animals in this study is noteworthy. In the NTP study, the survival of low-dose male mice (after week 74) and of high-dose male mice (after week 75) was significantly lower than controls (p<0.001) (NTP, 1986). The survival of high-dose female mice was significantly lower than controls after week 90 (p<0.001). Both concentrations of PCE produced a statistically significant increase of hepatocellular carcinomas in treated mice of both sexes (p<0.001) (Table 32). The incidence of these carcinomas in male mice was as follows: controls, 7/49 animals; low-dose, 25/49; and high-dose, 26/50. The incidence of hepatocellular carcinomas in treated female mice was: controls, 1/48 animals; low-dose, 13/50; and high-dose, 36/50. Hepatocellular adenomas occurred in both sexes of mice and at both concentrations of PCE (Table 3-2), The incidence of adenomas was not statistically significant. However, the combined incidence of hepatocellular adenomas and hepatocellular carcinomas was significant. In males, the combined incidence was: controls, 16/49 animals; low-dose, 31/49; (p - 0.002); and high-dose, 40/50 (p<0.001). In females, the incidence of hepatocellular adenomas and carcinomas was: controls, 4/48 animals; low- dose, 17/50 (p - 0.001); and high-dose, 38/50 (p<0.001). Summary of Evidence of Carcinogenicity in Animals The NCI bioassay of PCE found that administration of PCE by gavage was associated with a statistically significant increased incidence (pcO.001) of SL 3893 g 3-31 DRAFT hepatocellular carcin ma (NCI, 1977). This increase was documented in lowand high-dose PCE*treated B6C3F1 mice of both sexes. A decrease in the time to first tumor development was also observed in treated mice of both sexes and both dose groups. Early mortality in rats prevented an analysis of PCE's carcinogenic potential in this species. The NCI concluded that under the conditions of this study, PCE was a liver carcinogen to B6C3F1 mice of both sexes (NCI, 1977). In 1979, IARC reviewed the NCI study on PCE as well as the animal carcinogenicity studies of Rampy and associates (1978) and Theiss and colleagues (1977), Only two short-term assays were evaluated (Cerna and Kypenova, 1977; Greim et al., 1975). The IARC determined that there was "limited evidence" that PCE is carcinogenic in mice (IARC, 1979) . The IARC recently re-evaluated the evidence of carcinogenocity of PCE to animals (IARC, 1987). They concluded that there was sufficient evidence that PCE is carcinogenic to animals (IARC, 1987). The final report of the NTP inhalation bioassay on PCE was released in 1986 (NTP, 1986). The NTP determined that, under the conditions of this study, there was "clear evidence of carcinogenicity" of PCE for male F344/N rats, "some evidence of carcinogenicity" of PCE for female F344/N rats, and "clear evidence of carcinogenicity" of PCE for both sexes of B6C3F1 mice. In rats, these conclusions were based on an increased incidence of mononuclear cell leukemia in males and females. Male rats also developed renal tubular cell neoplasms (a rare type of tumor). The evaluation of carcinogenicity in mice was based on an increased incidence f hepatoc llular adenoma and 3-32 038960 Si* hepatocellular carcinoma in males, hepatocellular carcinoma in females. and an increased incidence of Si 38961 3-33 DRAFT 4. TOXIC EFFECTS IN HUMANS To assess correctly the health risks from a chemical, consideration of human toxicity data is essential. Unfortunately, information on human toxicity for many substances is limited or is anecdotal in nature. For PCE, however, there have been some controlled inhalation exposure studies to define occupational limits. In addition, epidemiological studies have been conducted to explore the relationship between exposures to PCE vapors and potential health effects. This section presents a brief overview of the health effects of PCE exposures. This discussion is followed by a review of different epidemiological studies dealing with PCE. Finally, human data on the toxic effects of PCE on specific organs and systems are examined. GENERAL TOXICITY Acute exposure to PCE can produce skin irritation and burns, as well as irritation of the eyes and respiratory tract. Central nervous system (CNS) depression is the most immediate effect of exposure, but high concentrations can also cause loss of consciousness and respiratory failure. Liver and kidney toxicity can result from single exposures (Stewart et al., 1961a; Stewart, 1969; Hake and Stewart, 1977), but the concentration and duration are typically greater than those that produce transient CNS effects. Chronic occupational exposure to PCE has caused headache, dizziness, hangover, intoxication, diminished cognitive abilities, and a decreased performance in the Romberg and Flanagan Coordination Tests (Stewart et al., 1961b; Stewart et al., 1970; Stewart et al., 1974). Extended exposure to PCE (2 1/2 months to several years) has also promoted changes in kidney and SL 038962 4-1 DRAFT liver function, cirrhosis, and toxic hepatitis (Coler and Rossmiller, 1953; Meckler and Phelps, 1966; Hake and Stewart, 1977). In experimental studies, human volunteers have been exposed to PCE by inhalation at various concentrations and for various durations. Because subjects were allowed to leave exposure chambers when they felt discomfort, observed adverse effects have been restricted to the respiratory tract and CNS (Carpenter, 1937; Rowe et al., 1952; Stewart et al., 1961b; Stewart et al., 1970; Stewart et al., 1974; Hake and Stewart, 1977). Accidental exposure to PCE has occurred primarily as a result of its use as an industrial solvent. Although specific exposure levels have not always been determined, concentrations have been high enough to cause liver and kidney dysfunction (Coler and Rossmiller, 1953; Hake and Stewart, 1977; Koppel et al., 1985). EPIDEMIOLOGIC EVIDENCE FOR CARCINOGENICITY IN HUMANS Epidemiologic studies of PCE exposure,.have been reviewed by Reichert and by the EPA (Reichert, 1983; EPA, 1985a). Blair and associates analyzed tht death certificates of 330 union laundry and dry-cleaning workers (out of a cohort of 10,000) (Blair et al., 1979). Of 330 decedents, 279 had worked solely in dry-cleaning establishments (while union members). The solvent(s) used by the dry cleaners were not identified. Length of union membership ranged from one to 25 years, with a mean of 13 years. The number of expected deaths from cancer was 67.9 (based on proportionate mortality of the U.S. population) while 87 deaths from cancer were observed. The authors also compared the number of years of union membership with cause of death. 4-2 Si. o^63 DRAFT With the exception of nonwhite males, length of union membership was nearly identical for cancer and noncancer deaths. Increased mortality from cancers of the respiratory tract, cervix, and skin was documented (p<0.05). When all malignancies were evaluated together, the number of observed deaths was also significantly greater than expected (p<0.05). The authors noted that the excess of cervical cancer may be related to the typically low wages and socioeconomic class of this occupational group. Although an excess of liver cancer and leukemia was also observed, these increases were not statistically significant. The increases in cancer deaths among the study group probably contributed to a lower than expected relative frequency of deaths from other causes. It is noteworthy that death from circulatory disease was significantly lower than expected (p<0.005) (Blair et al., 1979). The factors contributing to this phenomenon are not known. Causes of death were not determined separately for laundry and dry-cleaning workers in this study. The actual solvent(s) used were not identified, and smoking history was not documented. The lack of control for smoking proved a significant deficiency of the study because lung cancer was a major contributor to the total number of cancers. Although this study identifies a potential occupational hazard, data provided are not adequate to evaluate the carcinogenic potential of PCE, Katz and Jowett analyzed the mortality patterns of 671 white female laundry and dry-cleaning workers (Katz and Jowett, 1981). Data were obtained from the death certificates of individuals who died in the period 1966 to 1977. Occupational codes listed on the certificates did not distinguish between the two types of work. Data on the duration of employment were not 038964 4-3 b'l* available, nor were the investigators able to determine to which solvent(s) the individuals were exposed. Smoking history was not known. Causespecific proportionate mortality ratios were calculated for 25 causes of death. A significant increase in risk of death from cancer of the kidneys (p<0.05) and genitals (p<0.01) was documented. An excess risk from skin and bladder cancer was also found; however, neither increase was statistically significant. Individuals in the group under study had a greater risk of death from cancer of the cervix, ischemic (obstructive) heart disease, and diabetes mellitus. However, when the effect(s) of low-wage occupations were accounted for, only the risk for diabetes mellitus remained statistically significant (p<0.05). Other studies of laundry and dry-cleaning workers have also reported an increased risk of death from cervical cancer (Blair et al., 1979; Kaplan, 1980); however, these investigators have not compared mortality data by lo'.* wage occupation. Although not definitive, the findings of Katz and Jowet: suggest that factor(s) other than (or in addition to) solvent exposure are important contributors to cervical cancer (Katz and Jowett, 1981). Kaplan completed a retrospective mortality study of 1597 dry-cleaning workers exposed to PCE for at least one year (prior to 1960) (Kaplan, 1980). By the end of the study period, 1028 of the cohort were alive, 285 had died, and the status of the remaining 254 was not known. Although a considerable effort was made to determine the history of solvent exposure, the solvent history of approximately half of the dry-cleaning establishments was unknown. Of those shops with known solvent history, none had used trichloroethylene; individuals who had worked in shops that had used carbon 4-4 03965 DRAFT tetrachloride were eliminated from the study. However, prior to 1960 (the period of interest in this study), the majority of dry cleaners used petroleum solvents (NIOSH, 1980). In keeping with this information, Kaplan decided that employment in a shop with unknown solvent history probably involved exposure to petroleum solvents. Similarities in the physical properties and physiological effects of gasoline, which has been associated with kidney cancer in rats (Kitchen, 1983), and petroleum solvents suggest that use of these solvents may contribute to an increased risk of cancer. The inability of Kaplan to quantify solvent exposure adds an important confounding variable to the study (Kaplan, 1980). The mean exposure concentration of individuals to PCE was calculated to be 22 ppm for drycleaning machine operators and 3.3 ppm for all other jobs. These values are based on a NIOSH survey (NIOSH, 1980) cited by Kaplan (1980). A Standardized Mortality Ratio (SMR)* was used to compare the number of observed deaths to the number of expected deaths by cause. Kaplan found an elevated SMR (182) for malignant neoplasms of the colon (11 observed deaths, 6.77 to 6.98 expected deaths) (Kaplan, 1980). In discussing this observation, Kaplan pointed out that those individuals of high socioeconomic status are at greater risk for cancer of the colon than individuals of low socioeconomic status (Kaplan, 1980). Because dry-cleaning workers generally receive low wages, the study cohort may have overrepresented individuals of low socioeconomic status, and therefore, included a disproportionate number of individuals who are at low risk for colon cancer. If this risk trend is * SMR - gfrserved deaths in study population expected deaths in study population SL 038966 4-5 DRAFT valid, the elevated SMR reported for colon cancer may actually be an underestimate of FOG risk. In addition to colon cancer, SMR's for malignant neoplasms of the rectum (158), pancreas (152), respiratory system (140), urinary organs (198), and "other and unspecified sites (major)" (156) were observed (Kaplan, 1980). Although Kaplan did not evaluate SMR's for statistical significance, a review of this study by the EPA included an evaluation of significance of these malignant neoplasms (EPA, 1985a). The SMR's for cancer of the rectum, pancreas, respiratory system, urinary organs, and "other and unspecified sites (major)" were not significant at the p<0.05 level. However, cancers of the respiratory system, urinary organs, and "other and unspecified sites" were of borderline significance (0.10<p<0.05). For nonneoplastic diseases, elevated SMR's were reported for diseases of the blood and blood-forming organs (290) and for diseases of the stomach am duodenum (211) as well as hernia and intestinal obstruction (125) (Kaplan 1980). These values are based on data from only two to four individuals and cannot be considered definitive. They do suggest the need foi additional study to determine if a relationship does exist between PCE oi solvent exposure and these diseases. Although the relatively small cohort in this study limits conclusions about the carcinogenic potential of PCE, the study (Kaplan, 1980) results suggest a relationship between colon cancer and solvent exposure. Because the NTP study of PCE found hyperplastic and neoplastic changes in the kidneys of treated rats (NTP, 1986), the increase in the SMR from malignant neoplasms 4-6 038967 SL DRAFT of the urinary organs raises the possibility that occupational exposure to solvents may increase the risk of cancer in these organs. Since petroleum products have also been linked to kidney cancer in rats (Kitchen, 1983), and because it is probable that some of the cohort were exposed to petroleum solvents (as well as to PCE), the possible contribution of PCE to an increased risk of urinary organ cancer cannot be ascertained. An additional problem in this study was the inability of investigators to collect data on (and thus control for) smoking history. Since smoking is associated with an elevated risk of many types of cancers (including lung and kidney), its contribution to the elevated SMR's reported by Kaplan needs to be evaluated (U.S. DHEW, 1979). Duh and Asal studied the cause(s) of mortality among 440 laundry and drycleaning workers from Oklahoma who died during 1975 to 1981 (Duh and Asal, 1984). This study had the same problem as the studies of Blair and co workers and Katz and Jowett: smoking histories were not available and separation of the two groups by occupation was not possible (Blair et al., 1979; Katz and Jowett, 1981). Therefore, duration or characterization of individual exposure was not reported. However, Duh and Asal noted that the two groups of workers probably experienced substantially different solvent exposure. The NIOSH reported that, although 75% of dry-cleaning establishments in the U.S. use PCE, Oklahoma may be unique in that petroleum solvents account for more than 50% of total solvents used (NIOSH, 1980). 038968 SL 4-7 DRAFT A Standardized Mortality Odds Ratio (SMOR)* revealed elevated SMOR's for all digestive diseases (1.5), cirrhosis of the liver (1.3), and homicide (3.8). The SMOR's less than 1.0 were reported for diabetes mellitus (0.7), ischemic (obstructive) heart disease (0.8), emphysema (0.8), and suicide (0.2). A SMOR less than 1.0 suggests that laundry and dry-cleaning workers may be at low risk for these diseases. Analysis of deaths due to cancer showed an increase in the SMOR for cancers of the respiratory system (1.8), lung (1.7), and kidney (3.8). Deaths from breast cancer were considerably less than expected (SMOR - 0.1). Brown and Kaplan conducted a retrospective, cohort-mortality study of workers employed in the dry-cleaning industry to evaluate the carcinogenic potential from occupational exposure to FCE (Brown and Kaplan, 1987). The study cohort consisted of 1,690 members of four labor unions (located In Oakland, Detroit, Chicago, and New York City). Individuals selected for the study had been employed for at least one year prior to 1960 in dry-cleaning shops using PCE as the primary solvent. Complete solvent-use histories were not known for about half of the' shops included in the study. Because petroleum solvents were widely used by dry cleaners prior to 1960, most ot the cohort had known or potential exposure to solvents other than PCE (primarily, various types of Stoddard solvents). The investigators also identified a aubcohort of 615 workers who had been employed only in establishments where PCE was the primary solvent. *SM0R was defined by Duh and Asal as a method that compares the number of deaths by specific cause to the number of deaths due to other causes in the exposed population (the odds) to the expected odds derived from a comparison population (Duh and Asal, 1984). 