i 

## Technical Review of Acetonitrile (Methyl Cyanide)

CAS Registry Number 75
- 05
- 8

Office of Environmental Information Environmental Analysis Division Analytical Support Branch

January 17 , 201 3ii Purpose of this Hazard Assessment for Acetonitrile Before a chemical can be proposed for addition to or deletion from section 313 of the Emergency Planning and Community Right

- To

- Know Act (EPCRA) of 1986, EPA performs a hazard assessment on the chemical. The statutory chemical listing/delisting criteria of EPCRA section 313(d)(2) are generally based on human health and environmental hazards. The analysis focuses on identifying and characterizing a chemical’s potential for toxicity to hum an s and the environment. This technical report provides an overall assessment and review of the toxicity of acetonitrile, which includes a review and consideration of both human health and environmental hazards. It includes a n assessment of the severity of the effect(s) and the dose or concentration levels at which the effect(s) were observed. This report contains several chapters including chemistry and environmental fate, human health hazards, and ecological toxicity

. An “Executive Summary” is included to provide the user with a brief review of the key findings presented in the following chapters of the report

.iii 

### T ABLE OF 

### C ONTENTS

1. EXECUTIVE SUMMARY
................................ ................................ ................................ ....... 1

1.1 BACKGROUND
................................ ................................ ................................ .................... 1

1.2 CHEMISTRY AND E NVIRONMENTAL FATE AS SESSMENT
................................ ... 1

1.3 HUMAN HEALTH HA ZARD ASSESSMENT CONC LUSIONS
................................ .... 3

1.4 ECOLOGICAL TOXI CITY HAZARD ASSESSME NT CONCLUSIONS
..................... 4

1.5 REFERENCES
................................ ................................ ................................ ....................... 5

2. CHEMISTRY AND ENVIRO NMENTAL FATE
................................ ................................ 7

2.1 SUMMARY
................................ ................................ ................................ ............................ 7

2.2 SYNONYMS/IDENTIFIERS
................................ ................................ ................................ 9

2.3 PHYSICAL AND CHEMICA L PROPERTIES
................................ ................................ . 9

2.4 PRODUCTION METHODS
................................ ................................ ............................... 10

2.5 USE PROPERTIES
................................ ................................ ................................ ............. 11

2.6 ENVIRONMENTAL F ATE
................................ ................................ ............................... 12

2.7 REFERENCES
................................ ................................ ................................ ..................... 22

3. HUMAN HEALTH HAZARD ASSESSMENT
................................ ................................ .. 27

3.1 SUMMARY
................................ ................................ ................................ ........................... 27

3.2 KINETICS AND METABOL ISM
................................ ................................ ..................... 29

3.3 EFFECTS OF ACUTE EXPOSURE
................................ ................................ .................. 33

3.4 EFFECTS OF REPE ATED EXPOSURE
................................ ................................ ......... 34iv 3 .5 DEVELOPMENTAL AND RE PRODUCTIVE TOXICITY
................................ .......... 43

3.6 GENOTOXICITY
................................ ................................ ................................ ................ 49

3.7 NEUROTOXICITY
................................ ................................ ................................ ............. 50

3.8 POTENTIAL MODES OF A CTION
................................ ................................ ................. 51

3.9 THYROID TOXICITY
................................ ................................ ................................ ....... 52

3.10 HUMAN RELEVANCE OF A NIMAL DATA
................................ ............................... 53

3.11 CONCLUSION
................................ ................................ ................................ .................. 54

3.12 REFERENCES
................................ ................................ ................................ ................... 58

4. ECOLOGICAL HAZARD ASSESSMENT
................................ ................................ ......... 65
4.1. SUMMARY
................................ ................................ ................................ .......................... 65

4.2 ECOLOGICAL TOXICITY
................................ ................................ ............................... 65

4.3. REFERENCES
................................ ................................ ................................ .................... 751 CHAPTER ONE: EXECUTIVE SUMMARY
1.1 BACKGROUND On June 28, 2002, BP Chemicals Inc. (BP) petitioned the Agency to remove acetonitrile (methyl cyanide) from the list of chemicals subject to the reporting requirements of Section 313 of the Emergency Planning and Comm unity Right

- to

- Know Act (EPCRA) of 1986. The 2002 petition followed the EPA denial of an earlier petition submitted by BP for the removal of acetonitrile from the EPCRA Section 313 reporting requirements. Currently, the INEOS Group has assumed the aceton itrile petition. In the 2002 petition, BP claimed that acetonitrile does not satisfy the criteria for adverse acute or chronic health effects or adverse environmental effects. The critical issue of the 2002 petition is the assertion that the mortality re ported in the N ational Toxicology Program (N TP ) (1996) 13

- week study was an acute effect. EPA’s technical assessment of the mortality reported in the NTP study along with a review of other human health and environmental hazards and the environmental fate of acetonitrile are provided in this report.
1.2 CHEMISTRY AND ENVIRO NMENTAL FATE ASSESSMENT

Acetonitrile (CAS Registry No. 75

- 05

- 8), also known as methyl cyanide, is a colorless,

v olatile , flammable organic liquid ( Budavari, et. al., 1996 )

. Its molecular formula is C 2 H 3 N , molecular weight, 41.05 ( structure is shown above)

. Acetonitrile is produced as a by

- product of acrylonitrile production (McKinney and DeVito, 1998) and is used in an array of industrial processes as a solvent, chemical intermediate and catalyst (McKinney and DeVito, 1998)

. The production and wide use of acetonitrile may result in its release to the environment through various waste streams. In 2006 , the aggregated U.S. pr oductions levels reported to EPA were between 50 million to less than 100 million pounds per year (U.S. EPA, 20 11 a ). Of the millions of pounds of acetonitrile produced each year, a commercial market exists for approximately half of the inventory (U.S. EPA , 2011 ; U.S. EPA, 20 11 a )

. In 2009, releases of acetonitrile reported to the Toxic Release Inventory (TRI) totaled 17.7 million pounds, of which 17.4 million pounds2 were disposed to on

- site in underground injection wells (U.S. EPA, 2011) . An additional 17.1 million pounds were burned on

- site for energy recovery by the manufacturers

. Ac e tonitrile is predicted to volatilize into the atmosphere from soil or water surfaces and will primarily exist as a vapor in the atmosphere ( Bidleman, 1988)

. Model predi ctions using an emission scenario based on industrial release data reported to the TRI estimate the acetonitrile distribution to air, water, and soil , with no advection taken into consideration , to be 79. 8 , 1 16.7 and 3.5 percent, respectively (U.S. EPA, 20 1 1b ). In the atmosphere, wet deposition can be the most significant loss process in regions of frequent rainfall resulting in deposition of acetonitrile to soil and surface water (Hamm et al., 1984; Karl et al., 2004; Sanhueza et al., 200 4 ). Based upon i ts low soil sorption coefficient and high water solubility, it is predicted that acetonitrile will migrate rapidly through permeable soils with the potential for groundwater contamination. O nce in contact with underground water, acetonitrile will readily disperse throughout the water body. The most significant degradation process for acetonitrile is biodegradation ; n either hydrolysis nor photodegradation is expected to be an important degradation process for acetonitrile (Brown et al., 1975, Atkinson, 1994). Observed bio degradation rates of ac e ton itrile have been reported at 58% in 5 days ( Babeu and Vaishnav , 1987 ), and 40% after 10 days (Urano and Kato , 1986 a,b). The potential for anaerobic biodegradation has been indicated by studies conducted by Li et al., (2007) although studies were not conducted under natural environmental conditions but conditions more pertinent to sewage treatment plants. Based upon the information presented in section s 2.6.3.2 and 2.6.3.3 of this report, it can be anticipated that the actual biodegradation of acetonitrile in underground injection wells where high concentrations of the chemical are present and anaerobic or oxygen limiting conditions are likely to exist , will be considerably slower than observed biodegradation rates under aerobic conditions for relatively low concentrations of the chemical.3 1. 3 HUMAN HEALTH HAZARD ASSESSMENT CONCLUSION S
In review of the potential human toxicity of acetonitrile , b ased on the available data, and given the severity of the effect, mortality, EPA concludes that there is sufficient evidence to support a concern for moderately high toxicity from long

- term repeat ed exposure. The NTP (1996) 13

- week subchronic inhalation study was identified as the principal study for deriving the Reference Concentration (RfC) in the Toxicological Review of acetonitrile (U

. S

. EPA, 1999). The use of mortality as the critical effect for derivation of the chronic Reference Concentration (RfC) for the Integrated Risk Information System (IRIS) is due to the steep dose

- response curve for acetonitrile, which is consistent with the exposure

- response data for other cyanide

- containing chemica ls. In addition, the Toxicological Review identified the 400 parts per million ( ppm ) exposure level in the NTP (1996) mouse study as a Frank Effect Level (FEL) given the death of a mouse at week 2 at 400 ppm and the increased mortality at 800 ppm. In the NTP (1996) 13

- week mouse inhalation study, mortality was observed in the 800 and 1 , 600 ppm ( ~ 402 and 805 milligrams per kilogram per day ( mg/kg

- day ) ) treatment groups. The first incidence of mortality in the 800 ppm t reatment group was not observe d until day 20 and single deaths continued on days 21, 45, 69 and 89 of the 13

- week study. This pattern of mortality is dissimilar to that observed in the 1 , 600 ppm treatment group of the NTP (1996) 13

- week mouse inhalation study, where initial deaths wer e observed in the first week and all mice died by day 21. Based on the observed pattern of death in the 800 ppm treatment group of the NTP (1996) 13

- week inhalation study, beginning at the end of the third week and extending through the termination of the study, it can be reasonably anticipated that additional acetonitrile

induced mortality would have continued beyond the termination of the study and the sacrifice of surviving animals. Because the mortalities extended from the third week of the study to s tudy termination, the data indicates that the mortality observed in the 800 ppm treatment groups is not due to a single acute exposure to sufficiently high acetonitrile concentrations, but rather is best explained as being the result of long

- term repeated exposures. The observed exposure
- response relationship for acetonitrile demonstrates that a threshold exists at acetonitrile exposures which are sufficient to cause mortality

.4 In addition to the concern for increased morta lity following repeated dose exposures, immunotoxicity and developmental toxicity are also concerns. Evidence of immunotoxicity was observed in mice following a short

- term exposure to 100 mg/kg

- day, but was not observed in rats following subchronic and ch ronic exposures. Multiple studies demonstrated toxicity in the developing organism at oral doses ranging from approximately 30 mg/kg

- day in rabbits and 275 mg/kg

- day in rats and inhalation concentrations of 1 , 8 27 ppm in rats . It should be noted that these doses were also maternally toxic; however, the observation of maternal toxicity does not abrogate all concern with the observations of toxicity in the developi ng organism (U.S. EPA, 1991)

. 1. 4 ECOLOGICAL TOXICITY HAZARD ASSESSMENT CONCLUSION S Based upon a review of the limited acute and chronic toxicity data available for aquatic species (algae, invertebrates, and vertebrates) , adverse environmental effects d ue to acetonitrile can only be anticipated at relatively high exposure concentrations. This conclusion is based on observed acute toxicity values for aquatic organisms that are greater than or equal to 100 milligrams per liter ( mg/L ) , indicating relatively low acute aquatic toxicity (Table 4

- 2) . Acute toxicity values for fr eshwater fish and aquatic invertebrates ranged from 100 to 3 , 600 mg/L. The 96 hour LC 50 values for fish and green algae ranged from 100 to 3454 mg/L (Table 4

- 2)

. E xperimental chronic toxicity values for green and blue

- green algae, freshwater invertebra tes, freshwater fish embryos, and duckweed ranged from 160 to 7 , 300 mg/L indicating relatively low chronic toxicity. Experimental chronic toxicity studies on earthworms reported 100% survival at exposure concentrations of 500 micrograms per gram ( μ g/g ) soil and 200 microliters per gram ( μ L /g ) soil. No data were available to assess chronic toxicity to estuarine/marine fish or invertebrates. Acetonitrile is not expected to bioaccumulate in aquatic organisms

. Available studies on the toxicity of acetoni trile to terrestrial species are limited to experimental chronic toxicity values for earthworms. These studies reported 100% survival at exposure concentrations of 500 μ g/g soil for 14 days and 200 μ L/g dry soil for 15 days (Phillips et al., 1993 and Robi doux et al., 2000, respectively). The phytotoxicity of acetonitrile has been5 assessed experimentally in duckweed. Chronic toxicity values for duckweed ranged from 368 to 1 , 342 mg/L (Tong and Hongjun, 1997; 2001), indicating relatively low chronic toxicit y. No data on population effects were located.
1. 5 REFERENCES
Atkinson, R. 1994. Gas

- phase tropospheric chemistry of organic compounds. J. Phys. Chem. Ref. Data. Monograph. 2:2162]. Babeu, L. and D.D. Vaishnav. 1987. Prediction of biodegradability for selected organic chemicals. J. Ind. Microbiol. 2:107

- 115. Bidleman, T.F. 1988. Atmospheric processes. Wet and dry deposition of organic compounds are controlled by their vapor

- particle partitioning. Environ . Sci. Technol. 22:361

- 367. Brown, S.L., F.Y. Chan, J. Jones et al. 1975. Research program on hazard priority ranking of manufactured chemicals (chemicals 41

- 60). Phase II

- final report. Menlo Park, California, SRI International, April 1975. p. 85. N TIS PB

- 263163. Budavari, S., M.J. O’Neil, A. Smith et al. (eds). 1996. The Merck Index, 12th edition. Whitehouse NJ, Merck & Co., Inc., p13. Hamm, S., J. Hahn, G. Helas et al. 1984. Acetonitrile in the troposphere: Residence time due to rainout and up take by the ocean. Geophys. Res. Let. 11:1207

- 1210. Karl, T., M. Potosnak, and A. Guenther. 2004. Exchange process of volatile organic compounds above a tropical rain forest: Implications for modeling tropospheric chemistry above dense vegetation. J. Geophys. Res. 109:D18306. Li, T., J. Liu, R. Bai et al. 2007. Biodegradation of organonitriles by adapted activated sludge consortium with acetonitrile

- degrading microorganisms. Water Res. 41:3465

- 3473. McKinney, R.J., and S.C. DeVito. 1998. Nitrile s. In: Kirk

- Othmer Encyclopedia of Chemical Technology. Electronic version available through subscription to Wiley Interscience. John Wiley & Sons. Available at http://www.mrw. interscience.wiley.com/kirk/kirk_search_fs.html

. NTP (National Toxicology Program). 1996. Toxicology and carcinogenesis studies of acetonitrile (CAS NO. 75

- 05

- 8) in F344/N rats and B6C3F1 mice (inhalation studies). NTP TR 447. Phillips, C.T., R.T. Chec kai and R.S. Wentsel. 1993. Toxicity of selected munitions and munition

- contaminated soil on the earthworm (Eisenia foetida). Edgewood Research, Development & Engineering Center, Aberdeen Proving Ground, MD: 22. NTIS ADA264408.6 Robidoux, P.Y., C. Svendse n, J. Caumartin et al. 2000. Chronic toxicity of energetic compounds in soil determination using the earthworm ( Eisenia andrei ). Environ. Toxicol. Chem. 19(7):1765

- 1773. Sanhueza, E., R. Holzinger, B. Kleiss et al. 2003. New insights in the global cycle of acetonitrile: Release from the ocean and dry deposition in the tropical savanna of Venezuela. Atmos. Chem. Phys. Discuss. 3:5275

- 5288. Tong, Z. and J. Hongjun. 1997. Use of duckweed ( Lemna minor L. ) g rowth inhibition test to evaluate the toxicity of acrylonitrile, sulphocyanic sodium and acetonitrile in China. Environ. Pollut. 98(2):143

- 147. Tong, Z. and J. Hongjun. 2001. Aquatic toxicity of acetonitrile and its water quality criteria for the prote ction of aquatic life in China. Int. J. Environ. Pollut. 15(5):568

- 575. Urano, K, and Z. Kato. 1986a. A method to classify biodegradabilities of organic compounds. J. Hazard. Mat. 13:135

- 145. Urano, K. and Z. Kato. 1986b. Evaluation of biodegradatio n ranks of priority organic compounds. J. Hazard. Mat. 13:147

- 159. U.S. EPA. 1991. Guidelines for Developmental Toxicity Risk Assessment. Risk Assessment Forum, Washington, DC. EPA/600/FR

- 91/001. U.S. EPA. 1999. Toxicological Review of Acetonitrile . Office of Research and Development. Washington, DC. January, 1999. Available at http://www.epa.gov/iris/toxreviews/0205

- tr.pdf

. U.S.EPA. 20 11

. TRI.NET program. Generated values based upon 2009 reported TRI data. Report generated May 6, 2011. TRI.NET is available at EPA.gov.tri/tridotnet/
U.S.EPA 20 11 a . Inventory Update Rule (IU) 200 6 . Toxic Substances Control Act (TSCA) Inventory Update Database. Environmental Protection Agency (EPA). Information retrieved May 9, 2011. Available at http://www.epa.gov/oppt/iur/iur02/index.htm

. U.S. EPA 2011b. EPI SuiteTM v.4.0. © 2000

- 2009 U.S. Environmental Protection Agency. Report generated May 31, 2011. Available at http://www.epa.gov/opptintr/exposure/pubs/episuitedl.htm.

.7 CHAPTER TWO: CHEMISTRY AND ENVIRONMENTAL FATE

2.1 SUMMARY Acetonitrile (CAS Registry No. 75

- 05

- 8) is a member of the nitriles chemical class. The molecular formula is C 2 H 3 N and molecular weight is 41.05. It is a colorless, volatile, and flammable liquid with an ether

- like odor and is completely miscible with water and many organic solvents, but is immiscible with many saturated hydrocarbons (Budavari, et. al., 1996; WHO, 1993). Its high dielectric strength and dipole moment make it an excellent solvent for both inorganic and organic compounds including polymers (McKinney and DeVito, 1998). Acetonitrile is produced as a by

- product of acrylonitrile production. The principal uses of acetonitrile include: as a solvent in hydrocarbon extraction processes, as a specialty solvent, as a reaction solvent in the production of synthetic pharmaceutical s, as a solvent for the extraction of fatty acids, and as a solvent in analytical instrumentation (Lewis 2001; Budavari et al., 1996; WHO 1993). In 2006 the aggregated U.S. productions levels reported to EPA were between 50 to less than 100 million pounds per year (U.S. EPA, 20 11 a ). Of the millions of pounds of acetonitrile produced each year, a commercial market exists for about h alf of the inventory

. In 200 9 , releases of acetonitrile reported to the TRI totaled 17.7 million pounds, of which 17.4 millio n pounds were disposed to on

- site in underground injection wells (U.S. EPA 2011) . An additional 17.1 million pounds were burned on

- site for energy recovery by the manufacturers.
The environmental fate assessment of acetonitrile indicates that most of the chemical will volatilize into the atmosphere from terrestrial and aquatic surfaces. Based upon a Level III Fugacity Model (see section 2.6.4), air is predicted to be the environmental compartment where the largest mass of acetonitrile ( 80 percent ) wil l partition (U.S. EPA 2011b) . This is due in part to the slow abiotic degradation (photodegradation) of acetonitrile in the atmosphere. The predicted half

- life values for photodegradation are from months to over a year, depending upon the concentrations of hydroxyl radicals in the atmosphere as well as other environmental factors (HSDB, 2002; Atkinson, 1994). In the atmosphere acetonitrile is subject to wet deposition.8 Based upon a soil partitioning coefficient (K oc ) value of 5, acetonitrile may demonst rate a high mobility in soils and the potential for groundwater contamination. The most significant degradation process for acetonitrile is biodegradation. The observed degradation rate of acetonitrile at an initial concentration of 100 ppm in water by an acclimated mixed microbial culture was reported at 58% in 5 days (Babeu and Vaishnav, 1987) . Urano and Kato (1986b) using an activated sludge microbial ino culum in an aqueous media with an initial 100 ppm concentration of acetonitrile reported 40% degradation after10 days ( Urano and Kato (1986 a, b ). The potential for anaerobic biodegradation , has been indicated by studies conducted by Li et al., (2007) altho ugh studies were not conducted under relevant environmental conditions but conditions relevant to sewage treatment plants. The anaerobic studies by Li et al. demonstrated the total degradation of acetonitrile by acclimated bacteria to occur at 20 hours (0 .8 days) for concentrations of acetonitrile of 1 gram per liter ( g/L )

. Parallel studies conducted by Li et al. under aerobic conditions resulted in total degradation at 10 hours (Li et al, (2007)). However, b ased upon the additional information presented in section 2.6.3.2 and 2.6.3.3 of this report, it can be anticipated that the actual biodegradation of acetonitrile in underground injection wells where high concentrations of the chemical and anaerobic or oxygen limiting conditions are likely to exist , will be considerably slower than observed biodegradation rates under aerobic conditions for relatively low concentrations of the chemical. The abiotic degradation of acetonitrile is expected to be very slow. Neither hydrolysis nor p hotodegradation is expected to be an important degradation process for acetonitrile (Brown et al., 1975, Atkinson, 1994). Reactions of acetonitrile with hydroxyl radicals in the atmosphere are slow, with estimated half

- life values of months to over a year , depending on the concentrations of hydroxyl radicals in the atmosphere and other factors (Atkinson, 1994). The half

- life for the base

- catalyzed hydrolysis of acetonitrile at pH 7 has been predicted at 150,000 years (HSDB, 2002)

.9 2.2 SYNONYMS/IDENTIFIERS (CAS, 2002 ; ChemIDplus, 2011 ):
Acetonitrile Cyanomethane
Ethanenitrile Ethyl nitrile Methane, cyano

Methanecarbonitrile Methyl cyanide (MeCN)
NCI

- C60822
NSC 7593
2.3 PHYSICAL AND CHEMICA L PROPERTIES
Table 2

- 1, Physical and Chemical Properties of Acetonitrile (CAS Reg. No. 75

- 05

- 8)

Property Value a Conditions Reference
Molecular Formulae C 2 H 3 N Molecular Weight 41.05 Freezing point (°C)

- 45.7

- McKinney and DeVito, 1998 Boiling point (°C) 81.6

- McKinney and DeVito, 1998 Vapor pressure (mm Hg) 88.8 25 °C Boublik et al., 1984 Water solubility (mg/L) 1x10 6 25 °C Riddick et al., 1986 Log K ow - 0.34

- Hansch et al., 199510 Henry’s Law constant (atm

m 3 /mol) 3.45x10 - 5 25 °C Gaffney et al., 1987

Dissociation constant, pKa

- 4.30

- Perrin, 1965 Hydroxyl radical reaction rate constant (cm 3 /molec

sec) 2.63x10 - 14 25 °C Atkinson, 1994

Density (g/cm 3 ) 0.786 20 °C McKinney and DeVito, 1998 Vapor density 1.42 Air = 1 Clayton and Clayton, 1982
K oc 5 (HSDB, 2002; Lyman et al., 1990)

a Values are experimental unless otherwise indicated.
2.4 PRODUCTION METHODS Acetonitrile is produced commercially as a by

- product from the manufacture of acrylonitrile by propylene ammoxidation (McKinney and DeVito , 1998). In this process, known as the Sohio Process, the gas

- phase reaction of propylene with ammonia in the presence of ammonia (ammoxidation) and an oxidation catalyst (based on, for example, vanadium or molybdenum) produces acrylonitrile :