4-8 o^9 t9 SV DRAFT The PCE exposure in shops included in the study was evaluated independently (Ludwig et al., 1983). The geometric mean of time-weighted-average exposures was 22 ppm PCE for machine operators, and approximately 3 ppm for other workers. Brown and Kaplan calculated person-years-at-risk (PYAR) for each worker (Brown and Kaplan, 1987). The PYAR were then combined into five-year calendar periods and five-year age groups by the life-table-analysis-system (Waxweiler et al., 1983). The PYAR values were also evaluated by length of employment and by time lapsed since first employment in a shop that used PCE. The expected number of deaths was calculated by multiplying PYAR (by age and calendar period) by the U.S. mortality rates. Risk of death due to a specific cause was calculated by means of a SMR. Among the (main) cohort, the number of observed deaths from all causes (considered together) was less than expected (493 observed, 575.5 expected; SMR - 86) (Brown and Kaplan, 1987). No deaths occurred from liver cancer, although 3.5 were expected. There were also fewer deaths due to diseases of the circulatory system and nervous system (SMR - 70 and 73, respectively). However, observed deaths from 1 all types of neoplasms were higher than expected (142 observed, 122.9 expected; SMR - 116). Elevated SMR's from malignant neoplasms of the intestine (136) and pancreas (172) were documented. Malignant neoplasms of the urinary tract caused a significant excess of deaths (12 observed, 4.7 expected; SMR -- 255). Of these urinary tract cancers, kidney cancer caused four deaths (2.0 expected; SMR - 200), while eight deaths from bladder cancer were observed (2.7 expected; SMR 296). Mortality from calculi of the urinary system (a nonmalignant disease) 03S97o 4-9 DRAFT was also greater than expected (2.0 observed deaths, 0.3 expected; SMR 667). Although Brown and Kaplan note that there may be an association between calculi and malignant disease of the urinary tract (Brown and Kaplan, 1987), the association is speculative. An elevated SMR for cancer of the cervix (196) and a decreased SMR due to cancer of the breast (87) were attributed to factors associated with the low socioeconomic status of the cohort (Hoover et al., 1975). The subcohort (workers employed only in shops where PCE was the primary solvent) had only one death from urinary tract cancer (1.3 deaths expected). All deaths from urinary calculi (2) occurred in this group. In summary, a statistically significant excess of deaths from urinary tract cancer was observed in those workers potentially exposed to both PCE and petroleum solvents. Individuals employed in shops where PCE was the primary solvent did not have an increased risk of mortality from kidney or bladder cancer. Although these findings do not rule out PCE as the causative agent of urinary tract cancer, the data suggest that other factors or agents may have contributed to the development of neoplastic disease. The possible relationship between exposure to petroleum solvents and kidney cancer has already been noted (Kitchen, 1983). An excess risk of bladder cancer has been associated with cigarette smoking (Matanowski and Elliot, 1981). Brown and Kaplan were not able to document smoking history; however, they calculated the possible effects of smoking on the risk for bladder cancer (based on Axelson, 1978) (Brown and Kaplan, 1987). They concluded that the three-fold increase in bladder cancer among the cohort could not be attributed to smoking. 4-10 SL 038971 DRAFT The currently available epidemiologic scudies add limited information to the understanding of the health hazards associated with occupational exposure to PCE, Although there is some indication that use of dry-cleaning solvents poses a health risk, the contribution of individual solvents to the overall problem is far from clear. Until studies are completed that include a thorough analysis and quantification of PCE exposures, epidemiological studies will not be useful for the assessment of the human health risks of PCE. TOXICITY TO MAJOR ORGANS AND SYSTEMS Information on human toxicity following exposure to PCE has been obtained from case reports of accidental exposures, as well as from a limited number of experimental studies. Human health effects from short- or long-term exposures are similar to those observed in animals. Perchloroethylene initially affects the CNS, and larger doses cause various degrees of hepatic and renal damage. Table 4-1 summarizes human health effects resulting from experimental inhalation exposures to PCE. * Liver The earliest reports of PCE-induced liver damage are associated with its use as an anti-helminth in the 1920's and 1930's. Hundreds of thousands of individuals were treated with PCE for hookworm infestations. The PCE was typically given as a single oral dose of 0.12 mL/kg (maximum of 5 mL) (Reichert, 1983). The PCE reportedly produced hepatic necrosis on some occasions. Damage was transient and recovery took place within 1 to 2 weeks SL 038972 4-11 TABLE 4-1. HUMAN HEALTH EFFECTS FROM EXPERIMENTAL INHALATION EXPOSURE TO PCE. Concentration (ppm) Dose regime Exposure period Effect Reference DRAFT SL 038973 20 75 to 80 100 100 100 to 120 106 150 200 216 210 to 244 5 d/wk for 5 wk Single exposure Single exposure 5 d/wk for 5 wk , Single exposure Single exposure 5 d/wk for 5 wk Single exposure Single exposure Single exposure 7.5 h/d No change in EEGa Stewart et al.. 1974 1 to 4 min Slight eye irritation Stewart et al. , 1961b 7h Headache; sleepiness; Stewart et al., 1970 transient eye, nose, and throat irritation; abnormal modified Romberg test scores 7.5 h/d Increase in delta-wave activity on EEGa Stewart et al., 1974 4 to 6 min 1h Soft palate irritation Transient eye irritation; congestion of frontal sinuses Stewart et al. , 1961b Rowe et al ., 1952 7.5 h/d Lower scores in Flanagan testC Stewart et al., 1974 6 to 30 min Normal^Romberg test scores Stewart et al. , 1961b 45 min to 2 h Eye irritation; sinus congestion; inebriation Rowe e t al.., 1952 30 to 187 min Lightheadedness; difficulty maintaining normal Romberg test Stewart et al. , 1961b DRAFT SL 038974 Concentration (ppm) Dose regime TABLE 4-1. (Continued) Exposure period Effect Reference 280 475 600 911 1060 2000 Single exposure 2h Single exposure 130 min Single exposure 10 min Single exposure 95 rain Single exposure Single exposure 1 to 2 min 7.5 min Lightheadedness; eye irritation; sinus congestion; transient nausea Rowe et al. , 1952 Eye irritation; sinus congestion; slight feeling of elation Carpenter, 1937 Eye and nose irritation; dizziness; diminished motor coordination; some loss of inhibition Rowe e t al.., 1952 Lassitude; inebriation; exhilaration; mental fogginess Carpenter, 1937 Eye and upper respiratory tract irritation Rowe et al,., 1952 Exposure terminated because Carpenter, 1937 all subjects felt faint EEG: electroencephalogram. Romberg test: measurement of ataxia. Flanagan test: measurement of coordination. DRAFT (Hall and Schillinger, 1925; Lambert, 1933). Meckler and Phelps documented a case in which an individual developed hepatitis after a massive inhalation exposure to PCE (of undetermined concentration) (Meckler and Phelps, 1966). The individual's liver remained enlarged 6 months after the exposure. Interestingly, in a separate case of oral overexposure reported by Koppel and co-workers, an individual did not have any measurable liver or kidney damage after ingestion of 8 to 10 mL of relatively pure PCE. Clinical measurements of organ function that were within normal limits included SGOT, SGPT, alkaline phosphatase, red and white blood-cell counts, and serum creatinine. However, the individual was hospitalized for treatment within 1 hour of ingestion, which probably averted organ damage (Koppel et al., 1985). Stewart and co-workers have investigated several cases of acute overexposure to PCE that caused liver damage. The earliest report documented the clinical effects of occupational exposure to the vapor of a petroleum-based solvent mixture that contained approximately 50% PCE (Stewart et al., 1961). The individual was exposed to this mixture for about 3.5 hours, which caused a loss of consciousness. Simulation of exposure conditions gave an estimated concentration of 250 ppm, with levels that reached 1000 ppm for the last 30 minutes. Nine days after the incident, urinary urobilinogen and serum bilirubin levels were elevated, an indication of liver impairment. On the 18th day after exposure, a slight elevation of SGPT was measured (Stewart et al., 1961a). In a separate incident reported by Stewart, a worker was overcome by PCE vapors (of unknown concentration) (Stewart, 1969). The exposure lasted about 10 minutes and produced hepatic dysfunction. Clinical measurements of liver function were normal shortly 4-14 SL 038975 DRAFT after exposure. A slight increase in SGOT levels was measured on the third and fourth days following exposure, and urinary urobilinogen levels were elevated on the ninth day. Elevated serum enzymes were also measured by Hake and Stewart after a massive overexposure to PCE (estimates of the exposure concentration were not made) (Hake and Stewart, 1977). The individual was found lying in a pool of PCE, some of which was probably absorbed dermally as well as through the lung. Recovery was complete within 21 days. Eight of nine firemen exposed to PCE vapors for 3 minutes (unknown concentration) had elevated SGOT levels. Hepatomegaly and splenomegaly were also found in one individual. Normal function was regained within 22 to 63 days (Saland, 1967). The effects of chronic occupational exposure to PCE vapor were studied in seven individuals exposed for 2 to 6 years (Coler and Rossmiller, 1953). Short-term tests indicated that exposure levels ranged from 232 to 385 ppm. The authors assumed that these measurements were representative of levels that workers were exposed to 8 hours/day, 5 days/week. Of the 7 individuals, 3 had abnormal liver-function tests, and 1 was diagnosed as having cirrhosis. Kidneys In animals, kidney damage generally occurs at exposure levels greater than those that cause liver toxicity. This trend seems to hold true for humans as well. Several cases of overexposure great enough to cause loss of SL 038976 4-15 DRAFT consciousness have not produced measurable kidney damage (Stewart et al., 1961a; Stewart, 1969; Patel et al., 1973; Patel et al., 1977). In 1 case of overexposure (where an individual was unconscious in a pool of PCE for 12 hours), kidney damage was measured by proteinuria and hematuria. These effects lasted for 20 days and for 8 days, respectively (Hake and Stewart, 1977). Lungs Pulmonary edema has been documented in only one instance. Exposure to PCE was probably massive (>1500 ppm) because the individual was comatose and required mechanical ventilation upon admission to the hospital. Recovery was complete within 4 days (Patel et al., 1977). Skin and Eves Dermal contact with PCE causes localized irritation; prolonged exposure can cause erythema, first- and second-degree burns, and blistering (Gold, 1969; Stewart et al., 1961b; Hake and Stewart, 1977). The PCE vapor is also irritating to the eyes (Carpenter, 1937; Rowe et al., 1952). Exposure of human volunteers to 106 ppm produced transient eye irritation; this became more pronounced when the concentration was increased to 216 ppm (Rowe et al., 1952). Stewart and colleagues has also reported eye irritation when humans were exposed to 100 ppm for a period of 7 hours (Stewart et al., 1970). SL 038977 4-16 Connective Tissue DRAFT Sparrow described a patient who had a connective tissue disorder with similarities to a syndrome observed in vinyl chloride workers (Sparrow, 1977). The individual in question was exposed to PCE vapor during his work at a dry cleaners. No measurements of concentration were reported, but at least once a week (over a 4-year period), exposure was high enough to cause dizziness and sleepiness. The individual displayed pathological changes in the skin of the hands, acrocyanosis, and polymyopathy. Abnormalities in the immune system and in hepatic function (mild hepatitis) were also documented. The patient may have been abnormally sensitive to PCE, perhaps related to an existing abnormality in the immune system. Indications that the individual did have an abnormal immune system are suggested by intermittent alopecia areata (since childhood), vitiligo (an apparently autoimmune condition characterized by destruction of. melanocytes), and an absence of immunoglobulin A (IgA). This appears to be a unique case report, whose information is confounded by the abnormal medical history of the individual. Although PCE exposure may have contributed to the disease, it is not clear if PCE was the sole causative agent. Central Nervous System Acute exposure to PCE generally causes temporary CNS effects such as dizziness, headache, and confusion. However, massive single exposures can cause loss of consciousness (Patel et al., 1973; Patel et al., 1977; Hake and Stewart, 1977), and have been fatal in at least two instances (Lukaszewski, 1979; Levine et al,, 1981). Protracted exposure produces SL 038978 4-17 DRAFT symptoms similar to those observed after short-term exposure, although the effects apparently persist, even after exposure is terminated. Rowe and associates did not observe any CNS effects in humans exposed to 106 ppm (single exposure of unspecified duration) (Rowe et al., 1952). Minor CNS effects were produced by 216 ppm (45 minutes to 2 hours), and a 10minutes exposure to 600 ppm significantly affected motor coordination. Eleven volunteers exposed for a single 7-hour period to 100 ppm experienced headache, dizziness, and somnolence (Stewart, 1970). Three of the 11 had abnormal scores on the Romberg test, which measures ataxia. Tests of coordination, visual inspection, visual acuity, and depth perception were normal. Carpenter noted only minor and transient CNS effects in individuals exposed to 500 ppm for 2 hours (Carpenter, 1937). When the same subjects were exposed to 911 ppm, they complained of lassitude, exhilaration, and inebriation. In a study conducted by Stewart and co-workers, individuals were exposed to 20 to 150 ppm of PCE, 7.5 hours/day over a 5-week period (Stewart et al., 1974). No alterations in the EEG were notedat 20 ppm, but some aberrant EEG tracings were seen after 100 ppm. A decrease in Flanagan Coordination Test scores was observed following exposureto 150 ppm. In a subsequent experiment, individuals were repeatedly exposedto 0, 25, or 100 ppm of PCE, 5.5 hours/day over 11 week (Stewart et al., 1977). CNS effects were observed by measuring the subjects' response on the Romberg, Michigan Hand- Eye Coordination, and Flanagan Coordination Tests, and on the EEG. In contrast to the 1970 study, exposure to 100 ppm did not produce abnormal SL 038979 4-18 DRAFT scores on the Romberg Test. PCE at 100 ppm did cause a significant decrease in Flanagan Coordination Test scores. Coler and Rossmiller documented subjective complaints of malaise, dizziness, headache, lightheadedness, and intoxication in individuals regularly exposed to concentrations of 232 to 385 ppm PCE (Coler and Rossmiller, 1953). Gold described a case history of an individual exposed to PCE vapors, 6 to 7 days a week for 3 years (Gold, 1969). The individual was hospitalized after exhibiting confusion, disorientation, agitation, and an inability to concentrate. A neurological examination revealed a normal EEG. However, his performance on psychological tests that required concentration was "poor", and he showed "marked confusion." These problems persisted over a 12-month follow-up period, although there was no further exposure to PCE (the individual was lost to follow up at this time). Gold concluded that there was suggestive evidence of both cerebral and cortical damage, and basal ganglia involvement (Gold, 1969). However, it was not possible to obtain conclusive evidence of any neurological damage. Gregersen and colleagues examined 65 Danish workers exposed to various organic solvents for neurotoxic effects (Gregersen et al., 1984). Although only 15% of the cohort were exposed to PCE (approximately 100 ppm), the results are intriguing. Individuals were examined for intelligence, given a neuropsychological exam, and were subjected to a number of neuropsychological tests. The authors concluded that solvent exposure was correlated with acute neurotoxic symptoms, as well as longer-lasting symptoms of intellectual impairment. A relationship between exposure and signs of peripheral neuropathy was also observed. SL 038980 4-19 McMullen reported on a case in which an individual was exposed to PCE vapor (500 ppm) for an undetermined period of time (McMullen, 1976), CNS depression was observed; the effects were described as resembling alcohol intoxication. No clinical measurements of sensory or organ function were made, but the individual apparently recovered within 6 hours. Reproductive System No studies have addressed the question of whether exposure to PCE affects the human reproductive system. At present, there is no evidence of human reproductive toxicity from PCE exposure. Cardiovascular System Some organic solvents have been ' associated with cardiac arrest due to ventricular fibrillation. It is thought that these compounds sensitize the heart to epinephrine-induced arrhythmias. There is suggestive evidence that PCE has this effect in animals (see Section 3), and there is one report oi PCE-induced cardiac arrhythmias in humans. Abedin and associates observed that occupational exposure to PCE probably Caused dizziness and premature ventricular contractions in one case (Abedin et al., 1980). Although PCE may not have been the only factor in this response, removal from exposure to PCE alleviated the symptoms. TERATOGENIC EFFECTS There is no published information on the teratogenicity of PCE in humans. 