A cetonitrile and hydrogen cyanide are the principal by

- products of th is reaction

. Approximately 30

- 40 kilograms ( kg ) of acetonitrile and 140

- 180 kg of hydrogen cyanide are generated per 1000 k g of acrylonitrile produced (Pollak, Romeder, et. al., 1991). The acetonitrile is recovered as the water azeotrope, dried, and purified by distillation. The impurities present in commercial

- grade acetonitrile are water, copper, iron, and propionitrile; small amounts of allyl alcohol may also be present (McKinney an d DeVito, 1998). Some acrylonitrile producers H 2 C C H C H 3 + N H 3 3/2 O 2 + catalyst H 2 C C H C N H 2 O 3 +
byproducts: H 3 C C N H C N + propylene acrylonitrile

acetonitrile11 recover and purify acetonitrile, but most burn the by

- products as plant fuel and/or dispose of it into on

- site underground injection wells (Pollak, Romeder, et. al., 1991; U.S. EPA, 2 01 1 )

. In 2006 the aggrega ted U.S. productions levels reported to EPA were between 50 to less than 100 million pounds per year (U.S. EPA, 20 11 a ). Of the millions of pounds of acetonitrile produced each year, a commercial market exists for about h alf of the inventory

. In 2009, releases of acetonitrile reported to the T RI totaled 17.7 million pounds, of which 17.4 million pounds were disposed to on

- site in underground injection wells. An additional 17.1 million pounds were burned on

- site for energy recovery by the manufacturers (U.S. EPA, 2011)

. Acetonitrile can also be synthesized by several other routes: (1) dehydration of an acetic acid and ammonia mixture, acetamide, or ammonium acetate; (2) reaction of ammonia and acetylene; (3) reaction of ethanol and ammonia in the prese nce of a catalyst; and (4) reaction of cyanogen chloride with methane, ketones, ethanol, alkylene epoxides, and paraffins or olefins (Budavari, et. al., 1996; WHO, 1993).
Acetonitrile is also a naturally occurring substance . Small quantities are present in coal tar (Budavari, et. al., 1996). It is also formed during the burning of biomass and released to the atmosphere (Karl et al., 2004; Kondo et al., 2004; Sanhueza et al., 200 4 ).
2.5 USE PROPERTIES Because of its good solvency and relatively low boiling point, acetonitrile is used widely as a recoverable reaction medium, particularly for the preparation of pharmaceuticals. Its largest use is for the separation of butadiene from C 4 hydrocarbons by extractive distillation. It also has been propose d for the separation of other olefins (e.g., propylene, isoprene, allene, and methylacetylene from hydrocarbon streams). It is a superior solvent for polymers and can be used as a solvent for spinning fibers and for casting and molding plastics. It is us ed widely in spectrophotometry and electrochemistry. Since pure acetonitrile does not absorb UV light, it is commonly used as a solvent in high pressure liquid chromatography (HPLC) (McKinney and DeVito, 1998). It is also used as a solvent in the extract ion of fatty acids from vegetable, fish liver, and animal oils (Budavari, et. al., 1996).12 Acetonitrile is also used as a catalyst and as an ingredient in transition

- metal complex catalysts. There are many uses for it in the photographic industry and for the extraction and refining of copper and by

- product ammonium sulfate. It also is used for dyeing textiles and in coating compositions. It is an effective stabilizer for chlorinated solvents and it has some application in the manufacture of perfumes (McK inney and DeVito, 1998). Another use of acetonitrile is as a reagent for the preparation of a wide variety of compounds. It undergoes typical nitrile reactions to produce many types of nitrogen

- containing compounds such as amides, amines, higher molecula r weight mono

- and dinitriles, halogenated nitriles, ketones, isocyanates, pyridines, and imidazolines (McKinney and DeVito, 1998). The major use of acetonitrile in organic synthesis is in the production of malononitrile through reaction with cyanogen chl oride (Pollak, Romeder, et. al., 1991). The production and wide use of acetonitrile as a solvent, chemical intermediate, and catalyst may result in its release to the environment through various waste streams. It is also released to the atmosphere from incineration of polyacrylonitrile polymers, automobile exhaust, tobacco smoke, synthetic rubber manufacture, off

- gas of shale oil retorting, and manufacture of acrylonitrile and turbine engines. Acetonitrile has been detected in shale oil wastewater and in wastewater from coal gasification processes (HSDB, 2002).
2.6 ENVIRONMENTAL FATE 2.6.1. Environmental Transport
. When release d into the environment, acetonitrile is expected to volatilize from the surface of moist soil and water surfaces

. Based on a Henry’s Law constant of 3.45x10 - 5 atm

- m 3 /mol (Lyman et al., 1990; Gaffney et al., 1987) using the EPIWIN estimation software (U.S. EPA 2011b ), the volatilization half

- lives of acetonitrile are calculated to be approximately 0.5 days for a model river (assum ing a wind velocity of 5 meters per second ( m/sec ) , water depth of 1 meter, current velocity of 1 m/sec, temp at 25 °C ) and 7 .5 days for a model lake (assuming a wind velocity of 0.5 m/sec, water depth of 1 meter, current velocity of 0.05 m/sec, temp at 25 °C )

. The vapor pressure of 88.8 mm Hg (Boublik et al., 1984) suggests that acetonitrile may volatilize from dry soil surfaces as well (HSDB, 2002). This vapor pressure value also indicates that acetonitrile will exist primarily as a vapor when it enters the atmosphere13 (Bidleman, 1988). Due to the weak reactivity of acetonitrile in the atmosphere, transport of the compound from the troposphere to the stratosphere is expected to occur ; acetonitrile has been detected in the stratosphere (Arijs et al., 1983 ). However, wet deposition can be the most significant atmospheric loss process in regions of frequent rainfall, resulting in deposition of acetonitrile to soil and surface water (Hamm et al., 1984; Karl et al., 2004; Sanhueza et al., 200 4 ).
A soil partitioning coefficient (K oc ) value of 5 has been calculated for acetonitrile using its experimental log K ow value of

- 0.34 (Hansch et al., 1995) and a regression

- derived equation (HSDB, 2002; Lyman et al., 1990). Based on this K oc value, acetonitrile is not expected to bind to soil or sediments, but will be highly mobile in soil (HSDB, 2002; Swann et al., 1983). Acetonitrile may leach from soil to groundwater due to its high mobility. Deep

- well injection containment failures are described among the sour ces of groundwater contamination in the U.S. EPA Ground Water Handbook (U.S. EPA, 1987). The potential for dispersion and transport by groundwater would be high for acetonitrile accidentally released into an aquifer due to its high water solubility and re latively low affinity for soils. As discussed in the section 2.6.3 of this report, acetonitrile in subterranean environments may be subject to biotic degradation. The volatilization of acetonitrile from a contaminated ground wate r body is dependent upon multiple factors including available air head space above the groundwater body , air exchange rates , and groundwater p H , temperature and velocity. A cetonitrile released in vapor form from subterranean sources, including groundwater, may permeate through the ground and enter into the atmosphere depending on the geology, temperature and moisture content of the land mass above its release zone. In i nstances where acetonitrile contaminated groundwater makes it way to surface water, it potent ially can volatilize into the atmosphere, the predicted volatilization rates stated above for a river and lake body of water. 2.6.2. Abiotic Degradation. The abiotic degradation of acetonitrile released in the environment is expected to be very slow. Ac etonitrile does not undergo direct photodegradation (HSDB, 2002). T he main abiotic degradation processes are considered to be reaction with hydroxyl radicals in the atmosphere.14 Experimental data for acetonitrile reactions in the atmosphere are presented by Atkinson (1994) as second

- order rate coefficients. Half

- life estimates for the acetonitrile reaction with the hydroxyl radical in air vary from months to over a year, depending on variations in OH radical concentrations and other factors (EPIWIN,(U.S.EPA, 2000a)). The National Library of Medicine’s Hazardous Substance Data Ba nk (HSDB , 2002 ) reports a n estimated half

- life of 620 days based upon an experimental rate constant for the reaction of acetonitrile with hydroxyl radicals of 2.63E

- 1 4 cubic centimeter ( cm3 ) /molecule

- sec and a n atmospheric concentration of 5 x 10 +5 hydroxyl radicals per cubic centimeter

. Limited literature was available addressing the atmospheric reaction products of acetonitrile

. The reaction between acetonitrile and atmospheric hydroxyl radicals is expected to proceed through hydrogen abstraction or addition to the triple bond (Atkinson, 1994; Graedel et al., 1986). Based on other hydrogen abstraction reactions involving hydrocar bon side chains, oxygenated species are expected but no literature was identified stating specific abiotic degradation products. However, photochemical smog chamber studies show that acetonitrile is unreactive towards photochemically generated free radicals (Dimitriades and Joshi, 1977). Abiotic hydrolysis of acetonitrile is extremely slow under environmental conditions (Brown et al., 1975). HSDB (2002) reports a rate constant for the base

- catalyzed hydrolysis of acetonitrile of 5.8 E

- 3/M

- hr and a corresponding half

- life at pH 7 of 150,000 years.

As stated earlier in this report, in 200 9 some 17.1 million pounds of acetonitrile were reported to the TRI as being burned on
- site for energy recovery by the manufacturers (U.S.EPA, 20 1 1 ). During combustion in air the decomposition of acetonitrile initiates at 760 °C

. At 950 °C (with a 2 sec residence time) 99.99% of the acetonitrile will have decomposed ( Britt, 2002 ). The ignition of acetonitrile in excess oxygen results in the production of CO, CO2, H2O, N2 and NO as the major combustion products. The combustion of acetonitrile at temperatures below 900 °C and low oxygen concentrations can result in the production of HCN ( Britt, 2002 ). 2.6.3. Biotic Degradation. The most significant degradation process for acetonitrile is biodegradation

. The biological capability for acetonitrile degradation appears to be widespread over several genera of prokaryotes (organisms that lack a cell nucleus or any other membrane

bound organelles) , and euk aryotes ( organisms whose cells are internally organized into complex1 5 structures enclosed within membranes)

. Experimentally observed degradation rates vary based upon several factors including media type, temperature, acetonitrile concentration, available oxygen, specific microorganisms involved and the degree of acclimation. Within this section additional information is provided on specific microorganisms capable of degrading acetonitrile along with studies assessing the degradation of acetonitrile under aerobic and anaerobic conditions. It is important to note that many of the biodegradation results provided below are from studies that used test conditions likely representative of aerobic or anaerobic digestion within a sewage treatment plant. These con ditions are not necessarily reflective to relevant environmental conditions , in particular those conditions that are associated with underground injection wells or subterranean environments.
2.6.3.1 Mec hanisms of Biotic Degradation
. While mechanisms for the microbial degradation of acetonitrile have been proposed, they have been elucidated using pure microorganism strains and/or acetonitrile as the sole carbon source. Whether or not these mechanisms occur in the environment is unknow n. A single

- step mechanism for the direct intracellular conversion of acetonitrile to acetic acid and ammonia via the enzyme nitrilase (See Figure 2

- 1) has been proposed (Chapatwala et al., 1992; Li et al., 2007; Nawaz et al., 1989).

Figure 2

- 1
A two

- step enzymatic hydrolysis mechanism for the aerobic biodegradation of acetonitrile has also been proposed (see Figure 2

- 2 ) (Dadd et al., 2000; Rezende et al., 1999; Linardi et al., 1996; Babu et al., 1995; Nawaz and Chapatwala , 1991; Nawaz et al., 1989 ; Chapatwala et al., 1992; Li et al., 2007). In this mechanism, acetonitrile undergoes enzymatic degradation to form acetamide, which then readily biodegrades to form acetic acid and ammonia

.