4-20 SL 038981 MUTAGENIC EFFECTS DRAFT Ikeda and co-workers examined lymphocytes from individuals exposed to PCE for 3 months to 18 years (Ikeda et al., 1980). Chromosomal aberrations, sister-chromatid exchanges, and alterations of the mitotic index were the cytogenetic effects studied. The exposure level in 1 group of workers was 92 ppm (geometric mean), while a second group was exposed to 10 to 40 ppm (the authors did not give TWA exposure concentrations). Although a control group was included, the criteria used to select and/or match controls was not reported. Exposed individuals did not have a significantly greater frequency of chromosomal aberrations or sister-chromatid exchanges, nor were there any substantial differences in the mitotic Index. Trichloroethanol is a metabolite of PCE isolated from the urine of humans. Gu and associates reported a slight increase in the number of sisterchromatid exchanges per cell in human lymphocytes exposed to 178 mg/L of trichloroethanol (Gu et al., 1981). Neither of these studies is adequate to assess the mutagenic potential of PCE ^nd its metabolites in humans. SUMMARY OF EVIDENCE OF HUMAN CARCINOGENICITY Evaluation of the carcinogenic potential of a chemical is based on the results of short-term assays of mutagenesis, pharmacological data (e.g., distribution and metabolism), lifetime animal bioassays, and epidemiological evidence. Several agencies and groups have developed systems of classification for evaluating evidence on the carcinogenic activity of a substance. The International Agency for Research on Cancer (IARC) separates SL 038982 * 4-21 DRAFT strength of evidence of carcinogenic activity into 4 groups: sufficient evidence, limited evidence, inadequate evidence, and no evidence of carcinogenicity (IARC, 1982). Inclusion in any of these categories is based on data from short-term assays, as well as animal and human studies (if available). The EFA uses the same groupings, but places a substance in one or the other category solely on the basis of animal bioassay data (EPA, 1984a). To assess overall evidence of carcinogenic potential to humans, 6 additional categories are used: group A - Human Carcinogen, group B - Probable Human Carcinogen (further separated into B1 and B2), group C - Possible Human Carcinogen, group D - Not Classified (due to inadequate animal evidence), and group E - No Evidence of Carcinogenicity for Humans (EPA, 1984a). In the absence of sound epidemiological data, the greatest weight ei evidence in a carcinogen assessment is typically given to the result.- t lifetime animal bioassays. The criteria employed in analysis of bioasja\ data include an increase in the incidence of tumors in treated animals over those noted in controls, a decrease in latency (time to tumor development) development of rare tumors, and an increase in the number of tumors in individual animals. In 1982 and 1987, the IARC evaluated available information on PCE and determined that there was "inadequate" evidence to conclude that PCE Is carcinogenic to humans (IARC, 1982). An analysis of animal data gave sufficient evidence of carcinogenicity (IARC, 1987); evidence for any activity of PCE in short-term tests was "inadequate" to judge it.'. 4-22 SL 038983 DRAFT carcinogenic potential (IARC, 1982). The EPA Health Assessment Document for Tetrachloroethylene (Perchloroethylene) has also analyzed the evidence of carcinogenicity of PCE (EPA, 1985a). This evaluation included an extensive review of short-term test results, data from animal tests, and several epidemiological studies. The EPA concluded that the evidence for the carcinogenicity of PCE in animals is "limited," and that the epidemiological data were inconclusive. Perchloroethylene was placed in group C, a possible human carcinogen (EPA, 1985a). It should be noted however, that since these 2 analyses were published, other epidemiologic and animal study results have become available. It Is possible that a subsequent evaluation by either the EPA or IARC would result in a different conclusion. SL 038984 4-23 5. QUANTIFICATION OF PCE'S CARCINOGENIC POTENCY The term carcinogenic "potency" is used herein to mean the quantitative increase in tumorigenic risk per unit dose for very low exposure levels. Risks are predicted using the "linearized" multistage dose-response extrapolation model that has been generally used for cancer risk assessment by the California Department of Health Services (CDHS, 1985) and the U.S. Environmental Protection Agency (EPA) (EPA, 1985b; Anderson et al., 1983). The following carcinogenic potency assessment proceeds In 5 steps: (1) selection of those cancer bioassay data sets suitable for dose-response assessment; (2) determination of the empirical relationship between applied dose and cancer risk for each data set selected; (3) determination of the relationship between administered dose and the amount of PCE metabolized for animals and humans; (4) expression of carcinogenic potency in terms of rate of metabolite formation; and (5) expression of carcinogenic potency in terms of estimated human applied dose, taking PCE metabolism into account. SELECTION OF BIOASSAY DATA INDICATIVE OF PCE CARCINOGENICITY The bioassays used in this potency assessment are the National Cancer Institute (NCI) study of mice exposed to PCE by gavage (NCI, 1977) and the recent National Toxicology Program (NTP) studies of mice and rats exposed to PCE by inhalation (NTP, 1986). The EPA used the gavage study as the basis of a carcinogenic potency assessment in 1985 (EPA, 1985a) and used the inhalation study to update this assessment in 1986 (EPA, 1986). In each of these studies PCE caused a significantly increased tumor incidence (see Section 3). These 5-1 bioassay data sets are used here because they represent the only long-term exposure studies with well-defined, exposure-response data that indicate a positive carcinogenic response for FCE in animals. Human epidemiological data suitable for dose-response assessment are not available. Brief reviews of the administered dose and associated tumor response for the NCI gavage study with mice and for the NTP inhalation studies with rats and mice are given below. This information is summarized in Table 5-1, along with other information derived later, which also is tabulated here for convenience. From the doses (D) administered in those bioassays corresponding lifetime, time-weighted average (TWA) applied doses (A), are derived using certa.i standard assumptions (noted below). As discussed in Section 2 PCE is extensively metabolized in mammals. Themetabolites (rather than PCE) have been suggested to be responsible for almost i all acute, chronic, and carcinogenic effects (e.g., Pegg et al., 1970 Schumann et al., 1980). Postulated toxic metabolites of PCE that have bta . identified include trichloroacetyl chloride, trichloroacetic acid, a: ,i trichloroethanol; the formation of the reactive epoxide tetrachloroethylen? oxide is thought to be the first step in the metabolism of PCE (Bouse et al 1975; Grelm et al., 1975). Other minor metabolites have also been identified and/or proposed. If a reactive metabolite is indeed responsible for the carcinogenic effects of PCE, then the total metabolized dose may be useful as a surrogate measure of the actual delivered dose. Since the metabolized dose is likely to be smaller than the administered dose, using the metabolized dose to calculate cancer potency would result in a higher potency value than would the administered dose. Using available data, carcinogenic potencies have been SL 038986 5-2 DRAFT SL 038987 Study species (strain) NCI, 1977 Mice (B6C3F1) NTP, 1986 Mice (B6C3F1) TABLE 5-1. DOSE-RESPONSE DATA FOR SELECTED CANCER BIOASSAYS Sex and weight (kg) Kale 0.030 Administered dose or cone., D 0 536b 1072 TWA Applied dose, A. (me/kc-d) 0 215.0d 430.ld Type Tumor hepatocellular carcinoma Female 0.025 0 386b 772Kb 0 154.9d 309.7 hepatocellular carcinoma Male 0.037 0 100C 200C 0 146.6 293.2 hepatocellular carcinoma Incidence 2/20 32/48 27/45 0/20 19/48 19/45 7/49 25/47 26/50 Female 0.032 Male 0.037 0 iooc 200C 0 100c 200c 0 153.9 307.8 0 146.6 293.2 hepatocellular carcinoma hepatocellular adenoma/carcinoma 1/44 13/42 36/47 16/49 31/47 40/50 TABLE 5-1. DOSE-RESPONSE DATA FOR SELECTED CANCER BIOASSAYS Study species (strain) Sex and weight (kg) Administered dose or cone., D 0 TWA Applied dose, A. (rae/kg-d) 0 Type Tumor Incidence4* 4/44 Female 0.032 100c 200 153.9 307.8 hepatocellular adenoma/carcinoma 17/42 38/47 NTP, 1986 Rats (F344/N) Hale 0.44 0 200C 400c 0 143.0f 286.0 mononuclearcell leukemia 28/50 37/48 37/50 Female 0.32 0 200c 400C 0 159.0f 318.1 mononuclearcell leukemia 18/49 30/50 29/50 3Tumor -incidence denominator excludes animals dying before the occurence of the first corresponding tumor type observed in each study. ^Average administered daily gavage dose, D, in mg/kg-d, for a 5 d/wk exposure over 78 wk of a 90-wk bioassay. DRAFT 038988 CAverage administered inhalation exposure in ppm 6 h/d, 5 d/wk over 2-years. ^Time-weighted average (TWA) dose: A - D (5/7) x (78/90) (90/104)^ (following Anderson et al. 1983). c/i eTotal respired dose averaged over time: A = D (6/24) (5/7) (0.0345)(body weight/0.025) 2/3 ^ (6.78 mg PCE/m ppm). fTotal respired d^se averaged over time: A = D 6/2'0 (5/7) (0.105) (body weight/0.113) 2/3 (6.78 mg PCE/m -ppra)/body weight 'fo1lowing / :>dert.on et al.. 1983). DRAFT estimated on the basis of metabolized dose (denoted M) , in addition to those based on the administered dose or "applied" dose (denoted A) (EPA, 1985a, 1985b, and 1985c; EPA Draft, 1986; CDHS, 1985). From pharmacokinetic data an empirical relationship was obtained between administered dose and the total metabolized dose. PHARMACOKINETIC ANALYSES In two previous risk assessments, EPA calculated the metabolized dose (M) as a function of the administered dose (D) for the NCI and NTP bioassay data using a simple steady-state pharmacokinetic (SSPK) approach (EPA, 1985a; EPA Draft, 1986). In using this approach the EPA did not take into account; (1) bioassay duration relative to expected lifespan; (2) male/female animal weight differences; (3) the treatment of species-specific data on metabolic elimination pathways; (4) the scaling of certain kinetic parameters across species; and (5) the treatment of available human metabolic data. Therefore, a reanalysis of the relevant pharmacokinetic data was undertaken as part of the current potency assessment for PCE. This reanalysis, however, relied on the same basic SSPK approach that EPA used. The SSPK approach presented here and used by the EPA Is fairly crude, primarily because much of the available metabolic data were not gathered under steady-state physiological conditions at the very low levels of anticipated human exposure. Physiologically-based pharmacokinetics (PE PK) is a relatively new form of pharmacokinetic modeling designed to consider physiological processes, biochemical kinetics, and tissue levels of the chemical (or its metabolites) under investigation, based on available biological data. This approach SL 038989 5-5 DRAFT represents the body by compartments, into and out of which compounds are transferred by diffusion and blood flow. This aids in cross-species comparison of metabolizing enzymes, metabolism pathways, organ volume and blood flow, absorption and distribution phenomena, differences in body size and composition, transport characteristics, and differences in the relative routes of detoxification and excretion. Preliminary results from FB-PK models of PCE metabolism for bioassay and environmental contexts are available (EPA Draft, 1986; Reitz and Nolan, 1986). Although a detailed description of a PBPK approach is beyond the scope of the present analysis, results using such an approach along with results obtained using the simpler approaches are presented here. The use of metabolized, rather than applied, dose adds and/or emphasizes several uncertainties relevant to the analysis of the experimental uoso response data, including: uncertainty in identifying the active carcin<: = species; uncertainty in parameter estimates based on experimental animal ''a: uncertainty involved in extrapolating kinetic constants between species; uncertainty in some key constants estimated for humans based on limited humi' metabolism data; uncertainty in the application of the non-steady state PP model (which purportedly adjusts variables to steady-state conditions uncertainty in the detection of all human PCE metabolites; uncertainty in the relative metabolic rates at environmental exposure levels; and uncertainty regarding the extent to which inter-individual variability relates to pharmacokinetic parameters, particularly for humans. The problem of inter individual variability is, of course, common to many problems in predictive regulatory toxicology, not just pharmacokinetic analysis. For risk assessments using toxic endpoints other than carcin genesis, this is generally 5-6 SL 038990 UKAFT dealt with by incorporating a safety factor. With regard to the other sources of uncertainty mentioned, an attempt is made here and in other reports (EPA, 1985a, 1985b, and 1985c; EPA Draft, 1986) to assess available metabolism data for internal consistency. The alternative to using estimates of metabolized PCE dose in this analysis is to rely on the applied dose. This latter approach would not consider possible interspecies differences in: 1) absorption of the administered dose; 2) both the qualitative and quantitative metabolism of the PCE; and 3) relative rates of detoxification. However, the applied dose approach can be adjusted for absorption differences. The decision to use the applied dose or the metabolized dose model involves a trade-off between the additional parameter uncertainties associated with a physiologically-based pharmacokinetic model and a potential systematic error associated with an applied dose model. The optimum approach hinges on scientific/policy judgments concerning the quality of data underlying (1) identification of carcinogenic mechanism of PCE and (2) pharmacokinetic parameter estimates. The problems and benefits associated with both approaches are discussed below for the different bioassays. METHODOLOGY Evaluation Of Cancer Potency To calculate potential risks associated with PCE exposure, the staff of CDHS used a linearized multistage model, which provides a reasonably healthprotective risk estimate due to its property of being linear at low exposure levels. The lifetime probability of developing a tumor (p) induced bv n St, 3899i 5-7 DRAFT average daily intake of chemical (d) is often assumed to be (CDHS, 1985; Anderson et al., 1983): 21 p(d) - 1-exp-(qQ + q^d + q2d +... + 9jd > with constraints q^> 0 for all i. The q^ are parameters of the model which are taken to be constants and are estimated by maximum likelihood techniques from the data. The parameter represents the background lifetime incidence of the tumor. The q^ value or some upper bound is often called the cancer potency, since for small doses it is the ratio of excess lifetime cancer risk to the average daily dote received. For the present discussion, cancer potency will be defined as q' the upper 95% confidence bound of q^, as estimated by maximum likelihood techniques. When the dose is given in mg/kg-d, the parameters q^ and q| are given in units (mg/kg-d) Details of the estimation procedure are given In Crump, Guess and Deal (1977). To estimate potency in animals from experiments of duration Te, rather than the natural lifespan of the aninuu. (T), it is assumed that cancer incidence increases with the third power age: ^animal ' '! * <T/I*>3 Following Gold et al. (1984) and the EPA (Anderson et al., 1983), the natural lifespan of mice and rats is assumed to be two years, so that for experiments lasting Te weeks in these rodents: ^animal * q! * <104A=)3. 