Figure 2

- 2 C N H 3 C H 2 O N i t r i l a s e H 3 C O H O N H 3 +

C N H 3 C H 2 O N i t r i l e h y d r a t a s e H 3 C N H 2 O H 2 O A m i d a s e H 3 C O H O N H 3 +16 Since acetamide is not f ormed in the single

- step mechanism and is expected to be degraded in the two

- step process, acetic acid and ammonia are expected to be the predominant end products of the aerobic biodegradation of acetonitrile (Graham et al., 2000; Li et al., 2007; Nawaz et al., 1989). 2.6.3.2 Aerobic Biodegradation Studies
. In review of the studies available on the aerobic biodegradation of acetonitrile, several of the studies have been conducted using acclimated microorganisms in sludge or sewage media conditions and may not be reflective of environmental conditions of interest. As stated previously, experimentally observed biotic degradation rates are dependent u pon several factors including media , temperature, available oxygen, acetonitrile concentration , specific microorganisms involved and the degree of acclimation. For acetonitrile, t he Chemicals Evaluation Research Institute (CERI, 2003) reports an average of 32% degradation observed in 21 days based upon biochemical oxygen demand (BOD) testing (OECD Test Method 301C). In a study by Urano and Kato (1986b) that used activated sludge as a microbial inoculum, the lag period of acetonitrile degradation in an aqueous media was about 1 day, after which the compound, at an initial concentration of 100 ppm, w as degraded by 17% after 3 days and 40% after 10 days based upon the BOD. Acclimated mixed microbial cultures isolated by an enrichment culture technique provided a 58% degradation of 100 ppm acetonitrile in an aqueous solution in 5 days (BOD test) (Babeu and Vaishnav, 1987). Earlier studies reported that the biodegradation of acetonitrile using activated sludge or sewage proceed s slowly with unacclimated microorganisms in comparison to acclimated microorganisms (Kalmykova and Rogovskaya, 1978; Ludzack et al., 1958; Placak and Ruchhoft, 1947). Ludzak et al. (1961) reported the results of a 28

- day biodegradation study that used a semi

- continuous feed of previously acclimated activated sludge. Under the conditions of the briefly described study, 82

- 94% of the theoretical BOD was achieved when an initial concentration of 139 ppm of acetonitrile was introduced. A graphical representation in the Ludzack et al. paper indicated that acetonitrile degradation proceeded more slowly prior to acclimation of the s ludge. Acetonitrile at an initial concentration of 490 ppm reached 0

- 17% of the theoretical BOD after 24 hours in a screening test using an activated sludge inoculum that was apparently unacclimated (Placak and Ruchhoft, 1947 )

. The National Institute of Technology and Evaluation (NITE, 2002) reports that 84% degradation of acetonitrile from an initial17 concentration of 100 ppm was observed in 28 days using an activated sludge inoculum (OECD Test Method 301C). The biodegradability of acetonitrile was als o observed with river water; the 12

- day ThOD (theoretical oxygen demand) test measured the generation of CO 2 from Ohio River water spiked with acetonitrile at concentrations ranging from 0.1

- 25 mg/L. After 12 days, a total generation of CO 2 was observed e quivalent to 40% of the ThOD estimated for acetonitrile (Ludzack et al., 1959). Acclimation of the microorganisms was examined by re

- dosing; the degradation was 5 times faster after acclimation and was also 4 times faster at 20 °C than at 5 °C (Ludzack et al., 1958; Ludzack and Ettinger, 1960). In a more recent study, Li et al.(2007), studied the degradation of higher concentration s of acetonitrile by a mixed culture of microorganisms obtained from raw activated sewage sludge. I t is important to note that the conditions this study was done under are likely representative of anaerobic digestion within a sewage treatment plant and might not be applicable to anaerobic conditions found in the environment. The experimental medium use d was a synthetic mineral salt solution. The microorganisms used in the study were acclimated with acetonitrile

. A s analyzed by gas chromatographic analysis for acetonitrile, complete degradation of acetonitrile was realized in about 10 hours for initial acetonitrile concentrations of 1g per liter, and 75 hours for initial concentrations of 5g per liter acetonitrile. At 10 g per liter acetonitrile, even a greater reduction in the degradation rate was observed with only 75% degradatio n after 186 hours ( Li et al., 2007)

. Concerning media type and its impact on acetonitrile degradation, b ased upon studies conducted by Takemoto et al., (1981) t he biodegradation of acetonitrile is expected to be slower in seawater than in freshwater

. 2.6.3.3 Anaerobic Biodegradation Studies
. Limited literature was identified concerning the anaerobic biodegradation of acetonitrile. In the one recent study, Li et al., (2007), had sewage sludge acclimated aerobically for 2 months using acetonitrile as the sole source of carbon and nitrogen. I t is important to note that the conditions this study was done under are likely representative of anaerobic digestion within a sewage treatment plant and might not be18 applicable to anaerobic conditions found in the environment , in particular conditions relevant to underground injection wells or subterranean environments

. The acclimated mixed culture was used to access degradation rates of acetonitrile at initial concentrations of 1 and 2 g/L under both aerobic and anoxic conditions in a synthetic mineral medium salt solution . With acetonitrile at an initial concentration of 1 g/L, complete degradation was observed within 20 hours under anoxic conditions. The parallel study done by Li et al. (2007) under aerob ic conditions resulted in a faster degradation rate of 10 hours for complete degradation for an initial concentration of 1g/L. This is consistent with the findings of Boethling et al. (1995) which observed for various organic chemicals classes anaerobic b iodegradation rates were generally slower than aerobic biodegradation rates, with the exception of some chemical classes (e.g., nitroaromatic, halogenated hydrocarbons, but not nitriles). A t a higher concentration of 2 g/L acetonitrile, observed anaerobic degradation was significantly slower with complete degradation being observed within 212 hours (versus 20 hours) . A cetamide, acetic acid, and ammonia were detected during the anaerobic biodegradation process which suggests that these products may be poss ible under anaerobic conditions (Li et al., 2007). Another anaerobic study utilizing sewage sludge was conducted by Nagel et al., (1995). An inoculum obtained from low

- solids anaerobic digesters at the Denver Municipal Waste Water Reclamation Plant in Denver, Colorado , demonstrated an 83% reduction of acetonitrile on a dry municipal solid waste feedstock mixed with a liquid nutrient soluti on . The 83% reduction of acetonitrile was observed after a 90

- day incubation period at 37 o C using an initial acetonitrile concentration of 50 mg/L.
2.6.3.4 Biodegradation Organisms
. Several microorganisms have been identified as demonstrating the abili ty to degrade acetonitrile. Nawaz et al. (1989) described the inducible capability of a soil isolate, Pseudomonas putida , to degrade acetonitrile to acetic acid via the
two

- step process (Figure 2

- 2). The organism can metabolize a wide range of aliphatic nitriles and amides, but not aromatic nitriles. The amidase of this organism has a specific activity at least 150 times greater than that of the nitrile hydratase, indicating that when the microbe generates acetamide during acetonitrile metabolism, this will be rapidly converted to acetic acid and ammonia. Other Pseudomonads, including P. marginalis (Babu et al., 1995) and P.19 aeruginosa (Nawaz and Chapatwala, 1991; Nawaz et al., 1989) have been isolated from soils. These organisms have the capacity to degrade acetonitrile and a wide range of other nitriles, and to rapidly produce acetic acid and ammonia as end products. A metabolically diverse group of eukaryotic microorganisms can degrade acetonitrile, including environmental isolates of different strains of yeast and fungi, which perform this process in spite of very high concentrations of acetonitrile. The organisms c apable of this type of metabolism included Candida famata (Linardi et al., 1996), Debaryomyces spp., Geotrichum spp., Pichia spp., Cryptococcus spp., Aureobasidium spp., Rhodotorula spp., Tremella spp., Hanseniaspora spp. (Rezende et al., 1999), and Candid a guilliermondii (Dias et al., 2000). All isolated fungi or yeast that were able to degrade acetonitrile used the two

- step processes to convert it to acetamide, and then to acetic acid and ammonia. Several of these strains could grow on very high concent rations of acetonitrile: Candida famata could grow on concentrations as high as 4%, and Cryptococcus sp. strain UFMG

- Y28 was able to grow on acetonitrile at concentrations of 97 mM (approximately 0.4%) (Rezende et al., 1999).
2.6.4 Environmental Distributi on Assessment
. Table 2

- 2 presents a model (U.S. EPA, 20 1 1b ) estimate of how, on a percentage basis, acetonitrile would distribute in the environment if released continuously to the air, water and soil environmental compartments at rates of 98 1.2 , 1 8

. 2 and 0.0 8 kg/hour, respectively. Also, advection was turned off by setting the time values to 10,000 hours. This emission scenario is based on industrial release data reported to the T RI by U.S. manufacturers in 200 9 (U.S. EPA 20 11 ). The model estimate (EPI Suite version 4.0, (U.S. EPA 20 11b ) ) is based upon the estimated half

- life values stated in Table 2

- 2, the half

- life being the time necessary for a reduction in 50 percent of the mass of acetonitrile in the environmental compartment identified (air, water, soil or sediment), due to environme ntal fate processes. These half

- life values were also generated by the model ( U.S. EPA, 20 1 1b ). Based upon the model, air is the environmental compartment where the largest mass of acetonitrile will accumulate ( 79. 8 %), followed by water, and then soil

.20 Table 2

- 2, Environmental Distribution of Acetonitrile Compartment Mass amount (percent) Half
- life Values (hours)
Air 79. 8 1.54x10 3
Water 16.7 360
Soil 3.5 720
Sediment 0.03 2 9 3.24x10 3

Source: EPI Suite, (U.S. EPA, 20 11b ) Independent of the release scenario and model estimates reported above, it needs to be reiterated that of the millions of pounds of acetonitrile produced each year, the vast majority of releases are reported by manufacturers as being disposed of on
- site in underground injection wells (U.S. EPA 2011) . Based upon data reported to the T RI (U.S. EPA, 201 1 ), 17.7 million pounds were released in 200 9 , of which 1 7

. 4 million pounds were disposed of on

- site in underground injection wells. Based upon the informatio n presented in section 2.6.3.2 and 2.6.3.3 of this report, it can be anticipated that the a ctual bio degradation of acetonitrile in underground injection wells where high concentrations of the chemical and anaerobic or oxygen limiting conditions are likely to exist , will be considerably slower than observed bio degradation rates under aerobic conditions for relatively low concentrations of the chemical.
2.6.5. Toxic Transformation Products. As noted in section 2.6.3 , biodegradation is the most significant degradation process for acetonitrile in the environment. The major transformation products of acetonitrile in the environment are expected to be acetic acid and ammonia (Graham et al., 2000). Although some acetamide may be genera ted as a part of the two

- step biodegradation process noted in section 2.6.3, it is expected to biodegrade rapidly resulting in the formation of acetic acid and ammonia (Graham et al., 2000; Li et al., 2007; Nawaz et al., 1989 Urano and Kato, 1986a,b).21 Abio tic hydrolysis of acetonitrile is extremely slow under environmental conditions (Brown et al., 1975). HSDB (2002) reports a rate constant for the base

- catalyzed hydrolysis of acetonitrile of 5.8E

- 3/M

- hr and a corresponding half

- life at pH 7 of 150,000 years. Based on the reaction pathway for the abiotic hydrolysis of nitriles, the hydrolysis products of acetonitrile are expected to be acetamide, acetic acid, and ammonia (Smith and March, 2001). Acetamide will degrade under abiotic conditions, in particular in reaction with hydroxyl radicals. The National Library o f Medicine’s Hazardous Substance Data Ba nk (HSDB, 2010) reports an experimental rate constant for the reaction of acetamide with hydroxyl radicals of 2.1E

- 12 cm3/molecule

- sec and a corresponding half

- life of 7.6 days. Hydrolysis rates for acetamide are much slower. Using reported rate constants for the acid