5-8 SL 038992 DRAFT To estimate risk at low doses or exposure levels, the potency value is multiplied by the average daily dose. Interspecies_ .Scaling Once a potency value is estimated in animals using one of the techniques described above, human potency can be estimated. As described in the California risk assessment guidelines (CDHS, 1985), a dose in units of milligram per unit surface area is assumed to produce the same degree of effect in different species in the absence of information indicating otherwise. Under this assumption, scaling to the estimated human potency ^qhuman^ can be ac^ievet^ by multiplying the animal potency value (9anima^) by the ratio of human (bw^) to animal body weights (bw^) raised to the one-third power when animal potency is expressed in units (mg/kg-d) ^human ^animal X (bw^/bw^) Body weights are typically assumed to be 70 kg for humans, 0.35 kg for rats and 0.030 kg for mice, but each experiment should be interpreted with respect to the actual weight recorded during the bioassay. Dose Adjustments Several adjustments need to be made to the experimental exposures to calculate the lifetime daily exposure levels. Thus, for inhalation exposures, the reported dose must be multiplied by: SL 038993 5-9 DRAFT H/24: where H is the hours of exposure per day. This converts the exposure period to a time weighted average for 24 hours daily continuous exposure, D/7: where D is the number of days exposed per week. This converts the dosing schedule to a time weighted average for a seven day/week continuous exposure. Le/L: where Le is the length of the experiment and L is the lifespan of the animal (the longer of Le or 24 months). This converts the experimental protocol to a continuous lifetime exposure. NCI (1977) Mouse Study: Gavage Applied Dose and Tumor Incidence In the NCI study, PCE in corn oil was administered to male and female B-'v, ' i mice, beginning at about 5 weeks of age, by gastric intubation for 5 days/*, for 78-weeks, followed by 12 additional weeks of pre-sacrifice observation : : 1977). Incidence of hepatocellular carcinomas increased with dose for b~ n males and females (see earlier discussion in Section 3). Table 5-1 give- ,;ie average administered daily dose, D, for the male and female vehicle contr 1 low- and high-dose groups along with corresponding tumor incidence data ( EPA, 1985a). Note that D is the time-weighted average (TWA) dose (in units of mg/kg). In table 5-1, in contrast to Table 3-2 tumor-incidence data are given as the number of tumors found in animals which survived at least until the appearance of the first carcinoma in each study (which appeared at week 24 for 5-10 DRAFT female mice and at week 41 for male mice in the NCI study). The mortalityadjusted incidence data are used here in preference to unadjusted incidence data to partially correct for the influence of competing mortality risks. Metabolized Dose The pharmacokinetic data of Buben and 0'Flaherty (1985) for PCE metabolism in mice were used to calculate the metabolized dose from the applied dose. In this study, PCE metabolism was examined in 3- to 5-month-old, male Swiss-Cox mice during subchronic administration of PCE by gavage in corn oil for 5 days/week for 6 weeks, exposure conditions similar to those of the NCI study. Seven different animal groups received dose levels ranging from 20 to 2000 mg/kg. Metabolism was estimated by measuring the daily excretion of the sole urinary metabolite detected, trichloroacetic acid (TCA). Dose-dependent metabolism was observed to be described by the Michaelis-Menten equation: ,, DV M,, " _____ max.u D + Km (5-1) where Mu - TWA yield of urinary metabolite, mg/kg-d; D - TWA daily administered dose, mg/kg-d; Vmax u "aPParent maximum rate of urinary metabolite production, mg/kg-d; and - apparent Michaelis constant, mg/kg-d. The value of M^ is essentially equivalent to the dose rate in mg PCE metabolized, since the molecular weights of PCE and TCA differ by less than SL 038995 5-11 DRAFT 1.5%. In the mg/kg-d, while Buben and 0'FlahertyJ studJy, Vmax,u was estimated to be 136 was estimated to be 660 mg/kg-d (Buben and O'Flaherty, 1985). If for different species the dose estimated from the urinary metabolites, M , were a constant proportion of total metabolized dose (which we shall denote as M), then it would make no difference whether M or were used to derive a relationship between cancer risk and applied dose, D. But this proportion does not appear to be constant for different species. In B6C3F1 mice, for example, the mass-balance metabolic study by Schumann and co workers reported that the total amount of PCE metabolites found after either a 6-hour inhalation at 10 ppm or a single, oral dose of 500 mg/kg of radiolabeled PCE, were 80.0% and 82.3% of the administered dose, respectively, in the form of urinary metabolites (Schumann et al., 1980). These results were based on the assumption that metabolism was complete by 72 hours post-exposure and that cagewash radioactivity was attributable solely to urine content. For Sprague-Dawley rats, a metabolic study by Pegg and associates found that urinary metabolites comprised only 58.9% and 57.1% of the total PCE metabolized after exposures of 10 ppm for 6 hours and 1 mg/kg by injection, respectively, of radiolabeled PCE (and only 49.9% and 45.7% after the higher exposures of 600 ppm for 6 hours and 500 mg/kg by injection, respectively) (Pegg et al., 1979). Since both studies assumed complete metabolism by 72 hours post-exposure, the estimates of M would be decreased to a greater extent for rats than for mice. Thus, these two studies indicate that Mu was a greater fraction of M for the mice than for the rats. In addition, M was found to decrease with an increase in dose for u rats. 038996 5-12 DRAFT Based on the proportion M^/M indicated by Schumann et al. (1980), and the relationship between applied dose and urinary metabolite production indicated b*y Buben and O'Flaherty (1985), Mu and Vmax,u in mice are assumed to be 80% of M and Vmax , Thus, the parameter Vmax (which allows us to estimate the total amount of PCE metabolized) is 170 mg/kg-d (~ 136/0,80, using the vmax u estimated by Buben and O'Flaherty (1985). This assumes that PCE metabolism is quantitatively similar for B6C3F1 and Swiss-Cox mice. A single, oral dose of 500 rag/kg of PCE administered to B6C3F1 mice was found by Schumann et al. (1980) to yield an average metabolized dose of 85.02 mg/kg. This is 16% greater than the value of 73.26 mg/kg predicted by Equation 5-1 using a Vmax,u estimate of 170 mg/kg. Using a Vmax,u of 136 mg/kg yields an average metabolized dose of 58.6 mg/kg. These results did not agree with the PCE metabolism data obtained by Mitoma et al. (1985) for groups of four male 4-6 week-old B6C3F1 mice. Mitoma and co-workers reported that only 22 to 37% of radiolabeled gavage doses (900 and 225 mg/kg> respectively) administered after 4 weeks of similar, daily, nonlabeled dosing, was metabolized. Fitting the Michaelis-Menten equations (54 and 5-5 below) to these two data points results in parameter estimates of Vmax " 366 mS/kS and Km " 765 mg/kg. The advantage of using the data from Mitoma and associates in estimating Michaelis-Menten metabolic parameters is that the same strain of mouse and a similar exposure protocol were used in the 1977 NCI mouse bioassay (Mitoma et al., 1985; NCI, 1977). The drawbacks, on the other hand, are (1) that the mice studied by Mitoma and colleagues were quite young and therefore perhaps not representative of mice in the 90-week NCI bioassay, (2) the metabolite recovery period used by Mitoma and co-workers was only 48 hours, SL 038997 5-13 DRAFT and (3) total radioactivity recovered by Mitoma and associates accounted for only about 80% of the radiolabeled PCE administered. For these reasons the data from Mitoma and colleagues (1983) on PCE was not considered accurate enough for use in the calculation of a metabolized dose. However, it should be kept in mind that, for mice administered comparable dose levels, a 3- to 4-fold difference in the rate of PCE metabolism has been reported (22 to 82%). This leads to a two-fold mg/kg-d to 366 mg/kg-d) in mice. difference in the calculated V max (170 For gavage studies the apparent is unaffected by the transformation from to M (see Equation 5-5 below). Under the assumptions stated above, it is possible to relate administered dose to the metabolized dose for the 1977 NCI mouse bioassay using the equation M- 1 DV _____ * ____ntax.u 0.80 D + Km (5-2) To convert M to a lifetime TWA, equation 5-2 needs to be adjusted for the specific exposure scenario being conside 1. The EPA assumed that cumulative age-specific cancer rates increase as th third power of age (EPA, 1980; Anderson et al., 1983). Thus, given a bioassay duration of L and a natural test-species lifespan of L, the cancer potency value calculated using a multistage-risk-extrapolation model would 3 be increased by the factor (L/L ) (EPA, 1980; Anderson et al., 1983). Equivalently, the lifetime TWA dose that would yield the corrected potency value would be the actual bioassay TWA dose multiplied by the factor 3 (Lg/L) In the 1977 NCI study, for example, the mice used were 5 weeks 03"ft 5-14 DRAFT old at start, and their age at the end of the bioassay was 95 weeks; the L value used was 90 weeks (NCI, 1977). The median survival time for (control) mice was 104 weeks, i.e., L - 104 wk. For the NCI (1977) data, a partial 3 lifetime-exposure-correction factor of (90/104) was used. This is close to 3 the value of (95/108) , which reflects the actual age at bioassay termination and assumes an experimental lifespan for B6C3F1 mice of about 25 months. The EPA used this approach to derive potencies for trichloroethylene (EPA, 1985b), but this approach was not used by EPA to derive a potency value for PCE carcinogenicity (EPA, 1985a; EPA Draft, 1986). Thus, the daily metabolite doses (M) in the NCI (1977) study were derived from corresponding values of applied dose, D, using the following modification of Equation 5-2; M Vmax x 5 x 22 x 90 D + Km 7 90 L104J (5-3) where M and D are in units mg/kg-d, Vmax and Km were assumed to be 170 ragAg-d and 660 mg/kg-d, respectively.' NTP (1986) Rat Study; Inhalation In the 1986 NTP study, groups of fifty 8- to 9-week-old male and female F344/N rats were exposed to PCE by inhalation at 0, 200, or 400 ppm for 6 hours/day, 5 days/week over 2 years (NTP, 1986). The exposure concentration values for males and females appear in Table 5-1, along with the SL 3g99g 5-15 DRAFT TABLE 5-2. METABOLIZED DOSE CALCULATIONS FOR SELECTED CANCER DATA SETS Average Daily Metabolized Dose (M) in mg/kg-d Study species (strain) Sex and weight (g) Experimental applied dose or cone., LLNLb Reitz and EPA Nolan EPA equivalent equivalent equivalent C PB-Px PB-PKe NCI, 1977 Mice (B6C3F1) Male 0.030 Female 0.025 0 mg/kg 536 mg/kg 1072 mg/kg 0 mg/kg 386 mg/kg 772 mg/kg 0 30.6 42.2 0 25.2 36.8 0 30.6 42.2 0 25.2 36.8 -' *- --" NTP, 1986 Mice (B6C3F1) Male 0.037 0 ppm 100 ppm 200 ppm 0 43.3 61.4 0 35.0 53.1 0 13.5 19.5 0 65.4 93.8 Female 0.032 0 ppm 100 ppm 200 ppm 0 46.7 65.8 0 35.0 53.1 0 13.5 19.5 0 67.5 96.9 NTP, 1986 Rats (F344/N) Male 0.44 0 ppm 200 ppm 400 ppm 0 11.8 17.2 0 15.7 21.2 0 14.9 18.6 a------------------ - Female 0.32 0 ppm 200 ppm 400 ppm 0 14.0 20.0 0 15.7 21.2 0 14.9 18.6 ** d/wk exposure over 78 wk of a 90-wk bioassay; for the NTF study, average administered inhalation exposure in ppm for a 6 h/d, 5 d/wk exposure over a 2-y Rioassay. LLNL refers to the Lawrence Livermore National Labs' approach used herein. CHere M - M* W/F, where M* is the "estimate of dose metabolized", which for NCI study is given in Table 9-9 of EPA (1985a) and the NTP study given in Table 4-3 of EPA (1986). W is (78/90)(5/7)(90/104) for mice in the NCI study, and W is (5/7) for mice and rats in the NTP study. F equals 0.80 and 0.54 for mice and rats, Respectively. These values for the NTP mouse and rat studies are derived from information in Tables A-4 and A-3, respectively, in EPA (1986). Metabolized doses appear to be low for mice (by a factor of about 3) due to the underestimated mouse alveolar ventilation rate (EPA, 1986; see p. 4-15 and p. A-17). Derived from the values for amount metabolized ("AM" in mg/day) given on page 6 of Reitz and Nolan (1986). 5-16 sv 39 DRAFT corresponding incidence of mononuclear-cell leukemia (MLK), the only tumor type observed to be significantly increased in rats at either the low or high dose levels. Again, the incidence values in the denominators appearing in Table 5-1 represent animals surviving at least until the appearance of the first MLK (which was at week 53 for males and at week 60 for females (NTP, 1986)). The amount of PCE metabolized in the NTP study can be estimated by fitting the Michaelis-Menten equation to the data obtained from the metabolism study in Sprague-Dawley (SD) rats conducted by Pegg et al. (1979). This assumes F344 and SD rats metabolize PCE similarly. In the Pegg and associates study, rats exposed via inhalation to 10 or 600 ppm of radiolabeled PCE for 6 hours were found to have metabolized a total of 1.87 and 36.4 mg/kg of PCE, respectively, by 72 hours post exposure. Of these totals, 1.10 (or 59%) and 18,2 (or 50%) mg/kg, respectively, were in the form of urinary metabolites. Thus, both urinary and total PCE metabolite production reportedly demonstrated saturable kinetics that may be represented by Equation 5-1. When only 2 dose-response points, say , M^ and , Mj, are available, as in the data from Pegg et al., the corresponding MichaelisMenten parameter estimates are completely specified by the following equations: Vmax M,2 M. 1 and (5-4) K m M. 1 M,2 D 1 D2 SL 039001 5-17 (5-5) DRAFT Thus, V and K are estimated to be 24.6 mg/kg and 2X4 ppm based on the urinary metabolite data for a 6-hour exposure to PCE, and 53.0 mg/kg and 273 ppm based on the corresponding total metabolite data (Pegg et al., 1979). In its updated PCE potency assessment, the EPA used the former values based on urinary metabolism to derive the total metabolized doses for the NTP (1986) rat inhalation study (EPA Draft, 1986). The CDHS estimates of V and are based on the total amount of metabolites generated by rats inhaling PCE as measured by Pegg et al. (1979). The values expressed above in rag/kg of metabolites produced following a single 6-hour exposure were also used to estimate corresponding steady-state values (in mg/kg-d) for a repeated exposure of 6 hours/day. Before estimating the rate of metabolite formation for the 1986 NTP rat bioassay, adjustments must be made to account for the interrupted exposure pattern (5 days/week) used in the NTP (1986) rat inhalation study, and to account for the fact that the rats used by Pegg et al. (1979) were lighter in weight compared to those used in the NTP study. The weight differences can be expected to result in corresponding differences in metabolic capacity. The approach used by Lawrence Livermore National Labs (and used herein) takes into account these differences in animal weight. In the context of pharmacokinetic modeling, physiological parameters such as maximum enzymatic reaction rates and related metabolic clearance rates are generally assumed to vary with basal metabolic rate in proportion to bodv surface area (or, approximately, to body weight raised to the two-thirds power), rather than to body weight per se (Gehring et al., 1978; Dedrick and Bischoff, 1980; Anderson et al., 1980; Andersen et al., 1980; Calabrese, 5-18 SL 039002 DRAFT 1983; Ramsey and Andersen, 1984; EPA, 1985a; EPA Draft, 1986; NAS, 1986). Accordingly, given an estimatedin mg/hour for an animal of weight , the corresponding predicted value for animals of weight would be Vnicix (wz-/w i. ) 2/3 or, for Vmax expressed in mg/kg-hour, Vmax (wi /wz ) 1/3 . That is, a heavier animal would be expected to metabolize less per unit body weight than a lighter one in a given amount of time. In contrast, the value of the Michaelis-Menton constant, K^, for a given compound is generally assumed to be independent of body size (in the absence of data indicating otherwise) when this constant is expressed as the reactant concentration in units mg/L blood (Ramsey and Andersen, 1984; EPA Draft, 1986; NAS, 1986), Since blood weight is proportional to body weight for animals of widely varying weights (Adolph, 1949; Dedrick and Bischoff, 1980), the independence of and body size should also be expected to hold for values expressed In terms- of mg/kg body weight. Under this assumption, however, apparent Km values expressed in terms of an inhaled air concentration (e.g., ppm), as opposed to a blood concentration, would not be expected to be constant for animals of widely varying size unless pharmacokinetic steady-state conditions applied. Upon initial exposure or after an exposure of short duration (relative to the half-life of metabolism/elimination), the maximum rate at which an inhaled toxicant reaches an organ is related to the rate of pulmonary uptake. Since weight- normalized breathing volume decreases as body weight increases, an apparent inhalational in ppm for animals of a given weight is expected to be higher for heavier animals (Anderson et al., 1980), That is, the heavier the animal the higher the air concentration required to achieve a specified target organ concentration (corresponding to the actual Kffl). SL 039003 5-19 DRAFT Respiration rate is observed to be proportional to body surface area (or, approximately, to body weight to the two-thirds power) (Guyton, 1947; Adolph, 1949; EPA, 1980b; Anderson et al., 1983; Calabrese, 1983). Accordingly, for brief, nonsteady-state and/or discontinuous exposure scenarios such as those present in the Pegg et al. (1979) and Schumann et al. (1980) metabolic studies and (to a certain extent) in the 1977 NCI and the 1986 NTP bioassays, the assumed corresponding predicted value of Km for animals of weight, v^, would be K^ (w/w^) 1/3 ppm, given an estimated apparent inhalational in ppm for an animal of weight, w^. Under this assumption, lifetime TWA equivalent values for total PCE metabolized dose, M, in the 1986 NTP rat study is derived from corresponding values of applied dose, D, using the following modification of Equation 5-3; -11/3 V M D Vmax 11 w 2J D + K w,, 1/3 x5 7 (5-6) w 1J where M is in mg/kg-d and D is in ppm, Vmav is 52.982 mg/kg-d and is 273,32 ppm, and in which w^ and w^ are the weights of the Pegg et al. (197 9,' rats (0.25 kg) and the 1986 NTP rats (0.44 or 0.32 kg), respectively. Values of D and M derived from Equation 5-6 for the 1986 NTP bioassays for male and female rats are listed in Table 5-2, along with corresponding tumor-incidence data. NTP (1986) Mouse Study; Inhalation In the 1986 NTP study, groups of fifty 8- to 9-week-old male and female B6C3F1 mice were exposed to PCE by inhalation at 0, 100, or 200 ppm for 6 5-20 03900** si* draft hours/day, 5 days/week over 2 years (NTP, 1986), These exposure concentrations appear in Table 5-1, along with the corresponding incidence data for hepatocellular carcinoma (HC) and hepatocellular adenoma or carcinoma (HAC), the only tumor types observed to be significantly increased in the mice at either the low- or high-dose levels. Again, the incidence rate denominators appearing in Table 5-1 represent animals surviving at least until the appearance of the first HC (which was at week 60 for males and week 67 for females; cf. NTP, 1986). Since one female mouse was missexed in the control group, the tumor-incidence-rate denominator used for that group appearing in Table 5-1 is 44 (NTP, 1986). In a draft form of its updated carcinogenic risk assessment for PCE, the EPA used a denominator value of 46 for the control female mice (EPA Draft, 1986). The latter value is not consistent with the data published by the NTP (1986). Lifetime TWA values for the total metabolized dose (M) of PCE for the 1986 NTP mouse bioassay are derived from Equation 5-6. The mouse data of Schumann et al. (1980) on PCE metabolism following inhalation or injection provide insufficient information for estimating V and using Equation 5-6, because only one exposure level was used for each route. The value of Vmax for the total amount of metabolites formed was assumed to be 170 mg/kg-d, estimated from the PCE oral administration data of Buben and O'Flaherty (1985). The value of K m was estimated from the rat inhalation data of Pegg et al. (1979) and was calculated to be 126 ppm (- 273.32 x 1/3 (w^/0.25 kg) ' ) where w^ represents the body weight of one mouse and 0.25 kg is the approximate weight of one rat in the study by Pegg et al. (1979). The values for mice of differing weights were estimated using Equation 5-6, where w^ and w^ are the weights of the mice from the Schumann et al. study SL 039005 5-21 DRAFT (0.0245 kg) and the 1986 NTP mice (0.037 or 0.032 kg). One of the assumptions made in using this approach is that (lacking evidence to the contrary), there are no differences in the metabolism of PCE between oral and inhalation exposure. Thus, after scaling for differences in body size, it is pr ossible to estimate the Vmax and Km for PCE metabolism based on both a mouse oral study and a rat inhalation study. The Schumann et al, (1980) data on metabolism following PCE inhalation by mice for 6 hours at 10 ppm may be used to check the estimates of V and K^. They found that a total of 14.5 mg/kg PCE was metabolized by mice within 72 hours after being exposed to 10 ppm PCE for 6 hours. The 14.5 mg/kg PCE value is 16% greater than the predicted value of 12.5 mg/kg using Equation 5-6 with a D of 10 ppm, a Vmax of 170 mg/kg, and a K of 126 ppm. For this case w^ and w^ are equivalent and the factor 5/7 relating to che NTP exposure scenario is deleted. Results of Metabolized Dose Calculations Metabolized PCE doses calculated here by Lawrence Livermore National Lab:; (LLNL) are summarized and compared in Table 5-2 to values calculated by the EPA (1985a; 1986) and by Reitz and Nolan (1986) using alternative approaches to metabolic analysis. Their values have been adjusted to correspond to a lifetime TWA total metabolized dose. For the NTP inhalation study, the metabolized doses estimated for mice in either exposure group range 5-fold. However, the actual range may be smaller, due to an underestimate of the mouse alveolar ventilation rate by 5-22 SL 039006 DRAFT the EPA (1986), which would reduce the estimate of the amount of PCE metabolized (Table 5-2). This range of a factor of 5 indicate the magnitude of some of the uncertainties involved in the modeling and parameter selection from limited data. CARCINOGENIC POTENCY EXTRAPOLATION BASED ON ANIMAL BIOASSAY DATA Carcinogenic "potency" refers to a quantitative expression of increased tumorigenic response per unit dose at very low dose levels. The following carcinogenic potency assessment is based on a quantitative analysis of animal bioassay data sets, assuming that PCE is carcinogenic to both animals and humans at low environmental dose levels. The rationale for using this assumption for PCE is well-established and is discussed in detail elsewhere (EPA, 1980; Anderson et al., 1983; CDHS, 1985). Arguments against using this assumption for PCE focus on the.possibility that the carcinogenicity of PCE in bioassays conducted at high doses may be caused primarily by increased cellular and/or subcellular proliferation (i.e., by tumor promotion or some epigenetic mechanism, rather than by initiation or some genotoxic mechanism), and that associated dose-response relationships contain a threshold below which carcinogenic effects do not occur (e.g., Schumann et al., 1980; Stott et al., 1982; Buben and 0'Flaherty, 1985; Elcombe et al., 1985; Green and Prout, 1985; Prout et al., 1985; Mirsalis et al., 1985). Low-dose potency extrapolation from dichotomous tumor-response information in selected animal-bioassay data sets was performed as described by Crump SL 039007 5-23 DRAFT and Watson (1979) to numerically fit parameters response extrapolation model: of the multistage dose- 8i -S q,d i-1 Probability of Cancer - R - 1 - e with q^ > 0 for all i. (5-7) in which g is the number of exposed groups in the bioassay, and d is the dose level at which the risk function is evaluated. Following EPA, (1980), Anderson et al. (1983) and CDHS (1985), q^ or some upper statistical confidence bound is defined as the low-dose "potency" parameter since at low doses an estimated extra risk may be obtained by multiplying q^ by the dose. The input to this program for each data set consisted of the values for total lifetime TWA applied dose, A, or metabolized dose, M, (in mg/kg-d) and corresponding tumor incidence data given in Table 5-1. For each data set GLOBAL79 was used to calculate a maximum likelihood estimate (HLE) ana a one-tailed 95% upper confidence limit (UCL) for the linear parameter ^ The MLE estimates of q^ that relate A and M to predicted tumor risk are denoted q^(A) and q^(M), respectively, and the corresponding UCL estimates are denoted q^(A) and q*(M), respectively. When dose is given in units of -1 mg/kg-d, the units of q^ and q* are (mg/kg-d) . Calculated values of (A), q(A), q^(M) and q(M) corresponding to each of the 8 bioassay data sets considered here are given in Table 5-3. The male and female rat potency values of 0.064 and 0.04 (mg/kg-d) calculated from the 1986 NTP rat-leukemia data are somewhat larger than those calculated from both the 1977 NCI and the 1986 NTP mouse data. The maximum potency value of 0.064 (mg/kg-d) is about 9 times larger than the Si 39ooa 5-24 DRAFT smallest calculated potency value of 0.0073 (mg/kg-d) derived from HC incidence data for female mice in the 1986 NTP bioassay. In contrast, the calculated UCL potency values corresponding to the applied dose approach (q^(A)) -1 span only a 2-fold range (0.0026 to 0.0064 (mg/kg-d) ). Thus, the metabolized dose tends to indicate a different sensitivity between rats and mice while the applied dose indicates a more similar susceptibility to the carcinogenicity of PCE. Time-to-tumor data are readily available for the NTP mouse and rat bioassay data, as well as for the NCI mouse bioassay data. Thus, it is possible to use a more complex, time-dependent version of the multistage model that takes Into account the observed differential survival of control and exposed bioassay groups. In the NTP bioassays, survival was only slightly decreased in the PCE-exposed groups, so the multistage time-to-tumor model is not used as the basis for potency extrapolation in the present analysis. However, for comparative purposes a time-dependent version of the multistage model, WEIBULL82 (Howe and Crump, 1983; Crump and Howe, 1984) was used to derive potency estimates from the. NCI and NTP bioassay data. These potency values are given in the last two columns of Table 5-3. Note that the timedependent potency values are similar to values based on the time-independent form of the multistage model (also shown in Table 5-3). These timedependent values differ from time - independent values by no more than a factor of 1.5. The largest potency value, 0.071 (mg/kg/day) ^ for MLK in male rats, represents an increase of about 11% over the corresponding q. SL 039009 5-25 TABLE 5-3 CANCER POTENCIES ESTIMATED FROM SELECTED DATA SETS Administered Dose Metabolized Dose Time -Independent Time-Dependent Analysis Analysis Data Set Oral NCI (1977) B6C3F1 Mice MLE 9X(A) 95% UCL qf(A) MLE qx(M) 95% UCL q*(M) MLE q^H) 95% UCL qf(M) Male - Hepatocellular Carcinoma 0.0046 0.0064 0.024 0.032 Female - Hepatocellular Carcinoma 0.0023 0.0030 0.017 0.022 Inhalation NTP (1986) B6C3F1 Mice <_n i Male - Hepatocellular carcinoma 0.0025 0.0035 ro 01 Female - Hepatocellular carcinoma 0 0.0028 0.011 0 0.015 0.0073 0.012 0 0.022 0.0072 DRAFT Male - Hepatocellular adenoma/ carcinoma 0.0043 0.0059 0.0062 0.024 0.0032 Female - Hepatocellular adenoma/ tn carcinoma r4 0.00044 0.0039 NTP (1986) F344/N Rats 0 0.0098 0.0088 0X 0 & t 0 Male - Mononuclear-cell leukemia 0.0022 0.0040 0.037 0.064 0.043 Female - Mononuclear-cell leukemia 0.0015 0.0026 0.024 0.040 0.024 MLE - maximum-likelihood estimate; 95% UCL -= one-tailed 9 5% upper confidence limit. 0.026 0.012 0.071 0.040 DRAFT CARCINOGENIC POTENCY IN TERMS OF HUMAN APPLIED DOSE Assuming that the above calculated values of q*(M) are adequate to approximate expected tumor incidence among animals subjected to very low levels of effective dose M, three issues need to be addressed: first, the application these calculated potency values across exposure routes; second, the extrapolation of potency from animals to humans; and third, determination of the relationship between applied dose, D, and metabolized dose, M, for humans. Each of these issues will be addressed in turn. Exposure Route Various physiological processes, such as the "first pass effect" involving liver-mediated activation or detoxification, may result in differing values of metabolized dose, given the same applied dose but administered via different exposure routes. If biotransformation of PCE to a reactive metabolite is solely responsible for the carcinogenic effects of PCE, the actual exposure route and administered dose may not be as important as the determination of the metabolized dose. Different routes of exposure can allow differing amounts of PCE to reach the sites of metabolism or the target tissues, and thus effect the calculation of the metabolized dose. In this document it is assumed that a given dose of PCE is equipotent regardless of exposure route, as long as the distribution of this dose among susceptible target tissues is not affected by the exposure route. Comparison of the potency estimates for both the applied dose and the metabolized dose for the mouse gavage study and the mouse inhalation study indicate that the route of exposure may not be an important factor for PCE SL 039011 5-27 DRAFT (Table 5-3) (NCI, 1977; NTP, 1986). However, for the purposes of risk assessment of inhaled PCE, the range of the inhalation studies will be used. Interspecies Dose Equivalence Following the suggestion of Mantel and Schneiderman, the EPA and the California Department of Health Services assume that mg/(surface area) is an equivalent measure of lifetime TWA dose between species for carcinogens (Mantel and Schneiderman, 1975; EPA, 1980b, 1984a; Anderson et al., 1983; CDHS, 1985). Specifically, for purposes of carcinogen-risk assessment, the EPA assumes that a TWA lifetime dose (expressed as a dose rate in units of mg/kg ' ) is equivalent between species, since surface area is rough .y proportional to the 2/3rd power of body weight. Given this assumption and a lifetime metabolized dose rate for an animal of M^ mg/kg-d, then the equivalent human TWA dose rate would be M^/f, where the dose - equivalence factor f is here given by in which W. and W are the weights of humans (generally assumed to be 70 kg) na and the test animal, respectively. In the absence of definitive empirical data, however, it Is simply not known whether a mg/(surface area) or a mg/kg interspecies dose-extrapolation assumption better reflects reality in the context of extrapolating tumor response data in animals to anticipated response in humans. Existing data 5-28 SL 039012 DRAFT do not rule out either approach. For comparative purposes extrapolation of potency values from animals to humans is here carried out using both a dose per body weight (BW) and a dose per body surface area (SA) extrapolation method. Human Metabolism To estimate the carcinogenic potency of PCE in humans from q^(M), additional assumptions need to be made regarding human metabolism. The assumptions and extrapolations reveal additional uncertainties to the potency calculation when the metabolized dose is used. Several studies have examined the metabolism of PCE in humans following a controlled inhalation exposure to this compound. As discussed in Section 2, several PCE metabolites have been identified in human urine, the major one being trichloroacetic acid (TCA). There is a possibility that other metabolic products of' PCE in humans have not been identified. Some human exposure studies have sought to identify total urinary trichloro-compounds produced after inhalation exposure to PCE using a nonspecific, Fujiwarareaction analytical method (e.g., Ikeda and Imamura, 1973), but the actual products measured by this technique are in doubt (EPA, 1985a). The studies used here to estimate the extent of human PCE metabolism consist of four relatively recent studies that quantified only the cumulative amount of TCA produced for an observation period less than one estimated half-life (typically about 3 days), during and after a controlled PCE inhalation exposure (Ogata et al., 1971; Fernandez et al., 1976; Monster et al., 1979; Bolanowska and Golacka, 1972). SL 039013 5-29 DRAFT Before reviewing the above studies, three issues concerning how the results of these studies may be used to estimate the degree to which FCE in humans is metabolized during a steady-state (continuous) exposure will be discussed. These issues are: the excretion half-life of PCE in humans, estimation of total PCE metabolism and use of continuous versus peak exposures. Excretion Half-life in Humans The first of these issues involves the decay half-life of urinary metabolite formation. That is, the time required for half of the TCA formed from PCE metabolism to be excreted in the urine. Because the metabolism rate of PCE in humans is very slow, it is unlikely that the studies discussed herein were actually able to quantitatively determine the total amount of PCE metabolized following exposure. (Exposure time, t, for each of the studies discussed is listed in Table 5-4.). To extrapolate from the existing data, the remaining TCA produced after the end of urine collection had to be approximated. This was done by integrating the TCA production rate as a function of time from the termination of collection to time equals infinity, where this function is assumed to represent a simple first-order decay process. This is the same assumption used by the EPA (EPA, 1985a; EPA 1986) in assessing the results of the Bolanowska and Golacka (1972) humanexposure study. The EPA assumed a half-life for urinary TCA excretion of 100 hours, apparently based on a presumed empirical range of approximately 65 to 144 hours for this process (EPA, 1985a, where the TCA decay half-life is stated to be 144 hours). The lower end of this range is derived from studies indicating a terminal decline in TCA excretion with a half-life of 5-30 SL 039014 DRAFT from 65 to 90 hours in volunteers exposed to 72 and 144 ppm PCE for 4 hours (Monster et al., 1979). The upper value of 144 hours is derived from the study of Ikeda and Imamura (1973) as the mean of measured total urinary trichloro-compound decay half-lives of 13 subjects occupationally exposed to PCE, where the observed range was reported to be 123 to 190 hours. In this study, however, TCA itself was not measured, but total trichloro-compounds were measured using the nonspecific type of Fujiwara reaction method whose dependability has been questioned (EPA, 1985a). The values based on the studies by Monster and co-workers indicate that the rate of PCE release from adipose tissue, the most highly retentive tissue type for PCE, is an exponentially decaying function of time with a half-life of about 71.5 hours (Monster et al., 1979; Guberan and Fernandez, 1974). Thus, a value of 100 hours may overestimate the expected late-stage decay half-life of TCA excretion in humans. Accordingly, for the purpose of the metabolized dose analysis performed herein, a value of 75 hours was used as the presumed approximate late-stage TCA excretion half-life for humans. . Urinary Metabolites as a Fraction of Total Metabolites The second issue addresses the uncertainties of estimating total metabolism of PCE strictly from TCA measurements. It involves developing the relationship in humans of total PCE metabolized to urinary metabolites, and estimating the proportion of TCA metabolites excreted in the urine as compared to total elimination by all routes. For the purposes of this analysis it was assumed that 58% of the metabolized dose was found in the form of urinary TCA following low levels of exposure to PCE. The value of 58% is the average fraction of total metabolites found in the form of SL 039015 5-31 TABLE 5-4. COMPARISON OF HUMAN PCE METABOLISM AFTER INHALATION EXPOSURES Administered Dose a Urinary metabolite b Estimated total metabolites Exposure period Concentration Urine col lection Inspired Dose Study Ogata et al., 1971 (n=4> (h) 3 <*> 87 (h> (mg) 67 594.6 U(T) (mg) 4.3 U(E) (mg) 8.6 H (mg) H6 53 kg M P 53 kg 14.8 0.42 0.38 fj P corr 2.5 3.3 Fernandez et al., 1976 (n=2) 8 150 72 2734 24.6 35.4 61.0 0.65 0.63 2.2 3.5 DRAFT <VL 0 3 9 0 1 6 Monster et at., 1979 4 72 70 656.1 6.0 10.5 18.1 0.35 0.31 2.8 4.0 <n=6> Monster et al., 1979 (n*6) 4 144 70 1312 11.0 20.0 34.5 0.67 0.60 2.6 3.8 Bolanowska and Gotacka, 1972 6 56.4 tn=2> Weighted mean (weight * n, the sum of n S.D. of mean * 20) 23.5 771.1 3.5 14 24 0.34 0.25 3.1 3.8 2.6 3.7 0.05 0.06 99% Upper (2-tailed) confidence Unit of wean (df=19) 2.8 3.9 U(T) = observed cumulative urinary Metabolite (trichloroacetic acid only) production by any time T after exposure. See text for derivation of ll(T) values. U(E) * estiMated total urinary Metabolites if collection period Mere extended to infinity, t) Estimated total Metabolites produced if the monitoring period Mere extended so time = infinity. M = total amount of metabolites formed, t = amount of metabolites formed in 6 hrs. * = predicted total metabolite dose produced if the collection period Mere extended to infinity, and assuming that H = D V /(D K ), where PV 8.099 mg/kg and K = 273.32 ppm. 