- and base

- catalyzed hydrolysis of acetamide of 0.030/M

- hr and 0.17/M

- hr, respectively, a corresponding half

- life at pH 7 of 3,440 years was calculated. (HSDB, 2010). Concerning the biodegradation of acetamide, Urano and Kato (1986b) us ing activated sludge (30 mg/l) as a microbial inoculum and acetamide at an initial concentration of 100 mg/l ( 100 ppm ) in an aqueous medium at 20 °C , observed 50% degradation after approximately 4 days based upon the BOD

. In an earlier study, r eported aceta mide incubated with activated sludge resulted in a 29.6

- 45.7% of the theoretical BOD after 24 hours ( Malaney et al., 1962 as reported in HSDB, 2010) In the study by Li et al., (2007), the observed acetamide formed as a part of the biodegradation of acetonitrile under aerobic conditions peaked at 1800 mg/L about 8 hours after the initial exposure of acetonitrile to acclimated bacteria. Within 30 hours, concentration levels of acetamide were reduced to non

- detectable levels. As reported earlier in this report, some 17.1 million pounds of acetonitrile are burned on

site for energy recovery by the manufacturers (U.S.EPA 20 1 1 ). During combustion in air the decomposition of acetonitrile initiates at 760 °C . At 950 °C (with a 2 sec residence time) 99. 99% of the acetonitrile will have decomposed ( Britt , 2002 ). The ignition of acetonitrile in excess oxygen results in the production of CO, CO2, H2O, N2 and NO as the major combustion22 products (Britt , 2002) . The combustion of acetonitrile at temperatures below 900 °C and low oxygen concentrations can result in the production of hydrogen cyanide ( Britt , 2002 )

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3.1 S UMMARY Acetonitrile is metabolized to inorganic cyanide which is known to react with trivalent iron of cytochrome oxidase in mitochondria and block the chemical reduction of oxygen needed for cellular respiration, thus leading to cytotoxic anoxia (Albaum et al., 1946). The toxicity of acute acetonitril e exposure is believed to be mediated, in whole or in part, through this mechanism. Following acute inhalation exposure of 3,800 ppm acetonitrile to hamsters on a single day during gestation, an increase in maternal toxicity and mortality were observed, as well as an increase in severe fetal abnormalities, including exencephaly, encephalocoele, and rib fusions, at higher exposure concentrations (Willhite, 1983). Acute oral ingestion of acetonitrile on a single day during gestation, a decrease in fetal bo dy weight was observed in hamsters at 100 mg/kg, and dysmorphogenic features, including misdirected allantois and/or trunk and caudal extremity, were observed in rats at 2 , 000 mg/kg (Willhite, 1983). Mortality was not observed in dams exposed to 2 , 000 mg/ kg acetonitrile on GD 10; however, dams in all groups exhibited clinical signs of toxicity including piloerection, prostration, and/or tremors, and caused unspecified maternal weight loss between GDs 10 and 12 (Saillenfait and Sabaté, 2000).
Short

- term ( 10

- day) inhalation exposure to 200 ppm acetonitrile resulted in evidence of immunotoxicity in B6C3F1 mice; including a decrease in white blood cell counts, segmented neutrophil and monocyte counts, IgG levels, and in the number of plaques formed per spleen, as measured by the Cunningham Plaque assay (ImmuQuest Laboratories, 1984; Hazelton Laboratories, 1983a). In addition, a decrease in red blood cell counts, hematocrit, and hemoglobin was also observe d. Subchronic inhalation exposure to acetonitrile resulted in hepatocellular cytoplasmic vacuolization and increased liver weights in mice (Hazelton Laboratories, 1983a), as well as an increase in mortality in rats at 1,600 ppm (calculates to approximately 50 5 mg/kg

- day ) and in mice at 800 ppm (calculates to approximately 40 2 mg/kg

- day ) (NTP, 1996). Effects were not observed in rats or mice following chronic exposure to 400 ppm (calculates to approximately 12628 mg/kg

- day ) acetonitrile in rats a nd 200 ppm (calculates to approximately 100 mg/kg

- day ) acetonitrile in mice (NTP, 1996). NTP (1996) concluded that the evidence for carcinogenicity via inhalation of acetonitrile in male F344/N rats was equivocal, and that there was no evidence of carcino genicity in female rats or in either male or female B6C3F1 mice. Examination of the pattern of death in the 13

- week mouse NTP study suggests that the mortality observed at the 800 ppm exposure level is not the result of an acute exposure but rather the re sult of repeated exposures. The first mortality in this group was not observed until day 20 and deaths continued on days 21, 45, 69 and 89 of the 91 day study. This pattern of mortality is dissimilar to that observed in the 1 , 600 ppm group where initial deaths were seen in the first week and all mice died by day 21. Since the 2

- year NTP study was conducted at dose levels below those that caused mortality in the 13

- week study, there are no available studies that provide data on the mortality pattern from longer term exposure to concentrations at or near the doses that caused mortality in the 13

- week study. Oral and inhalation exposure during gestation (GD6

- 18, 19, 20) in rats and rabbits resulted in both maternal and developmental toxicity. Maternal morta lity was observed in rats at inhalation concentrations of 400 ppm (Mast et al., 1994 ), 1,827 ppm ( Saillenfait et al., 1993 ), oral doses of 375 mg/kg

- day ( Johannsen et al., 1986), at exposure to 3,800 ppm in hamsters (Willhite, 1983), and at oral doses of 30 mg/kg

- day in rabbits (Argus Research Labs, 1984). In rats, an increase in post

- implantation loss and in the number of fetuses with unossified sternebrae and a decrease in number of live fetuses per dam were observed at a dose of 275 mg/kg

- day (Johanns en et al., 1986)

. A n increase in the percentage of nonlive implants and early resorptions per litter were observed at exposure to 1,827 ppm (Saillenfait et al., 1993). A decrease in the average number of live fetuses per litter was observed in rabbits at 30 mg/kg

- day (Argus Research Labs, 1984). The observations of developmental toxicity were also observed at doses at which maternal toxicity or mortality were observed. However, it is inadequate to assume that the developmental effects result only from m aternal toxicity, and the results may indicate that both lifestages, the adult and developing offspring, are sensitive to the dose level (U.S. EPA, 1991).29 Under the Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005), the available data provide ina dequate information to assess the carcinogenic potential of acetonitrile. The data from a NTP 2

- year cancer bioassay indicate that the carcinogenic potential of acetonitrile in the rat and mouse is low (NTP, 1996). Based on the available data, and given the severity of the effect, mortality, EPA concludes that there is sufficient evidence to support a concern for moderately high toxicity from exposure to acetonitrile.
3.2 K INETICS AND METABOLISM 3.2.1. Absorption
. H uman and animal studies indicate that acetonitrile is readily absorbed from the gastrointestinal tract, through the skin and the lungs (U.S. EPA, 1999; WHO 1993). In humans, male and female test subjects absorbed 74% of acetonitrile from un

- inhaled cigarette smoke when held in the mouth for 2 seconds, and 91% was absorbed when cigarette smoke was inhaled to the lungs (Dalhamn et al., 1968a,b). Acetonitrile has been shown to absorb rapidly in dogs exposed to 16,000 ppm (26,880 mg/m 3 ) for 4 hours. Resulting blood cyanide levels peaked and rea ched steady state concentrations of 305

- 433 μg/100 mL after 3 hours (Pozzani et al., 1959).
3.2.2. Distribution
. Acetonitrile is rapidly, widely, and fairly uniformly distributed throughout the body in both humans and laboratory animals (U.S. EPA, 1999; WHO, 1993). A human autopsy of an individual who died 2 days following inhalation of acetonitrile vapors found cyanide in the brain, spleen, lungs, and kidneys, but not liver (Amdur, 1959; WHO, 1993). Animal studies have shown hydrocyanic acid (HCN , i.e. , hydrogen cyanide ), a metabolite of acetonitrile, has been found in the brain, heart, kidney and spleen of rats after inhalation of acetonitrile vapors (Haguenoer et al., 1975). Whole

- body autoradiography in male mice injected intravenously with acetonit rile radiolabeled with 14 C in the methyl group indicated that radioactivity was widely distributed throughout the body (e.g., liver, thymus, reproductive organs, nasal secretions, oral cavity, esophagus, and stomach contents) (U.S. EPA, 1999). Additionall y, oral exposure of animals resulted in metabolites found primarily in the spleen, stomach, and skin (U.S. EPA, 1999).30 3.2.3. Metabolism
. Acetonitrile metabolism has been well established in the scientific literature. The basic mammalian metabolic pathw ay is detailed in Figure 3

- 1. Acetonitrile is metabolized to inorganic cyanide through the intermediate production of hydrogen cyanide. Data from Freeman and Hayes (1988) demonstrate that the metabolism to cyanide is oxygen

- and NADPH

dependent, and medi ated by cytochrome P450 isozyme 2E1 (or P

- 450j) production of a reactive intermediate, methyl cyanohydrine (Ahmed et al., 1992; Feierman and Cederbaum, 1989; Willhite and Smith, 1981). The cytochrome P450 2E mixed

- function oxidases are responsible for the biotransformation of many xenobiotics and are highly inducible at the gene level by many substrates, including alcohol (Lu and Cederbaum, 2008). Formaldehyde and formic acid are also by

- products of acetonitrile metabolism (Ahmed et al., 1992). Cyanide i s further oxidized and conjugated to thiocyanate, a less toxic compound that is excreted in urine, but one that has been shown to interfere with thyroid function (Hartung, 1982). Conversion is mediated by rhodanese , a sulfurtransferase found in liver and human nasal respiratory mucosa (Hartung, 1982, Lewis et al., 1991). Rhodanese is an inducible enzyme and levels of rhodanese increase in response to increasing blood and tissue cyanide concentrations. Sites with the greatest abundance of rhodanese, such as hepatocytes in close proximity to the blood supply to the liver, bronchial epithelial cells, and proximal tubule cells of the kidney, are located to maximize conversion of cyanide to thiocyanate following both oral and inhalation exposure ( Sylvester and Sander, 1990 ). In vitro studies by Lewis et al. (1991) were unable to saturate human nasal tissue rhodanese enzymes, suggesting the concentration of sulfur

containing donor molecules, such as thiosulphate, is the rate

- limiting factor for the conversion o f cyanide to thiocyanate. It has been suggested that if thiocyanate is not readily excreted it may be partially reconverted to cyanide by thiocyanate oxidase (Hartung 1982). Cyanide also can be oxidized to cyanate ion with further oxidation to formic aci d (McMahon and Birnbaum, 1990).31 Figure 3
- 1. Metabolic pathway for acetonitrile. Case reports of acute human exposures, as well as studies in animals, have demonstrated that toxicity from acute acetonitrile poisoning occurs hours after exposure due to the slow conversion of acetonitrile to cyanide (Mueller and Borland, 1996; Michaelis et al., 1991; Freeman and Hayes, 1985). Freeman and Hayes (1985) found peak blood cyanide concentrations 7.5 hours after dosing rats with acetonitrile, and these concentrations remained significantly elevated 34

- 36 hours after exposure. Michaelis et al. (1991) reported elimination half

- lives for acetonitrile and cyanide via oral ingestion in human as 32 hours, and 15 hours, respectively. In a study by Pozzani (1959), rats and dogs were exposed via inhalation for 7 hours/day, 5 days/week for 90 days. In rats, thiocyanate was not completely eliminated between daily exposures, but was almost completely eliminated during the 2.5

- day rest period over the weekend. In dogs, however, thiocyanate continued to be eliminated beyond the 2.5

- day rest period over the weekend demonstrating species differences. Acetonitrile conversion to cyanide appears to be about 10

- fold greater in rat nasal ethmoturbinate microsomes than in liver microsomes (Dahl and Waruszewski, 1989), due to high P450 content in the rat nasal cavit y (U.S. EPA, 1999). The nasal cavity of the rat, in particular