'MX m c max m 33 ^ Assumes an alveolar ventilation rate of 5.6 L/win = 2.016 m /6h, and 6.78 mg PCE/m per ppm. eP = the percent of respired dose that is metabol;ted. P - where P is corrected to reflect a steady-state, low level exposure. The correction factor is based on equation 5-9. corr DRAFT TABLE 5-5. PREDICTED UPPER 95% UPPER CONFIDENCE LIMITS ON POTENCY (MG/KG-D) SUMMARY (ASSUMING HUMANS METABOLIZE 4% OF ABSORBED PCE). Study species Tumor strain type a Sex and weight (kg) Metabolized dose^ - qJ(M) BWd SA* Human applied doseC - q|(A) BWd SAe NCI, 1977 Mice HC B6C3F1 NTP, 1986 Mice HC B6C3F1 NTP, 1986 Mice HAC B6C3F1 NTP, 1986 Rats MLK F344/N M 0.030 F 0.025 M 0.037 F 0.032 M 0.037 F 0.032 M 0.44 F 0.32 0.032 0.022 0.015 0.0073 0.024 0.0098 0.064 0.040 0.42 0.31 0.19 0.095 0.30 0.13 0.35 0.24 0.0013 0.00089 0.00060 0.00029 0.00096 0.00039 0.0026 0.0016 0.017 0.012 0.0074 0.0038 0.012 0.0051 0.014 0.0096 s. HC - hepatocellular carcinoma, HAC - hepatocellular adenoma or carcinoma, MLK - mononuclear-cell leukemia. Human equivalent lifetime, time-weighted-average metabolized dose, M, in mg/kg-d. Q Human equivalent lifetime, time-weighted-average applied dose, D, here assumed to be equal to M/0.04 when D is very small; thus, q^(D) - 0.04q^ (M) for very small D. in mg/kg-d, is BW -- Body Weight inter-species dose extrapolation method; equivalent doses assum d to be in mg/kg, so M. -M numan animal eSA - Surface Area intarspecles dose extrapolation method: equivalent doses assumed to be in rag/kg , so M^^ - M^^f (animal weight)/70 kg]17 . SL 039017 5-33 DRAFT urinary metabolites produced by rats exposed to 1 mg/kg PCE by intubation (yielding 57.1% urinary metabolites) or 10 ppm PCE for 6 hours (yielding 58.9% urinary metabolites), in the study by Pegg et al. (1979). Rat, as opposed to mouse, data are used as the basis for this assumption only because of the greater similarity in body weight between rats and humans. However, this assumption may greatly underestimate the fraction of PCE metabolized in humans. Since humans produce lower levels of trichloro urinary metabolites thando rats, they may produce more nonurinarv metabolites or more urinary metabolites that are not trichloro- compounds. As indicated in studies below and in Section 2, only TCA has been used to estimate urinary excretion of PCE metabolites. The production of other urinary metabolites in humans, although documented, has not been adequately inv stigated as to their relative importance in the metabolism of PCE, Continuous Versus Peak Exposures The third and final issue concerns the assumptions about a continuous human- exposure scenario. Results of the human studies considered below were compared to predictions of PCE metabolism based on extrapolation from the data on PCE metabolism in rats exposed to 10 or 600 ppm PCE for 6 hours (Pegg et al., 1979), using the assumptionsmade in Equation 5-6. The V calculated from this rat study was 53.0 mg/kg. Using Equation 5-6, the equivalent human V would be the latter value multiplied by (0.25/70) 1/3 or 8.10 mg/kg for a 6-hour exposure to PCE. Similarly, the corresponding rat inhalational apparent value of 273 ppm would be divided by (0.25/70) ' to yield an equivalent (initial-exposure) human Michaelis- Menten constant of 1790 ppm. Using these parameter values, the predicted 5-34 SL 039018 draft metabolized dose, M (in mg/kg), as a function of applied dose, D (in ppm), is presented in Table 5-4 for each of the various studies considered below. The brief exposure periods used in the human experiments were conducted at relatively high concentrations compared to the average ambient concentrations of PCE identified in air. At lower doses, a greater percent of the dose is likely to be metabolized. Using the Michaelis-Menten relationship, the maximal metabolized dose per unit applied dose occurs as D approaches 0 in the equation M - D(Vmax/Km ). Thus, an estimate of the maximum factor by which the ratio M/D is increased at very low doses is given by D Vmax K Lim M - . m D-*0 0 M 'D V max -1+D . m (5-9) D + Km This factor is used iin Table 5-4 to adjust the percent, P, of respired dose, I, that is metabolized to yield a new percentage value, P which is corrected to reflect the steady-state, very low-level exposure conditions that relate to the human environmental exposure scenario being modeled. In this factor, was assumed to be 273 ppm, based on the Pegg and associates (1979) metabolism data (this value was not adjusted upward to account for initial rapid uptake prior to steady-state, since Pcorr is meant to reflect steady-state exposure conditions). Having addressed these issues and assumptions regarding human metabolism, the four studies used for calculating the rate and extent of human PCE metabolism will now be reviewed. SL 039019 5-35 Ogata and Co-workers (1971) DRAFT Urinary excretion of TCA was measured by Ogata and co-workers in four volunteers exposed to 87 ppm PCE for 3 hours (Ogata et al., 1971). From the graphical information presented in this study, the amount of TCA excreted over the 67-hour collection period was calculated from the area under the time versus TCA excretion-rate curve, and found to be about 4.3 mg. This value is presumed to be excess excretion over background, but a pre-exposure TCA excretion rate was not provided in this study. At 67 hours, the observed excretion rate was about 0.040 mg/hour, so that remaining excretion from that time on is estimated to be (0.040 mg/hour x 75 hours)/ln(2) - 4.3 mg. Total metabolized dose, , (assuming that urinary metabolites in humans represent 58% of total metabolites, as was seen in the rat study by Pegg et al,, 1979), from this 3-hour exposure was (4.3 + 4.3)/0.58 - 14.8 mg, which when multiplied by (6/3) yields an approximate equivalent 6-hour metabolized dose, Mg, of 0.42 mg/kg (assuming a 70-kg average subject weight). Fernandez and Co-workers (197.61 Urinary excretion of TCA was measured by Fernandez and colleagues in two volunteers exposed to 150 ppm PCE for 8 hours (Fernandez et al., 1976). This study reported an average total TCA excretion of 24.6 mg over the 72hour collection period applied. Urine was not collected during the exposure period, thus the total metabolites produced were underestimated. During the period from 48 to 72 hours post-exposure the reported TCA excretions for the 2 subjects were 3.63 and 3.70 mg, equivalent to an average excretion rate of 5-36 39020 DRAFT about 0.15 mg/hour. It was estimated from the data that by 72 hours the average excretion rate had declined to about 0.10 mg/hour, so that remaining excretion from that time on is estimated to be (0.10 mg/hour x 75 hours)/ln(2) - 10.8 mg. Total metabolized dose, Mg, (assuming a proportion similar to rats), from this 8-hour exposure was (24.6 + 10.8)/0.58 - 61.0 mg, which multiplied by (6/8) yields an approximate equivalent 6-hour metabolized dose, , of 0.65 mg/kg (again assuming a 70-kg average subject weight). Monster and Co-workers (1979^ Urinary excretion of TCA in the Monster and colleagues study was measured in six volunteers at rest exposed to 72 ppm PCE for 8 hours and again in these same volunteers at a later date exposed at rest to 144 ppm PCE for 4 hours (Monster et al., 1979). (Data from this study that were obtained from volunteers subjected to physical activity routines are not used for the present analysis). From the published graphical urinary excretion information, it was estimated that average total TCA excretions of 6,0 and 11.0 mg were observed among the 72 ppm and 144 ppm exposure groups, respectively, over the 70-hour collection period applied. (It appears that urine was not collected during the exposure period. These estimates take into account the background 0.025 rag/h TCA excretion rate observed in this study.) During the 24-hour period from 46 to 70 hours post-exposure, the reported TCA excretions were about 1.0 and 2.0 mg (and fairly constant over the period 22- to 70-hours post-exposure) for the 72 ppm and 144 ppm exposure groups, equivalent to average terminal excretion rates of about 0.042 and 0.083 mg/hour. Thus, the corresponding remaining excretions from SL 039021 5-37 DRAFT 70 hours on are estimated to be (0.042 mg/hour x 75 hours)/ln(2) - 4.8 mg and (0.083 mg/hour x 75 hours)/ln(2) - 9.0 mg, respectively. Corresponding total metabolized doses, M^, from these 4-hour exposures are (6.0 + 4.5)/0.58 - 18.1 mg and (11.0 + 9.0)/0.58 - 34.5 mg, respectively, which multiplied by (6/4) yields corresponding approximate equivalent 6-hour metabolized doses, Mg, of 0.35 and 0.67 mg/kg (using the reported average subject weight of 77 kg). Bolanowska and Golacka (1972) In evaluating the human metabolism study of Bolanowska and Golacka (1972), 3 the EPA assumed that five subjects were exposed to 390 mg/m PCE for 6 hours and were followed for 20 hours; the EPA did not take into account the background TCA excretion rates observed in the study. Actually, urinary excretion of TCA was measured from only 2 volunteers exposed to 391 and 374 3 mg/m PCE for 6 hours (interspersed with two 30-minutes "rest" periods). Urine from these subjects was collected for a total of 22 and 25 hours, respectively, including the exposure time. Background TCA excretion rates for the two subjects are given as approximately 0.035 and 0.010 mg/h. Taking these background rates into consideration, it is estimated from the graphical urinary excretion information given in this study that average total TCA excretions of 2.3 and 4.7 mg were observed from these subjects over the (approximately) 23.5-hour average collection period. During the final hours of urine collection, excretion rates for these subjects were approximately 0.045 and 0.15 mg/hour (where 0.15 mg/hour represented a weighted average of 9 hours at 0.14 mg/hour followed by 3 hours at 0.19 mg/hour). Thus, the corresponding remaining excretions from about 24 hours 5-38 SL 039022 DRAFT on are estimated to be (0.045 mg/hour x 75 hours)/ln(2) - 4.9 mg and (0.15 mg/hour x 75 hours)/ln(2) - 16.2 mg. Corresponding total metabolized doses, Mg, from these 6-hour exposures are thus 12 mg ((2.3 + 4.9)/0.58) and 36 mg ((4.7 + 16.2)/0.58), yielding an average value of about 0.34 mg/kg for this study (assuming an average subject weight of 70 kg). Metabolic Parameters Estimated from Different Human Data Sets Table 5-4 summarizes the metabolic data obtained from the studies discussed herein and presents comparisons between the predicted and the observed total metabolite dose for each experiment. The internal consistency of the general approach used in the metabolic analyses is supported by the observation that the predicted total metabolite doses for a 6-hour exposure to PCE for humans are quite similar to the corresponding empirically derived values, Mg, as shown in Table 5-4 (correlation coefficient r = 0.99, degrees of freedom - 3, p < 0.01). To calculate the average percentage, P, of PCE metabolized on the basis of the studies considered here, the inspired dose I had to be estimated from the PCE concentrations specified in these studies. Although PCE retention values were given in the Ogata et al. (1971) study, the basis for their calculations was not provided. Fernandez and associates used an estimated alveolar ventilation rate of 5.6 L/minute (equivalent to 340 L/hour or approximately 4.8 L/kg-hour for a 70 kg person) in their calculation of retained dose, a value close to those currently used in proposed PB-PK models for human metabolism (Fernandez et al., 1976; EPA Draft, 1986; NAS, 1986). The 5.6 L/minute value is also close to the effective value used by SL 039023 5-39 DRAFT Monster and co-workers who assumed that alveolar retention was approximately equal to 60% of the minute volume (which averaged 10 1.8 L/minute among their 6 subjects) (Monster et al., 1979). To calculate inspired dose, I, in mg as a function of applied concentration in ppm for the purpose of the present analysis, the alveolar ventilation 3 rate of 5.6 L/minute is assumed, and 1 ppm PCE - 6.78 mg/m . This 5.6 L/minute estimate has been used in several pharmacokinetic models, but underestimates the average breathing rate of humans (20 L/minute) by almost 3-fold. The resulting values of P for the 5 data sets range from about 2 to 3%, with a weighted average of 2.6%, where the values of n, the number of people studied in each experiment, are the weights (see Table 5-4). On the basis of the predictive model adjusted for continuous human exposure discussed above, corrected values, ^corr> fr t*'e average percent PCE metabolized range from about 3 to 4%, with a weighted average of 3.7%. The upper 99% confidence limit on the average P was calculated to be approximately 4%, assuming a normal error distribution of the mean and 19 degrees of freedom associated with the sample estimate (i.e., 20 persons studied, minus one). The UCL was approximated in this way because needed data for each individual studied could not be obtained for most of the studies considered. The calculated UCL value of 4% is also equal to the highest value of P calculated here, derived from the study of Monster et al. (1979). It should be emphasized here that these results are based on relatively brief experimental exposures to relatively high concentrations of PCE in air. Linear extrapolation of animal studies from high experimental exposure 5-40 SL 039024 concentrations DRAFT to low ambient concentrations suggests that essentially 100% of the PCE may be metabolized at ambient levels by mice and rats. At ambient levels humans would be expected to metabolize PCE to a much greater extent than that reported in brief, high exposure studies. The Pegg and associates inhalation (and oral) studies showed that for rats metabolism increases with decreasing PCE exposure (Pegg et al., 1979). In these studies, rats were exposed to either 10 or 600 ppm PCE for 6 hours. With a 60-fold decrease in concentration, there was a 2.7-fold increase in percent PCE metabolized. These acute exposure experiments were conducted at concentrations 10,000- to 100,000-fold greater than ambient levels (0.001 ppm). Assuming an inverse linear relationship between percent metabolism and concentration, rats and mice would be predicted to metabolize essentially 100% of the PCE at concentrations below 0.1 ppm. Similar estimates can also be made for humans by applying the relationship of percent metabolism and.concentration established for rats. The fraction of applied dose metabolized in humans, based on measured urinary trichloro compounds is estimated to be 4%, based on human experiments where exposure concentrations averaged 102 ppm. This exposure concentration is 100,000 times greater than ambient levels. Using the same linear relationship between metabolism and exposure conditions described above, extrapolating from the limited human exposure data to ambient PCE concentrations implies that humans would be expected to metabolize more than 60% of the PCE absorbed. This 15-fold increase in the expected extent of PCE metabolism at ambient concentrations indicates some uncertainty in the use of the PCE pharmacokinetic data for risk assessment. SL 039025 5-41 DRAFT Using a PB-PK modeling approach, Bogen and McKone estimated that the maximum plausible rate of POE metabolism in humans at steady-state for a given, extremely low PCE concentration (C.^) i-n a^r *-s a physiologically defined fraction (f**) of the alveolar input rate, mg/hour, where is the alveolar ventilation rate in L/hour (Bogen and McKone, 1987), This fraction is defined by f** _ m VQ1 +1 (5-10) in which is the blood flow to liver (metabolizing) tissue in L/hour and is the PCE blood/air partition coefficient for humans. This relationship reportedly holds for continuous, steady-state infusion by other exposure pathways as well. Note that Equation 5-10 assumes that the rate of PCE metabolism becomes infinite as PCE concentration in air approaches z^ro. Using the parameter specifications of - 3,8Q^ and P^ - 10.3 taken from recent studies of PCE pharmacokinetics (Reitz and Nolan, 1986; Ward et al., 1987), it follows that for PCE, f** - 73%. Thus, the physiological upper bound on metabolism of inspired PCE is predicted to be near 73%, i.e., 0.7 C^Q^ at any given extremely low concentration C^. Using a PB-PK approach and an analysis of the data of Ikeda and co-workers and Ohtsuki and colleagues on PCE metabolism among occupationally exposed humans, it was estimated that the actual fraction of PCE metabolized at steady-state for extremely low levels of PCE absorption (by any route) is likely to be between 2% and 50% (Ikeda et al., 1972; Ohtsuki et al., 1983; Bogen and McKone, 1987). This conclusion is consistent with predictions of other investigators who used dynamic PB-PK modeling (EPA Draft, 1986; Hattis et al., 1987) . 5-42 SL 039026 DRAFT PCE Carcinogenic Potency as a Function of Human Applied Dose (Taking PCE Metabolism Into Account) To estimate the carcinogenic potency of PCE in humans for a given steadystate applied dose (in mg/kg-day) administered by ingestion or inhalation, the value of 4% derived above was used as the fraction of applied dose that is metabolized in humans. Under this assumption, extrapolated potency is reexpressed in Table 5-5 as a function of human applied dose for each bioassay data set considered here under both the body weight (BW) and surface area (SA) assumptions for extrapolating equipotent doses between species. Using the BW approach, the calculated values of q*(M) (based on animal bioassay data) can be directly as estimates of human carcinogenic potency. That is, by this approach, a human metabolized dose in mg/kg has equal tumorigenic potency compared to an animal metabolized dose in mg/kg, regardless of body weight differences. Accordingly, the values o f qf(M) appearing in Table 5-3 are repeated in Table 5-5, which summarizes the predicted tumorigenic potency of PCE to humans as calculated in this analysis based on the 8 bioassay data sets considered here. Corresponding values of q|(M) based on the SA approach are also listed in Table 5-5. The latter values range from 0.095 to 0.42 (mg/kg-d) -1 (a 4.4-fold range) and so are somewhat more homogeneous than the q*(M) values based on the BW approach. The range of q*(M) values is reduced by a factor of about 2 substituting the SA for the BW approach (but note that the tumor incidences in control rats were much higher than those for mice, which influences potency calculations with the model used here). Using the SA approach, the SL 039027 5-43 DRAFT .i maximum qJ(M) value is 0.42 (mg/kg-d) based on HC incidence in male mice in the 1977 NCI bioassay. This value is about 6.6 times greater than the largest value obtained using the BW approach. Table 5-6 compares the 95% UCL potencies from Table 5-5 with corresponding values based on different approaches to calculating dose than the SSPK approach used to generate the potency values appearing in Table 5-5. In interpreting the values presented in Table 5-6, 4 points should be kept in mind. 1) The EPA approach to the NCI bioassay data implies that following ingestion humans metabolize PCE similar to micf (the "EPAI" method in Table 5-6), whereas following inhalation exposure humans are assumed to metabolize a somewhat lesser amount of PCE than mice or rats (the "EPAR" method in Table 5-6) (EPA 1985a; NCI, 1977). 2) All approaches based on the use of metabolic data involving respiratory exposure to PCE (namely, the "EPAR", "EPPK", "RNPK", and "SSPK" methods referred to in Table 5-6) were adjusted to reflect a single assumption regarding human alveolar respiration. The alveolar ventilation rate (AVR), assumed here to be 5.6 L/minute, is the same value used by the "SSPK" method and explained above in the subsection dealing with the comparison of human PCE metabolism after inhalation. For reference, it is noted that the actual AVR's 5-44 SI* 039028 DRAFT TABLE 5-6. TUMORIGENIC POTENCY OF PCE: SUMMARY OF VALUES BASED ON DIFFERENT APPROACHES TO DOSE CALCULATION. S tudy/ Species/ Sex Tumor Calc. type3 method*5 95% UCL potency0 values As a function of metabolized dose (M) As a function of human applied dose (A): q*(M) in (mg M/kg-d)"1 q*(D) (mg D/kg-d) f cr BW SAS f BW SAg NCI, 1977 Mouse Male Mouse Female NTP, 1986 Mouse Male Mouse Female Mouse Male Mouse Female HC HC HC HC HAC HAC SSPK EPAI EPAR APPL SSPK EPAI EPAR APPL SSPK EPAR EPPK RNPK APPL SSPK EPAR EPPK RNPK APPL SSPK EPAR EPPK RNPK APPL SSPK EPAR EPPK RNPK APPL 0.032 0.032 0.032 0.022 0.022 0.022 -.- 0.015 0.018 0.048 0.0099 0.0073 0.0093 0.025 0.0050 0.024 0.029 0.077 0.016 0.0098 o.on 0.033 0.0067 0.42 0.42 0.42 0.31 0.31 0.31 0.19 0.22 0.59 0.12 0.095 0.12 0.32 0.064 0.30 0.36 0.95 0.20 0.13 0.17 0.43 0.087 0.0013 0.0064 0.00066 0.0064 0.00089 0.0044 0.00046 0.0030 0.00060 0.00037 0.0047 0.00040 0.0035 0.00029 0.00019 0.0024 0.00020 0.0028 0.00096 0.00060 0.0075 0.00064 0,0059 0.00039 0.00026 0.0033 0.00027 0.0039 0.017 0.085 0.0088 0.085 0.012 0.062 0,0064 0.042 0.0074 0.0046 0.058 0.0049 0.043 0.0038 0.0025 0.031 0.0026 0.036 0.012 0.0074 0.093 0.0079 0.073 0.0051 0.0034 0.042 0.0035 0.051 039029 SL 5-45 Study/ Species/ Sex DRAFT TABLE 5-6. (Continued) c 95% UCL potency values AS a function of metabolized dose (M) As a function of human applie< dose (D): Tumor Calc. type a method*5 q*(M) in (mg M/kg-d) 1 q*(D) (mg D/kg-d) BWf SAS BWf SA Rat Male Rat Female MLK MLK SSPK EPAR EPPK APPL SSPK EPAR EPPK APPL 0.064 0.051 0.057 0.040 0.037 0.041 "* 0.35 0.28 0.31 0.24 0.22 0.25 *" 0.0026 0.0016 0.0056 0.0040 0.0016 0.0011 0.0040 0.0026 0.014 0.0085 0.030 0.022 0.0096 0.0068 0.024 0.016 HC - hepatocellular carcinoma, HAC - hepatocellular adenoma or carcinoma MLK - mononuclear-cell leukemia. ^Methods of calculation used are as follows (see text for more detailed discussion): SSPK - simple steady-state pharmacokinetic method, corresponding values taken from Table 5-5. EPAI - EPA-equivalent ingestion method: assumes that M - 20% of ingested PCE is metabolized EPAR * EPA-equivalent respiratory method; assumes that for humans M - 0.0207A or M - 0.0306A for potencies based on mouse or rat data, respectively, and the alveolar ventilation rate (AVR) is 5.6 L/min. RNPK - Reitz & Nolan-equivalent PB-PK method; assumes that for humans M -- 0.04A and AVR - 5.6 L/min. EPPK - EPA-equivalent PB-PK method; assumes that for humans M - 0.0977A and AVR - 5.6 L/min. APPL - applied dose method; corresponding values taken from Table 7-1, where they are labelled qt(A). '95% UCL ** one-tailed 95% upper confidence limit. Human equivalent lifetime, time-weighted average metabolized dose, M, in mg/kg-d, & defined as a function of D as described above, 'Human equivalent lifetime, time-weighted - average applied dose, D (or A, in the context of the APPL method described above). `BW Body weight inter-species dose extrapolation method; equivalent doses assumed to be in mg/kg, so M, -M , ^ SA - Surface Area interspecies^Hose ixtrapolation method; equivalent doses 1/3 assumed to be in mg/kg , so M,numan - Manima,l [(animal weight)/70 kg 5-46 SL 039030 DRAFT used in the "EPPK" and "RNPK" approaches were 3.42 L/minute and 5.80 L/minute, respectively. No assumption regarding AVR was incorporated into the "EPAR" method, since that approach extrapolated directly from experimental human metabolic data to predicted risk based on animal-bioassay data for which metabolized doses were calculated (EPA 1985a). Standardization using the 5.6 L/minute AVR assumption was undertaken for the purpose of presenting meaningful comparisons in Table 5-6, It is emphasized that it would be incorrect to use any of the potency values associated with any of the respiratory methods just discussed (viz., the "EPAR", "EPPK", "RNPK", or "SSFK" methods) with an AVR (or other respiratory volume clearance rate) not equal to 5.6 L/minute. 3) Both the "EPAI" and "EPAR" methods were adjusted in the context of the 1977 NCI bioassay data sets to incorporate the partial method, 3 lifetime factor of (104/90) used in the "SSFK" * again to facilitate a meaningful comparison of alternative potency calculations. 4) If the assumption is made that humans (and animals) metabolize some fraction, fm (besides the 4% value assumed in this analysis), of all respired or ingested PCE at very low levels of applied PCE dose, the appropriate upper-limit potency value to use would be calculated as q*(D) - f q*(M) based on the values of q*(M) in Table 5-6. Such alternative SL 039031 5-47 DRAFT assumptions are plausible, given that there are no animal or human experimental data that specifically address the issue of what fraction of an extremely small applied FCE dose (e.g., continuous exposure to one ppb in air) is metabolized. A total of 112 alternative potency values appear.in Table 5-6, with 48 values representing potency expressions, q*(M), in terms of metabolized dose and 64 values representing potency expressions, q^(D), in terms of applied dose (or PCE potentially available for uptake). The q*(M) values range from 0.0050 to 0.42 (mg M/kg-d)or an 84-fold range, and the q*(D) values range from 0.00019 to 0.085 (mg D/kg-d) , or a 447-fold range. Clearly then, the model, study, route of exposure, and tumor type taken to calculate the appropriate dose in dose-response assessment is a significant factor in cancer-risk extrapolation. Recommendation The staff of CDHS recommends that*-the metabolized PCE dose not be used in the recommended range of risks for several reasons. The metabolized doL adjustment does not reduce the uncertainty of the final potency values primarily due to poor characterization of PCE metabolism in humans. The dose calculation uncertainties must be considered in addition to the genem.i uncertainties of low dose extrapolation. The current data available on clr quantity of PCE metabolized by different pathways across species remains uncertain. This uncertainty is evident in the very large range of thc potuncies estimated using different pharmacokinetic techniques. 1. summarize, the following concerns remain: 1) The metabolized dos 5-48 DRAFT calculated by several methods, produces a 5-fold range of doses in male mice (19.5 mg/kg-day to 93,8 mg/kg-day, as shown in Table 5-2) corresponding to a greater than 10-fold range in human potency estimates for the same tumor site (see Table 5-6). 2) The metabolized dose approach assumes that the metabolic pathway leading to the production of the carcinogenic metabolite has been identified, but a cause-effect relationship has not been clearly established for any proposed pathway. 3) The metabolism of PCE in humans has not been clearly defined. Studies on metabolism have detected small amounts of two metabolites in urine. In humans, metabolism that produces urinary trichloro compounds accounts only for 1-2% of the administered dose following exposure to high concentrations, whereas 60-80% of the PCE inhaled is excreted unchanged through the lungs during the exposure. This implies that 20-40% of the absorbed PCE is stored (presumably in adipose tissue) and is slowly released over time. However, the high storage capacity of PCE in adipose tissue requires approximately 125 hours to reach steady-state equilibrium in adipose tissue upon continuous exposure. The slow release of PCE from adipose tissue results in a half-life in adipose tissue of approximately 91 hours. The difficulty in tracing the fate of these residual amounts of PCE has prevented the construction of a mass balance relationship between PCE absorbed and the PCE metabolized or exhaled unmetabolized by the lungs from human experiments. The residual PCE could be metabolized by the trichloro pathway over time, or by an unrecognized pathway producing metabolites such as C02 and oxalic acid that could not be easily recognized as products of metabolism without using radiolabeled compounds. SL 039033 5-49 Uncertainty remains DRAFT concerning the extent to which PCE is metabolized in animals and humans. Quantitative mass-balance studies of PCE metabolism during chronic (radiolabeled) PCE exposure conditions are available only for mice via ip-injection and drinking water exposure routes, while quantitative studies of urinary PCE-metabolite yield in chronically exposed mice are also available for the gavage exposure route. Such studies are not available for mice using an inhalation exposure route or for any exposure route using rats. This information necessitates the use of approximations <>t metabolized dose based on acute PCE exposures to interpret the results of chronic animal bioassays using inhalation exposures. Ln the context of extrapolating animal bioassay results to human risk predictions, further uncertainty arises from the current lack of any metabolic data on humans exposed to PCE under we 11-controlled, long-term experimental conditions te.g., ontinuous exposures lasting more than 1 week) -- a data gap t.at ta likely to persist. Even the interpretation of available data on f"'F. metabolism in experimentally exposed humans is uncertain, since these studies measured the production of only certain metabolities , rathi l ?han the total amount of metabolites produced. Data on the amount of PCE metabolized at ambient concentrations (Less than 1 ppb) are not available. However, several studies indicate that PCE metabolism increases as the concentration decreases. Extrapolation ot metabolism rates to ambient levels indicates that up to 100% of the TCE mav he metabolized by mice and rats. Humans may metabolize more than 601 at ambient levels compared to the level estimated in the human studies. Pharmacokinetic models generally assume chat PCE is metabolized at the same rate at different concentrations unless an ambient absorption factor is SL 039034 5-50 DRAFT added. 5) Pharmacokinetic models do not account for individual differences in metabolism and storage. A high variability of body burden of PCE was found for different people tested (Guberan and Fernandez, 1974; Hake and Stewart, 1977; Stewart et al., 1970). The body burden was found to depend on such factors as age, sex, exercise or workload, body mass and adipose tissue mass, pulmonary dysfunctional states, and individual differences in the intrinsic capacity to metabolize PCE. The pathways of PCE metabolism are speculative at present, but it is known that cytochrome P-450 is involved. The presence and basal activity of the enzyme class of cytochrome P-450 are determined genetically, but such systems are also readily altered by components of the diet. Certain phenolic food additives such as BHA and BHT, for example, influence enzyme levels involved in the detoxication and conjugation as well as the activation in animals. Epidemiologic evidence relating enzyme induction to components of human diet are not available at this time, but such induction is likely. Thus, although there is no definitive evidence, a large variability in metabolism in humans appears likely, but cannot be accounted for in the calculation of the metabolized dose other than through the incorporation of generic safety factors. None of the available pharmacokinetic models has incorporated a safety factor to account for variability in humans due to dietary or genetic predisposition. 