C H 3 C N CYP2E1 C H 2 O H C N C H 2 O

H C N S C N H C O O H C N acetonitrile methyl cyanohydrin formaldehyde sulfur donor (thiosulfate)

hydrogen cyanide rhodanese thiocyanate

cyanate formic acid

2-aminothiolazine-4-carboxylic acid32 the olfactory region, also has a high concentration of rhodanese (Dahl and Waruszewski, 1989). Rhodanese has also been found in the respiratory epithelium of human nasal tissue; however, Lewi s et al. (1991) showed that the rhodanese in human nasal tissue had a lower affinity (higher K m ) for cyanide and a lower rate (lower V max ) for cyanide metabolism than rhodanese from rat nasal tissue. K(m) is the concentration of substrate which permits th e enzyme to achieve half V(max)(the maximum rate of the reaction )

. Taken together these data suggest that cyanide may be detoxified to thiocyanate more rapidly in rat nasal tissue than in animal species with lower amounts or bioactivity of the enzyme , suc h as humans (U.S. EPA, 1999). Lewis et al. (1991) further reported an induction of rhodanese activity and an inability to saturate rhodanese enzymatic activity at high potassium cyanide concentrations using human nasal tissue in vitro. This was in contras t with a previous study that showed no substrate induction of rhodanese activity and saturation of rhodanese activity at high substrate levels using rat nasal tissue in vitro. Due to the dynamic nature of rhodanese activity it is not possible to directly compare the human and rat capacity or rate of cyanide conversion to thiocyanate via rhodanese. Further confounding the issue, Lewis et al. (1991) reported that rhodanese activity was statistically lower in smokers than in nonsmokers, and individual K m a nd V max values suggest that rhodanese in smokers may have a decreased affinity for cyanide. It is postulated by the authors that “increased exposure to cyanide or other pollutants in the tobacco smoke appears to have suppressed enzyme activity.” If chron ic exposure to cyanide does result in decreased rhodanese activity, it may provide a mechanism by which cyanide would accumulate after chronic exposure to acetonitrile.
3.2.4. Excretion
. A pproximately 80% of an ingested dose of inorganic cyanide is conver ted to thiocyanate, the majority of which is rapidly excreted in the urine (WHO, 2004). The Pozzani (1959) study showed variable excretion in rats and dogs exposed to acetonitrile via inhalation for 7 hours/day, 5 days/week for 90 days. In rats, thiocyan ate was not completely eliminated between daily exposures, but was almost completely eliminated during the 2.5

- day rest period over the weekend. In dogs, however, thiocyanate continued to be eliminated beyond the 2.5

- day rest period over the weekend demon strating species differences. It should be noted that this may33 be due to variable excretion rate and/or variable metabolism rates. Formic acid is also readily excreted in the urine of humans and animals. A secondary urinary metabolite that has been dete cted after administration of acetonitrile in drinking water of rats is 2

- aminothiolazine

- 4

- carboxylic acid (Swenne et al., 1996). In addition, acetonitrile may also be excreted unchanged in the urine or in expired air, although at extremely low levels.
3.3 EFFECTS OF ACUTE EXPOSURE Humans acutely exposed to sublethal doses of acetonitrile developed effects that are also generally attributed to metabolism of acetonitrile to cyanide (U.S. EPA, 1999). WHO (1993) reported several cases in which children or adults ingested ~ 250 to 4 , 000 mg/kg acetonitrile. Symptoms exhibited by poisoning victims include anxiety, confusion, hyperpnea, dyspnea, rapid pulse, unconsciousness, and convulsions (WHO, 1993). Cyanide was detected in the blood of these individuals. Case reports of acute occupati onal exposure to acetonitrile indicate that workers exhibited nausea, shallow and/or irregular respiration, and impaired motor activity. An autopsy of a worker who died shortly after exposure revealed cerebral, thyroid, liver, splenic, and renal congestio n (WHO, 1993). Gastric erosion has been reported in individuals who ingested acetonitrile (Ballantyne, 1983; Way, 1981).
In animals, oral LD 50 values have been reported for the mouse (269

- 453 mg/kg) and the rat (1 , 730

- 4 , 050 mg/kg) and inhalation LC 50 val ues of 12,000, 16,000, and 7 , 551

- 12,435 ppm have been reported for the rat for 2, 4, and 8 hour exposures, respectively, and for the mouse following 1

- 2 hour exposures (2 , 300

- 5 , 700 ppm) (WHO, 1993). Willhite (1981) reported the 1

- hr LC 50 in mice for sever al aliphatic nitriles including propionitrile at 163 ppm, n

- butyronitrile at 249 ppm, and acetonitrile at 2 , 693 ppm. A recent study (Moore et al., 2000) reported a slightly higher oral LD 50 of 617 mg/kg for Crl:CD

- 1(ICR)BR mice and an inhalation LC 50 of 3 , 587 ppm for this strain. Observational signs of toxicity reported in animals after acute exposure to acetonitrile include dyspnea, tachypnea, tremors, and convulsions in various studies (WHO, 1993).34 

### 3.4 EFFECTS OF REPEATED EXPOSURE A summary of key short

- term, subchronic, and chronic studies of acetonitrile is included in Table 3

- 1.

Table 3

- 1. Summary of Effect Levels from Key Short

- term, Subchronic, and Chronic Exposure Studies of Acetonitrile

Study Summary Exposure Levels ppm NOAEL LOAEL or FEL Effect (incidence)

10

- day inhalation, 6 hours/day, 5 days/week, 14

- day period; B6C3F1 mouse (ImmuQuest Laboratories, Inc., 1984; Hazelton Laboratories, 1983a) 0, 100, 200, 400 ppm 100 200 Decreases in red and white blood cell, segmented neutrophil and monocyte counts, hematocrit, and hemoglobin; decrease in serum immunoglobulin G (IgG); decrease in number of plaques formed per spleen 92

- day inhalation, 6 hours/day, 5 days/week; B6C3F1 mouse (Hazelton Laboratories, 1983a) 0, 25, 50, 100, 200, 400 ppm 100 200 Increases in hepatocellular vacuolization and liver weight measures

92

- day inhalation, 6 hours/day, 5 days/week; F344 rat (Hazelton Laboratories, 1983b) 0, 25, 50, 100, 200, 400 ppm 400 NA Effects were not observed

90

- Day inhalation, 6 hours/day, 5 days/ week, F344/N rat (NTP, 1996) 0, 100, 200, 400, 800, 1 , 600 ppm 800 1600 Mortality (1/20 at 800 ppm , 9/20 at 1 , 600 ppm )

90

- Day inhalation, 6 hours/day, 5 days/ week, B6C3F1 mouse (NTP, 1996) 0, 100, 200, 400, 800, 1 , 600 ppm 400 800 Mortality (1/20 at 400 ppm , 5/20 at 800 ppm , 20/20 at 1 , 600 ppm ).35 Table 3

- 1. Summary of Effect Levels from Key Short

- term, Subchronic, and Chronic Exposure Studies of Acetonitrile
Study Summary Exposure Levels ppm NOAEL LOAEL or FEL Effect (incidence)

2

- Year inhalation bioassay, 6 hours/day, 5 days/week, F344/N rat (NTP, 1996) 0, 100, 200, 400 ppm 400 NA No effects on clinical signs, body weight, survival, complete histopathology, hema tology, targeted organ weights

2

- Year inhalation bioassay, 6 hours/day, 5 days/week, B6C3F1 mouse (NTP, 1996) 0, 50, 100, 200 ppm 200 NA No effects on clinical signs, body weight, survival, complete histopathology, hematology, targeted organ weights

3.4.1. Effects of Short
- term Exposure

. In an unpublished short

- term immunotoxicity study, female B6C3F1 mice were exposed to 100, 200, or 400 ppm acetonitrile for 6 hours/day, 5 days/week, for 10 days during a 14

- day period (ImmuQuest Laboratories, Inc., 1984; Hazelton Laboratories, 1983a). The duration

- adjusted doses in mice were calculated as approximately 0, 49, 98, and 196 mg/kg

- day 2 . Statistically significant decreases in red and white blood cell counts, segmented neutrophil and monocyte counts, hem atocrit, and hemoglobin were reported in mice exposed at 200 and 400 ppm acetonitrile, but not in mice exposed to 100 ppm, and the magnitudes of change were described as low by Hazelton Laboratories (1983a). A dose

-

2 Throughout this chapter, concentrations used in inhalation studies are converted to equivalent doses in terms of mg acetonitrile per kg body weight per day (mg/kg

- day) to permit comparison of effect levels. The conversion uses default assumptions (inhalat ion rate and body weight for laboratory animals) from USEPA (1988) along with the exposure frequency and daily exposure time. The generic equation for conversion is as follows: ppm to mg/kg

- day =[ppm x (molecular weight/24.5) x rat ventilation rate (m 3 /da y) x duration of exposure (hr/day)] ÷ rat body weight (kg) = mg/kg

- day For example, for Fisher 344 rats: 100 ppm x (41.05 (molecular weight) / 24.5) x 0.19 m 3 /day x (6hr / 24hr) x (5 days / 7day) ÷ 0.180 kg = 32 mg/kg

- day. For B6C3F1 mice: 100 ppm x (41. 05 / 24.5) x 0.033 m 3 / day x (6hr / 24hr) x (5 days / 7days) ÷ 0.020 kg = 49.9 mg/kg

- day36 dependent decrease of 25.8, 33.1, and 48 .1% in serum immunoglobulin G (IgG) levels was also observed following short

- term exposure to 100, 200, or 400 ppm acetonitrile (approximately, 49, 98, and 196 mg/kg

- day), respectively (ImmuQuest Laboratories, Inc., 1984). Serum immunoglobulin M (IgM) was significantly decreased at 200 and 400 ppm , although the decrease was not dose

- dependent. The number of plaques formed per spleen, as measured by the Cunningham Plaque assay, was reduced by 12.9, 22.2, and 18.5% at 100, 200, and 400 ppm ; and, compared to the increase in mean spleen weights of 19 and 14% at 200 and 400 ppm , is suggestive of a depression in this immune function (ImmuQuest Laboratories, 1984). Lymphocyte counts were decreased in female mice at 200 and 400 ppm . ImmuQuest Laboratories, Inc. (1984) also reported thymic atrophy in mice exposed to concentrations of 200 and 400 ppm. The observed effects were considered to be compound

- related (ImmuQuest Laboratories, Inc., 1984). 3.4.2. Effects of Subchronic Exposure
. In two unpublished studies, male and female B6C3F1 mice and F344 rats (10 per sex per group) were exposed by inhalation to acetonitrile concentrations of 0, 25, 50, 100, 200, and 400 ppm for 6 hours/day, 5 days/week for 92 days (Hazelton Laboratories, 1983 a, b). The duration

- adjusted doses in rats calculate to approximately 0, 8, 16, 31, 63, and 125 mg/kg

- day. Effects were not observed in the F344 rats (Hazelton Laboratories, 1983b; as cited in U.S. EPA, 1999). The duration

- adjusted doses in the mice wer e approximately 0, 12, 24, 49, 98, and 196 mg/kg

- day. One male mouse died during the course of the study in each of the 50 , 200 , and 400 ppm exposure groups; however, the mortality was not considered exposure

- related (Hazelton Laboratories, 1983a). Hepatocellular cytoplasmic vacuolization was observed in mice at each exposure level, including the control group, and the vacuolization observed in the two highest dose groups was of greater severity than in the other dose groups. A statistically significant increase in relative liver weight was observed in male mice at 400 ppm and in female mice at 100, 200, and 400 ppm

. A n increase in liver/bra in weight ratio was observed in male mice at 400 ppm and in female mice at 100 and 400 ppm . A NOAEL of 100 ppm and LOAEL of 200 ppm were identified based on the observed hepatocellular vacuolization and increases in liver weight measures.37 In the 13

- week s ubchronic study (NTP, 1996), F344/N rats and B6C3F1 mice (10/sex/group) were exposed whole

- body to acetonitrile concentrations of 0, 100, 200, 400, 800, or 1 , 600 ppm, 6 hours/day, 5 days/week for 13

- weeks. This calculates to approximately 0, 32, 63, 126, 253, and 50 5 mg/kg

- day for Fisher 344 rats and approximately 0, 50, 101, 201, 402, and 805 mg/kg

- day for B6C3F1 mice. The mortality observed in the NTP (1996) 13

- week inhalation study of acetonitrile is detailed in Table 3

- 2. The mortality incidence of the 13

- week repeated exposure rat study (See Table 3

- 2) was 0/20 in each of the 0, 100, 200 and 400 ppm groups, 1/20 in the 800 ppm group (one death occurring on day 5), and 9/20 in the 1 , 600 ppm group (four deaths occurring on day 2, one each on days 7, 9 , 10, 11, 23). Clinical signs at the two high

- concentration groups included hypoactivity and ruffled fur during the first week. Ataxia, abnormal posture, and clonic convulsions occurred in the 1 , 600 ppm males that died. Decreased serum concentrations of triiodothyronine were observed in 1 , 600 ppm exposed female rats, only. Differences were not observed in rat thyroxine or thyroid stimulating hormone levels, and thyroid hormone effects were not seen in the mouse. In addition, a statistically significant decrease in hematocrit, hemoglobin, and erythroc y tes was observed in male rats at 1 , 600 ppm and in female rats at ≥800 ppm. Statistically significant changes in organ weights were also observed, primarily, at the highest dose in male rats and at ≥800 ppm in female rats, and include decreases in absolute and relative thymus weight, increases in absolute and/or relative liver and kidney weight, and decreases and increases in brain and heart weight, respectively. Histopathologic effects were limited to rats that died at 800 and 1 , 600 ppm; effects observed include congestion, edema, and hemorrhage in the lung alveoli.
In the 13

- week mouse studies, the mortality incidence (see Table 3

- 2) was 0/20 in each of the 0, 100 and 200 ppm groups, 1/20 in the 400 ppm group (death occurring on day 13), 5/20 in the 800 ppm group (deaths occurring on days 20, 21, 45, 69, 89) and 20/20 in the 1 , 600 ppm group (all deaths occurring by day 21) (NTP, 1996). Statistically significant changes in organ weights were observed, inc luding, increased absolute and/or relative liver weight at ≥100 ppm in males and ≥400 ppm in females and increased relative lung weight at ≥200 ppm in males. Liver cytoplasmic vacuolization and forestomach hyperplasia were observed in male and female mice38 Table 3
- 2. Mortality in NTP 13
- week Acetonitrile Inhalation Study (NTP, 1996)

Species Sex Concentration (ppm) Total number dying/number in study Week of death

Week 1 Week 2 Week 3 Week 4 Weeks 6
- 13 a

Rats M 800 1/10 1

F 800 0/10

M 1 , 600 6/10 4 1 1

F 1 , 600 3/10 1 2

Mice M 400 0/10

F 400 1/10 1

M 800 1/10 1

F 800 4/10 2 2

M 1 , 600 10/10 2 8

F 1 , 600 10/10 4 6

a In the 6

- 13 week column, the 800ppm male mouse died on day 45, the 800 ppm female mice died on day 69 and day 89

surviving through the end of the study from several treatment groups. These forestomach effects may not be specific to acetonitrile and have been observed in mice from other NTP inhalation studies, suggesting some other factors associated with the conditions of exposure (NTP, 1996; Melnick et al. 1990). However, gastric erosions have been observed in humans following cyanide exposure and may represent an effect of acetonitrile exposure (NTP, 1996).39 Therefore, from the 13

- week NTP study , for rats the NOAEL is 400 ppm and the frank effect level ( FEL ) is 800 ppm, with one death. For mice (the most sensitive species), the NOAEL is 200 ppm and the FEL is 400 ppm, with one death. The FEL is a level of exposure or dose that produces irreversible, adverse effects at a statistically or biologically significant increase in frequency or severity between those exposed and those not exposed. 3.4.3. Effects of Chronic Exposure
. In the chronic 2

- year study, groups of 55

- 56 male and 56 fema le F344/N rats were exposed to 0, 100, 200 or 400 ppm ( calculates to approximately 32, 63, or 126 mg/kg

- day) acetonitrile by inhalation for 6 hours per day, 5 days per week for 2 years. Groups of 60 male and 60 female B6C3F mice were exposed to 0, 50, 100 , or 200 ppm ( calculates to approximately 25, 50, and 101 mg/kg

- day) acetonitrile by inhalation for 6 hours per day, 5 days per week for 2 years (NTP, 1996). Excess mortality, above that seen in the control groups, was not observed in rats or mice at any dose level tested. The doses levels for the NTP (1996) 2

- year rat inhalation study were selected “based on reduced survival of the 800 ppm males and 1 , 600 ppm males and females in the 13

- week study” (NTP 1996). The dose levels for the NTP (1996) 2

- year mouse inhalation study were selected “based on reduced survival and gross and histopathological lesions in the 400, 800, and 1 , 600 ppm groups of male and female mice in the 13

- week study” (NT P 1996). The NTP (1996) report states that, “the dose response f or acetonitrile toxicity is quite steep, with little evidence of adverse effects in animals exposed to concentrations lower than those that resulted in death, or in survivors in exposure groups in which some animals died.” As detailed

- previously in this technical review, using mortality as the critical effect for acetonitrile exposure, the LOAELs for the NTP (1996) 13

- week rat and mouse studies were 1 , 600 ppm and 800 ppm, respectively. For both the rat and mouse, the highest doses used in the 2

- year chronic studies were lower than the lowest concentrations that resulted in mortality in the 13

- week subchronic study. The levels at which effects were obs erved in the 13

- week study were not tested in the chronic study, and, in addition, two of the three principal reviewers of the NTP (1996) study suggested that the highest doses applied in the chronic study (200 ppm mouse; 400 ppm rat) were too low and one reviewer suggested dose levels should have been as high as 800 ppm (NTP, 1996). In discussing the dose selections for the 2

- year chronic studies, NTP (1996) stated that:40 “The primary factor influencing the selection of exposure concentrations for the 2

- year studies was the mortality observed in the 13

- week studies. The greater number of deaths of mice than rats at 400 and 800 ppm, coupled with the forestomach effects in mice, led to the selection of 200 ppm as the highest exposure concentration for mi ce and 400 ppm for rats. Thus, the high exposure concentrations selected were within a factor of two of those that were lethal in the 13

- week studies.” As discussed previously, a single death in the 400 ppm mouse and 800 ppm rat treatment groups is not st atistically significant when compared to control groups. Therefore, the doses selected for the 2

- year study were four

- fold lower than those that were lethal in the 13

- week studies. 3.4.4. Carcinogenicity
. There are no studies evaluating the carcinogenicit y of acetonitrile in humans. Other data pertinent to the assessment of potential carcinogenicity include a National Toxicology Program (NTP) cancer bioassay in mice and rats, and genotoxicity assays.
In the NTP 2

- year inhalation cancer bioassay of acetoni trile (NTP, 1996), F344/N rats (56/sex/group) were exposed to acetonitrile concentrations of 0, 100, 200, or 400 ppm, 6 hours/day, 5 days/week, for 103 weeks ( calculates to approximately 0, 28, 57, and 114 mg/kg

day) and B6C3F1 mice exposed to concentratio ns of 0, 50, 100, or 200 ppm for 111 weeks ( calculates to approximately 25, 50, and 101 mg/kg

- day )

. An interim necropsy at 15 months involved 8 rats (each sex) and 10 mice (each sex). Complete histopathological examinations were conducted on all animals at this time and hematological parameters and liver, kidney, and lung weights were measured. Clinical signs and body weight were assessed throughout the study. After 2 years, animals were necropsied and examined for gross and microscopic alterations.

At the 15

- month interim necropsy, neoplastic or nonneoplastic lesions attributable to exposure were not observed in rats. Significant changes in survival, body weights, or clinical appearance were not observed in rats following a 103

- week exposure to acetoni trile. Histopathological examination did not reveal neoplastic or nonneoplastic lesions in any organs of41 exposed females. In male rats, a statistically significant increase in the incidence of basophilic hepatic foci was observed in the 200 and 400 ppm g roups (15/48, 22/47, 25/48, and 31/48), but the foci were not atypical in appearance (NTP, 1996). Although there was a marginally significant positive trend in the incidence of adenoma, carcinoma, or adenoma and carcinoma (combined) in the liver of male r ats, no significant dose

- related trend was present after incidences were adjusted for survival using the life table test (NTP, 1996). The incidences of hepatocellular adenomas and carcinomas in male rats were not statistically significantly increased in t he treated animals based on pairwise comparison with incidences in control animals. Also, the tumor incidences at 400 ppm were only slightly higher than the historical control range. In mice, no changes in the survival of the treated animals were observ ed, compared to the survival in control animals. Body weights were similar for all groups, and treatment

- related clinical signs were not observed. In contrast to the 13

- week studies, there were no concentration

- related effects of liver weight. At the 15

- month interim sacrifice, the only nonneoplastic change observed was a significant increase in the incidence of squamous hyperplasia in the forestomach of 200

- ppm females (incidences were 0/10, 1/10, 0/10, and 6/10 in control, low

- , mid

- , and high

- dose gro ups, respectively). At terminal sacrifice, the incidence of alveolar/bronchiolar adenomas was significantly increased in male mice following administration of the high concentration (6/50, 9/50, 8/48, and 18/50 at 0, 50, 100, or 200 ppm , respectively). C ombined incidences of alveolar/bronchiolar adenomas or carcinomas (10/50, 14/50, 14/48, and 21/50 at 0, 50, 100, or 200 ppm , respectively) were also significantly increased in 200 ppm males. NTP (1996) concluded that the evidence for carcinogenicity via inhalation of acetonitrile in male F344/N rats was equivocal (See T able 3

- 3 ) . Although there was a statistically significant positive trend in the incidences of hepatocellular adenomas, carcin omas, and adenomas and carcinomas (combined) in male rats only, the incidences were not statistically significant by pairwise comparison or by life table analysis. According to NTP (1996), the only partially positive neopl a stic finding in rats was a margin al increase in the inciden c e of hepat o cellular adenoma or carcinoma (combined) in male rats exposed to 400 ppm. Statistical significance was42 achieved (P=0.045) for a dose

- related trend, but pairwise comparisons of the exposed group incidences versus the c ontrol incidence were not significant. In 400 ppm males, the incidence of hepat o cellular adenoma or carcinoma (combined) (5/48, 10%) is higher than the historical control incidence of 3.8% for inhalation studies and slightly higher than the upper range of incidences observed in any one control group (8%). A 10% incidence has been noted in four previous control groups in dosed feed studies. Male rats exhibited an increased incidence of basophilic foci in liver that was statistically significant in the 200 and 400 ppm groups. Although the appearance of these foci was not atypical, as those more closely related to the carcinogenic process, altered hepatocellular foci are generally considered to be preneoplastic. There was no evidence of carcinogenicity in female rats or in either male or female B6C3F1 mice (NTP, 1996).
Table 3

- 3 : Data from NTP 2

- Year Inhalation Study (NTP, 1996) 0 ppm 100 ppm 200 ppm 400 ppm
Number examined 48 47 48 48

Basophilic Focus 15 22 25* 31**
Clear Cell Focus 3 1 2 5
E o sinopilic Focus 3 7 5 10

Mixed Cell Focus 1 1 1 5

Hepatocellular Adenoma or Carcinoma Overall rate(a) 1/48(2%) 1/47(2 %) 1/48(2%) 5/48(10% )

Adjusted rate(b) 3.3% 3.4% 4.8% 25.2% Terminal rate(c) 0/11(0%) 0/13(0%) 0/9(0%) 3/17 (18% )43 First incidence (days 637 623 673 693