6) Chronic studies in mice and occupational studies in humans indicate that mice and humans have similar sensitivity to PCE, so the 50-fold species difference suggested by some of the metabolized dose calculations do not appear justified. Thus, it is premature to use the metabolized dose in current estimations of human risk to PCE. SL 039035 5-51 For the low-dose risk assessment, the multistage model was used to estimate cancer potency. A time dependent analysis was found to result in values slightly larger. The surface area correction factor was used to adjust animal exposure data to a human exposure scenario. Thus as shown in Table 5-6, the suggested carcinogenic risk potency range is 0.016 to 0.073 (mg/kg per day). Using the assumption of a 70 kg individual breathing 20 m^/day, 3 and the conversion factor of 1 ppm - 6.89 mg/m , potency can be converted as follows: 0.016 (mg/kg-day) ^ x 6.89 ((mg/m^)/ppm) x 20 (m^/day) 1 " kg - 0.031/ppm - 31 x 10 Vppb and 0.073 (mg/kg-day) -1 x 6.89 ((mg/m3 /ppm) x 20 (ro3 /day) + 1- kg - 0.144/ppm - 144 x 10 ^/ppb. Based on the data of both rats and mice from the 1986 NTP si riv. it . estimated that the uDoer 95% confidence interval ran^ of risks indicai-us that 31 to 144 excess cancer cases mav result in everv million individuals 3 exposed over their lifetime to 1 rmb (6.78 ue/m ) PCE. Based on r nr... data. the unit risk for a lifetime continuous exDOSure to 1 ue/ra^ of PCK is estimated to be between 5 and 21 x 10 -6 The calculations represent the upper range of plausible excess cancer risk: the actual risk, which cannot be estimated, may be insiginificant. The potency based on a metabolized dose method suggested by EPA (1986) for the upperbound risk due to lifetime exposure to 1 pg/m3 of PCE in air ranges from 2,9 x 10 -7to 9.5-x710 However, EPA reported that the complete range of potencies for all methods 3 and dose-tumor incidence data evaluated for 1 pg/m of PCE in air was 2.9 x 10 ^ to 1,1 x 10 ^ (EPA 1986). Based on the CDHS potency evaluation of the a. 5-52 DRAFT annual average of 0.43 ppb of PCE in Che South Coast Air Basin and assuming that there are approximately 10 million residents in that area, an excess number of additional lifetime cancer cases of 133 to 619 might result from PCE exposure. Based on these findings, CDHS concludes that at ambient concentrations. PCE mav cause or contribute to an increase in mortality or serious illness due to the induction of cancer and thus pose a hazard to human health. SL 39037 5-53 DRAFT APPENDIX A DOSE-RESPONSE INFORMATION FOR ACUTE, SUBCHRONIC, AND CHRONIC TOXICITY IN ANIMALS (EXCLUDING TERATOGENIC, MUTAGENIC, AND CARCINOGENIC EFFECTS) Animal dose - response data should be evaluated together with applicable human data to establish safe exposure limits for PCE to prevent acute, subchronic and certain (specifically, noncarcinogenic and nonmutagenic) chronic toxicological endpoints. Here, dose-response data are reviewed according to exposure period (e.g., acute, subchronic, and chronic), route of exposure, and species for different toxic effects to major body organs and systems in animals. Acute Exposures Most of the studies reviewed did not clearly identify a no-observed-adverse effect level (NOAEL) for acute exposures. At the upper bound of acute, oral toxicity, LD50 values range from 3005 mg/kg to over 10,000 mg/kg (rats). Representative LD50 values are presented in Table A-l. Other acute responses to oral and inhalation exposures of PCE are shown in Tables A-2 and A-3, respectively. Even . at doses as low as 10 mg/kg, ventricular arrhythmias were observed in rabbits, and cardiovascular effects were noted in dogs at an average dose of 13 mg/kg. Table A-4 summarizes the acute responses to intravenous (IV) and intraperitoneal (IP) administrations. It is difficult to evaluate these studies with regard to their implications for predicted human toxicity because neither the doses nor the routes of administration are typical of human exposure. SL 039038 A-l Subchronic and Chronic Exposures DRAFT Carpenter and Rowe and co-workers conducted studies that identified subchronic and chronic exposure levels in the rat and guinea pig at which no effects were observed (Carpenter, 1937; Rowe et al., 1952). Carpenter exposed rats to 70 ppm of PCE 8 hours/day, 5 days/week for 7 months (Carpenter, 1937). Blood samples were analyzed for glucose and calcium. Periodic counts of white cells, polymorphonuclear neutrophil leukocytes, lymphocytes, and eosinophils were made. Urine was evaluated for bilirubin and albumin. In all instances, values were within normal limits. At the end of the exposure period, animals were sacrificed and examined for histological and/or pathological changes. No damage to organs or to the peripheral and central nervous systems were observed. Rowe and associates exposed female guinea pigs to 100 ppm of PCE 7 hours/day for a total of 13 exposures in 17 days (Rowe et al., 1952). Animals were observed for changes in behavior, appearance, growth, mortality, and final body weight. Tissues were also examined microscopically. No evidence of adverse effects was reported. However, a separate group of guinea pigs that were administered PCE by the same dose regime 132 times over 185 days displayed some evidence of toxicity. Females had a significant increase in liver weight (p - 0.01), and animals of both sexes had some abnormal deposition of fat in the liver. A recently completed bioassay of PCE used 100 ppm as the lowest concentration administered to mice (6 hours/day, 5 days/week for 103 weeks) (NTP, 1986). Although th results hve been discussed previously (see 03939 sv A-2 DRAFT Section A), it is important to note that numerous adverse effects were observed. Therefore, it appears that 70 ppm may be a NOAEL for inhalation exposure of rats (based on the 1937 Carpenter study), but NOAELS for mice and guinea pigs have not been identified. Hayes and co-workers studied the subchronic toxicity of PCE administered in the drinking water of rats (Hayes et al., 1986). Animals received 14, 400, or 1400 mg/kg-d in water daily, for 90 days. The primary effect was a significant decrease in body weight of high-dose males and of females given 400 to 1400 mg/kg-d (p<0.05). Although a loss of body fluids was not considered to be a factor in the lowered body weights, the actual cause of weight loss was not determined. Blood and serum were analyzed for a number of parameters. The only consistent effect was a significant elevation (p<0.05) of 5'-nucleotidase in high-dose animals of both sexes, as well as in males that received 400 mg/kg-,d. The authors suggest that this may be indicative of cholestasis (suppression of bile flow). However, no other serum indicators of hepatotoxicity were observed. A dose-dependent increase in liver and kidney/body weight ratios was found, but no statistical difference in liver and kidney/brain weight ratios occurred. No adverse hematological or pathological effects were noted in low-dose animals (14 mS/kg-d) The observed NOAEL in rats for ingestion of PCE in drinking water is 14 mg/kg. Tables A-5, A-6, A-7, and A-8 summarize the responses to subchronic and chronic administration of PCE. SL 039040 A- 3 draft TABLE A-1. MEAN LETHAL DOSES (LD50) OF TETRACHLOROETHYLENE TO LABORATORY ANIMALS Route Species Dose (mg/kg) Reference IPa IP IP Oral IP Oral Oral Oral Oral Oral Rat (female) Dog Rat (male) Rat Mouse Mouse Mouse Mouse Rat Rat 3005 3400 3835 3980 to 4680 4700 6400 to 8000 8115 8571 8850 13,000 Hayes et al., 1986 Klaassen and Plaa, 1967 Hayes et al., 1986 Withey and Hall, 1975 Klaassen and Plaa, 1966 Von Oettingen, 1964 Wenzel and Gibson, 1951 Dybing and Dybing, 1946 Lewis and Sweet, 1984 Smyth et al., 1969 Intraperitoneal administration. A-4 , DRAFT TABLE A-2. ACUTE ORAL TOXICITY OF TETRACHLOROETHYLENE TO ANIMALS Species Dose (mg/kg) Effect Reference Cat Cat Rabbit Rat Dog Cat Rabbit 812 No observed effects Maplestone and Chopra, 1933 1623 Drowsiness; unsteadiness Maplestone and Chopra, 1933 2158 50% Increase in serum lipoprotein; transient elevation in serum-enzyme activities Fujii, 1975 4700 Death 6492 to 8115 Death in 2 to 9 h Smyth et al., 1969 Lamson et al., 1929 6492 6492 8115 Death in 5 h Death in 36 h Death in 17 to 24 h Lamson et al., 1929 Lamson et al., 1929 Lamson et al., 1929 o390A2 A- 5 DRAFT TABLE A-3. ACUTE INHALATION TOXICITY OF TETRACHLOROETHYLENE TO ANIMALS Concentration Exposure Species (ppm) period Effect Reference Mouse Mouse Mouse Mouse Rat Rat Mouse Rat Mouse Mouse Rat Dog 200 400 800 1600 2000 2300 2917 3000 3700 3700 5163 9900 4 h Moderate fatty infiltration Kylin et al., 1963 of liver 4h Moderate to massive fatty Kylin et al., 1963 infiltration of liver Decrease in hepatic ATP levels; increase in total hepatic lipids and triglycerides Ogata et al., 1968 4 h Massive fatty infiltration Kylin et al., 1963 of liver; statistically significant increase in liver fat content Various CNS depression; loss of exposure consciousness; possible durations cardiac failure; respiratory failure Rowe et al., 1952 4 h Ataxia; 80% loss of avoidance and escape response Goldberg et al., 1964 4 h 100% Mortality NTP, 1986 "Several Loss of consciousness hours" Rowe et al. , 1952 24 min Anesthesia Gehring, 1968 470 min Liver dysfunction Gehring, 1968 4 h 100% Mortality NTP, 1986 Narcosis Lamson et 1929 al. , SL 039043 DRAFT SL 039044 TABLE A-4. ACUTE INTRAVENOUS (IV) AND INTRAPERITONEAL (IP) TOXICITY OF TETRACHLOROETHYLENE TO ANIMALS Species Dog Rabbit Dog Rat Rat Dog Dog Mouse Dose 13 mg/kg (mean dose) Route of administration IV (animals were anesthetized with pentobarbital) 10 mg/kg (mean dose) IV (animals were anesthetized with urethane) 20 to 40 mg/kg IV 1.3 mL/kg IP 0.3 to 2.0 mL/kg 1200 mg/kg 2300mg/kg 3900 mg/kg IP IP IP IP Effect Depression of myocardium; increased vulnerability of ventricles to epinephrine-induced arrhythmias; ventricular arrhythmias Increased vulnerability of myocardium to tachycardia Significant depression rate of rise of left intraventricular pressure Increase in BDPF*1; increase in C1-, K+, decrease in protein in BDPF 20 to 440% Increase, SG0TC Increased SGPT^, 50% of animals Increased retention PSP , 50% of animals ED,. ` elevation of SGPT Reference Kobayashi et al,, 1982 Kobayashi et al., 1982 Kobayashi et al., 1982 Hamada and Peterson, 1977 Cornish et al., 1973 Klaassen and Plaa, 1967 Klaassen and Plaa, 1967 Gehring, 1968 TABLE A-4. (Continued) Species Mouse Dose (mg/kg) 2.5 mL/kg Route of administration IP Mouse 4700 mg/kg IP Effect Necrosis and swelling proximal convoluted tubule Enlargement of hepatocytes; cellular infiltration and vacuolation; slight hepatic necrosis; slight necrosis convoluted tubule Reference Plaa and Larson, 1965 Klaassen and Plaa, 1966 I fl oo ^IV = Intravenous injection; IP - intraperitonea} injection. BDPF: Bile duct-pancreatic flow. ^SGOT: Serum glutamic oxalacetic transaminase. SGPT: Serum glutamic pyruvate transaminase. PSP: Phenosulfonephthalein. DRAFT SL 039045 SL 039046 TABLE A-5. SUBCHRONIC ORAL TOXICITY OF TETRACHLOROETHYLENE TO ANIMALS Species Dose (mg/kg) Dose regime Effect Reference DRAFT A-9 Rat 14 Daily for 90 d No observed adverse effects Hayes et al., 1986 Mouse 100,250,1000, 1500 or 2000 5 d/wk for 6 wk Significant increase absolute liver weight ; significant Buben and 0'Flaherty, 1985 increase liver triglycerides Mouse 100,250,500 or 1000 Daily for 11 d Significant elevation in absolute liver weights ^ hepatocellular swelling ; significant decrease of hepatic DNA content per gram of liver ; dose-related increase of hepatic DNA synthesis (1000 mg/kg) Schumann et al., 1980 Rat 100,250,500 or 1000 Daily for 11 d Significant elevation of liver weight (1000 mg/kg group only); minimal hepatic changes (1000 mg/kg group only) Schumann et al,, 1980 Mouse 100,200,500,1000, 5 d/wk for Dose - dependent increase in 1500 or 2000 6 wk liver degeneration and karyorrhexis Buben and O'Flaherty, 1985 Dog 300 Degenerative changes of liver; Hall and Schillinger, extensive atrophy of liver 1925 TABLE A-5. (Continued) Species Dose (mg/kg) Dose regime $ Effect Reference l< / 6 0 Rat 400 House 500,1000,1500, or 2000 Daily for 90 d 5 d/wk for 6 vk Significant decrease in body weight (females); significant increase in 5'-nucleotidase (males) 3" b Significant decrease in G6P ^ significant increase in SGPT Hayes et al. ,, 1986 Buben and 0'Flaherty, 1985 House 1000 5 d/wk for 6 wk Significant decrease in hepatic DNA content (indicative of hypertrophy) Buben and 0'Flaherty, 1985 oi-v Rat 1400 Daily for 90 d Significant decrease in body weight (both sexes); significant increase in 5'-nucleotidase (both sexes) Hayes et al.., 1986 ^All dose levels. G6P: Glucose 6 phosphatase. SGPT: Serum glutamic pyruvate transaminase. DRAFT SL 039048 TABLE A-6. SUBCHRONIC INHALATION TOXICITY OF TETRACHLOROETHYLENE TO ANIMALS Species Concentration (ppm) Dose regime Exposure period Effect Reference Guinea pig (females) Guinea Pig Rat Rat Mouse Rat Rat 100 100 200 13 exposures in 17 d 132 exposures in 185 d . 4 exposures in 4 d 230 5 d/wk 21 exposures 200 to 1600 Daily for 3d 400 130 exposures in 183 d 800 Daily for 1 month 7 h/d 7 h/d 6 h/d 8 h/d 8 h/d 7 h/d 12 h/d No observed effects Rowe et al. , 1952 Significant increase in liver weight (female); fat deposition in liver Decreased RNA content of brain; increased nonspecific cholinesterase activity; behavioral changes Granular swelling and congestion of kidneys Inhibition of growth; increase in mortality No observed effects in mortality Significant decrease in ACh Rowe et al. , 1952 Savolainen al., 1977 o Carpenter, 1937 Schumacher et al., 1962 Rowe et al. , 1952 Horuna et a) 1980 Rat 1600 8 exposures in 10 d 7 h/d Slight degeneration of germinal epithelium of testes; increase in liver weights; moderate fatty degeneration of liver Rowe et al. , 1952 6*60 Co 3> H- ro Species Concentration (ppm) Dose regime TABLE A-6. (Continued) Exposure period Effect Reference Rabbit Guinea Pig Rabbic Rat 2211 2500 2500 2500 6 d/wk for 45 d 18 exposures in 24 d 28 exposures in 39 d 13 exposures in 18 d 4 h/d 7 h/d 7 h/d 7 h/d Significant decrease in rate of glomerular filtration; decrease of renal tubular capacity High mortality; loss of equilibrium, coordination, and strength; rapid weight loss; increase in liver and kidney weights; fatty degeneration of liver; swelling of tubular epithelium Slight hepatic degeneration; CNS depression to point of helplessness CNS depression with loss of consciousness; 90% lethality Brancaccio et al. 1971 Rowe et al. , 1952 Rowe et al. , 1952 Rowe et al. , 1952 ACh: acetylcholine. DRAFT 0S06E0 'IS TABLE A-7. CHRONIC ORAL TOXICITY OF TETRACHLOROETHYLENE TO ANIMALS (Noncarciaogenic effects only) Species a Mouse (female) Dose*1 (mg/kg) Dose regime 386C 5 d/wk for 78 wk Exposure frequency Daily Effect Reference Early mortality (50% mortality by wk 62); toxic nephropathy (96%) NCI, 1977 Rat (male) *71 5 d/wk for Daily Early mortality (50% mortality NCI, 1977 78 wk by wk 72); toxic nephropathy (88%) DRAFT Rat (female) *7* 5 d/wk, for 78 wk Daily Early mortality (50% mortality by wk 74); toxic nephropathy (58%) NCI, 1977 Mouse (male) 536 5 d/wk for Daily Early mortality (50% mortality NCI, 1977 78 wk by wk 78); toxic nephropathy (82%)) Mouse 772 5 d wk for Daily Early mortality (50% mortality NCI, 1977 (female) 78 wk by wk 50); toxic nephropathy (100%) Rat (male) R.ti ( t t'HWl 11- ) 941 949 5 d/wk for 78 wk Daily 5 d/wk for 78 wk Daily Early mortality (50% mortality by wk 44); toxic nephropathy (94%) NCI, 1977 Early mortality (50% mortality by wk 66); toxic nephropathy (76%) NCI, 1977 o tP v> O O' A -14 TABLE A-7. (Continued) Species Dose (mg/kg) Dose regime Exposure frequency Effect Reference Mouse (male) 1072*1 5 d/wk for 78 wk Daily Early mortality (50% mortality by wk 43); toxic nephropathy (94%) NCI, 1977 0 ^Each dose group initially had 50 animals. cAll values are time-weighted average doses. These dose levels are associated with an increased incidence of cancer in experimental animals. See Section 5 for additional information. SL 039052 TABLE A-8. CHRONIC INHALATION TOXICITY OF TETRACHLOROETHYLENE TO ANIMALS Species Rabbit Rat Mouse Concentration (ppm) Dose regime Exposure period 15 7 to 11 months 3 to 4 h/d 70 5 d/wk for 7 months 100 5 d/wk for 103 wk 8 h/d 6 h/d Mouse (females) Guinea Pig 100 5 d/wk for 103 wk 6 h/d 200 158 exposures 7 h/d in 220 d Mouse Rat Rat (males) 200 5 d/wk for 103 wk 6 h/d 230 5 d/wk for 8 h/d 7 months (150 exposures) 4.00 5 d/wk for 103 wk 6 h/d Effect Reference Reduction in agglutinin formation Mazza, 1972 No observed effects Carpenter, 1937 Significantly lower survival than controls (after wk 74); increased incidence of liver degeneration and liver necrosis Increased incidence of nephrosis NTP, 1986 NTP, 1986 o ~n Significant depression of growth; increase in liver Rowe et al. , 1952 and kidney weight; slight to moderate fatty degeneration of liver Significantly lower survival than controls (after wk 78); increased incidence of liver degeneration, liver necrosis, and nephrosis NTP, 1986 Granular swelling of the liver; Carpenter, 1937 decrease in glycogen storage Decreased survival (after wk 82) NTP, 1986 039053 Species Mouse Guinea Pig Rabbit Monkey Rat Rat Rat TABLE A-8. (Continued) Concentration (ppm) Dose regime Exposure period Effect Reference 400 400 169 exposures in 236 d 169 exposures in 236 d 7 h/d 7 h/d 400 400 470 159 exposures in 220 d 179 exposures in 250 d 5 d/wk for 7 months 150 exposures 7 h/d 7 h/d 8 h/d 600 5 d/wk 6 h/d for 12 months 600 5 d/wk for 12 months 6 h/d Increase in kidney weight; swelling of tubular epithelium Significant depression of growth; increase in liver and kidney weights; moderate fatty degeneration and slight cirrhosis of liver No observed effects Rowe et al. , 1952 Rowe et al. , 1952 Rowe et al. , 19521!~ No observed effects Rowe et al. , 1952r _ Liver congestion and cloudy swelling; increased secretion, cloudy swelling and desquamation of kidney; congestion and increased pigmentation spleen Carpenter, 1937 T| "s Increase in mortality Leong et al 1975 Increase in mortality (males only); inflammation of kidney cells; nephrosis Rampy et al 1978 DRAFT REFERENCES Abedin, Z., R. C. Cook, Jr., and R. M. 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