Logistic regression test P= 0.045 P= 0.757 P=0.758 P=0.164

*Significantly different (P<=.05) from control by the logistic regression test. **; P<= 0.01 (a) Number of animals with neoplasm per number of animals with liver examined microscopically (b) Kaplan

- Meyer estimated neoplasm incidence at the end of the study after adjustment for intercurrent mortality (c) Observed incidence in animals surviving until the end of the study

3.5 DEVELOPMENT AND REPRODUCTIVE TOXICITY
The onl y study of repeated human exposure to acetonitrile was a case

- referent study evaluating reproductive endpoints. This study of medical records of Finnish women employed in state laboratories was designed to determine if there was an association between spo ntaneous abortion and exposure to acetonitrile. The odds ratio for seven cases involving exposure to acetonitrile was close to unity (Taskinen et al., 1994). However, the small number of cases and other limitations (e.g., self

- reported exposure) preclude any definitive conclusions from this study. The developmental toxicity of acetonitrile in animals has been studied in several species with both inhalation and oral exposure using both single and repeated dosing regimens (Table 3

- 4 ) . Pregnant Syrian golden hamsters were exposed via inhalation to 0, 1 , 800 (6 animals), 3 , 800 (6 animals), 5 , 000 (6 animals), or 8 , 000 (12 animals) ppm acetonitrile for 1 hour on GD 8 and then sacrificed on GD 14 (Willhite, 1983). One dam exposed to 3 , 800 ppm approximately 160 mg/kg

- day) died 3 hours after dosing, after exhibiting dyspnea, tremors, hypersalivation, ataxia, and hypothermia. At 5 , 000 ppm ( approximately 210 mg/kg

- day), all animals exhibited irritation and excessive salivation; one dam in this group died after displaying dyspnea, hypothermia, and tremors (Willhite, 1983). Six abnormal offspring from two litters of this group exhibited exencephaly, encephalocoele, and rib fusions. In the 8 , 000 ppm group ( approximately 338 mg/kg

- day), clinical effects included r espiratory difficulty, lethargy, ataxia, hypothermia,44 irritation, and gasping (4/12 dams), followed by tremors, deep coma, and death (3/12 dams). Fetotoxicity occurred in offspring of dams exposed to 8 , 000 ppm, as evidenced by decreased fetal body weight compared to controls (Willhite, 1983). Five of the nine surviving litters at 8 , 000 ppm developed severe axial skeletal dysgraphic disorders; one fetus exhibited extrathoracic ectopia cordis with accompanying defects in the sternum. This study identifies a maternal LOAEL of 3 , 800 ppm for maternal mortality (1/6) and a LOAEL of 5 , 000 ppm and NOAEL of 3 , 800 ppm for the observed developmental effects (exencephaly and rib fusions). Willhite (1983) also examined the developmental effects of ingested acetonitril e on pregnant hamsters. Pregnant Syrian golden hamsters (6

- 12/group) received gavage doses of 0, 100, 200, 300, or 400 mg/kg acetonitrile (in distilled water) on GD 8, and were sacrificed on GD 15. Death occurred in 1/6 and 4/12 dams given 300 and 400 mg /kg, respectively. A statistically significant decrease in fetal body weight was observed at ≥100 mg/kg and an increased number of resorptions was observed at ≥200 mg/kg. The 300 and 400 mg/kg doses were associated with a significant increase in grossly malformed offspring, primarily rib anomalies. Treatment with thiosulfate prevented both maternal and developmental toxicity, suggesting that the observed effects may be attributable to cyanide production from acetonitrile. A LOAEL of 100 mg/kg was identi fied for developmental effects based on decreased fetal body weight and a NOAEL of 200 mg/kg and a LOAEL of 300 mg/kg for maternal toxicity. Saillenfait and Sabaté (2000) administered a single oral dose of acetonitrile (2,000 mg/kg dissolved in olive oil) to pregnant Sprague

- Dawley rats (3 rats/group) on GD 10. Dams were sacrificed on GD 12 and the implantation sites and embryos were examined. No deaths were reported in acetonitrile

- treated dams; however, dams in all groups exhibited clinical signs of
tox icity including piloerection, prostration, and/or tremors. Except for sodium cyanide and allylnitrile, all of the compounds caused unspecified maternal weight loss between GDs 10 and 12. Embryo viability was not affected by any treatment (Saillenfait and Sabaté, 2000). In the acetonitrile group, dysmorphogenic features (including misdirected allantois and/or trunk and caudal extremity) were observed in 12 of 34 (35.3%) embryos examined (2 litters affected/3 litters examined), compared to 0 of 52 controls (4 litters examined). These specific dysmorphogenic features were also observed in embryos of sodium cyanide

- treated dams. The45 occurrence of developmental effects consistent with exposure to cyanide in animals exposed to acetonitrile implicates exposure to free cyanide as a likely source of the effects. The single dose of acetonitrile (2 , 000 mg/kg) was identified as a lowest

- observed

- adverse

- effect level (LOAEL) for developmental toxicity (dysmorphogenic features). Four groups of 53 Sprague

- Dawley rats were subjected to respiratory exposure (whole

body) concentrations of 0, 100, 400, and 1 , 200 ppm ( calculates to approximately 0, 45, 181, and 543 mg/kg

- day) acetonitrile for 6 hours/day, 7 days/week during GDs 6

- 19 (Mast et al., 1994). Each group inc luded 10 non

- pregnant animals for comparison, 10 pregnant animals for evaluating maternal blood for acetonitrile and cyanide, and 33 pregnant animals for evaluating developmental toxicity. On GD 20, the animals were sacrificed and the uteri were weighed a nd examined and the number and position of implants was determined. Mortality occurred in pregnant (2/33) and in non

- pregnant female rats (1/10) exposed at 1 , 200 ppm and in pregnant female rats (1/33) at 400 ppm. The death at 400 ppm was suggested by the investigators to be the result of a possible spontaneous cerebral hemorrhage. No treatment

- related effects on fertility, fetal viability, or developmental endpoints were observed. An increase in the incidence of supernumery ribs was observed at 100 ppm, however this effect was not statistically significant nor observed in other treatment groups. The 400 ppm exposure level was identified as a NOAEL and 1 , 200 ppm as LOAEL for increased maternal mortality. The developmental toxicities of eight aliphatic m ononitriles, including acetonitrile, were studied in rats (Saillenfait et al., 1993). Five groups of 20

- 23 Sprague

- Dawley rats were subjected to respiratory exposure (whole

- body) concentrations of 0, 1 , 000, 1 , 287 , 1 , 592 and 1 , 827 ppm ( calculates to approx imately 0, 453, 583, 721, and 827 mg/kg

- day) acetonitrile vapor 6 hours/day on GDs 6

- 20. All dams surviving to study termination were sacrificed on GD 21 and the uteri were weighed, fixed, and examined. Maternal deaths (8/20 dams; 40%) occurred at the hi ghest concentration, 1,827 ppm. A markedly increased percentage of nonlive implants per litter (resorptions and dead fetuses; one litter was completely resorbed) and early resorptions per litter were observed at 1,827 ppm, along with a decrease (not signif icant) in the mean number of live fetuses per litter. No effects on maternal or developmental endpoints were observed at lower doses. A LOAEL of 1 , 827 ppm and NOAEL of 1 , 592 ppm was determined for maternal46 mortality; 1 , 827 ppm was also identified as a LOAEL for developmental toxicity based on the increased percentage of nonlive implants per litter. Smith et al., (1987) evaluated in utero and early post

- natal growth and viability of offspring in pregnant Long

- Evans Hooded r ats (five groups of n=20

- 25) exposed to 0, 50, 150, 300, and 500 mg/kg acetonitrile via gavage administration from GD 7

- 21. Dams were allowed to deliver and the duration of pregnancy, litter size, sex ratio, and litter weight were determined. Any dam not delivering by GD 23 was euthanized and examined. At PND 6, the litters were randomly culled to 6

- 8 offspring per litter and then further culled to a total of 4 (2 males, 2 females). On PND 41

- 42, the pups were sacrificed and subjected to a gross necropsy . At the two highest doses of 300 and 500 mg/kg ACN, excessive maternal morbidity and mortality occurred. At 300 mg/kg/day, 11 females were moribund and had to be sacrificed at mid

gestation , and at 500 mg/kg/day, 16 females died prior to termination (>5 0% mortality). Acetonitrile did not produce any changes in pup birth weight, weight gain, or viability of the surviving litters at any dose level tested. Likewise, there were no changes in pregnancy rates, resorptions, or signs of toxicity in the offsprin g and no frank malformations were observed. It was noted by the authors that even at doses that were lethal to the dam, no overt signs of toxicity were observed in the fetuses. A NOAEL of 150 mg/kg

- day and LOAEL of 300 mg/kg

- day was identified for matern al toxicity. Johannsen et al. (1986) exposed pregnant Sprague

- Dawley rats via gavage to 0, 200, 375, 750, 1 , 500, and 3 , 000 mg/kg

- day acetonitrile on GDs 6

- 19 as a pilot range finding study. None of the exposed animals survived to study termination at ≥75 0 mg/kg

- day, and maternal deaths were noted at 375 mg/kg

- day with a concomitant increase in the mean number of early resorptions and a decrease in maternal body weight. Signs of maternal toxicity were not observed at 200 mg/kg

- day. In another developmenta l study (Johannsen et al., 1986), four groups of 25 pregnant Sprague

- Dawley rats were exposed via gavage to 0, 125, 190, and 275 mg/kg

- day on GDs 6

- 19. On GD 20, any surviving dams were sacrificed and evaluated for the number and location of viable and no n
- viable fetuses, number of early and late resorptions, number of total implantations and number of corpora lutea. At the high est dose of 275 mg/kg

- day, two females47 died by the end of the study (8% mortality) and four others were described in the study re port as appearing emaciated. A dose

- dependent, biologically significant increase in post

- implantation loss per dam and a decrease in number of live fetuses per dam were observed in the high dose group; however, neither effect was statistically significant . The number of fetuses with unossified sternebrae was increased dose

- dependently, although not statistically significant, at the high dose. A LOAEL of 275 mg/kg

- day and NOAEL of 190 mg/kg

- day was identified for developmental toxicity, as represented by the increase in post

- implantation loss, the decrease in number of live fetuses per dam, and the increase in number of fetuses with unossified sternebrae. Four groups of 25 artificially

- inseminated New Zealand white rabbits were exposed by gavage to 0, 2, 1 5, and 30 mg/kg

- day acetonitrile on GDs 6

- 18 (Argus Research Laboratories, 1984). Surviving rabbits were sacrificed on GD 29 and examined for implantation, corpora lutea, resorptions, and fetal survival. Treatment

- related deaths (5/25, or 20%) were repor ted in maternal rabbits exposed to 30 mg/kg

- day, with deaths on GD 12, 15, 17, 18, and 19; thus, no rabbits died until exposure had continued for at least 6 days. The dams also demonstrated a statistically significant increase in anorexia at 30 mg/kg

- day, as well as ataxia, decreased motor activity, bradypnea, dyspnea, and impaired righting reflex. It should be noted that an observation made by the Study Director, Dr. A.M. Hoberman, expressed April 1, 2002, to Mr. D.E. Strother, stated that gastrointestin al irritation and decreased food consumption were coincident adverse effects. The average number of live fetuses per litter was statistically significantly decreased at 30 mg/kg

- day. The high dose (30 mg/kg

- day) represents a LOAEL for maternal mortality (5/25) and for developmental toxicity based on the decreased number of live fetuses per litter. The mortality occurred at the highest dose tested in the study (LOAEL = 30 mg/kg

- day); however, the high mortality (20%) in dams also indicated a FEL for this d ose level. Because the deaths occurred with a short duration of exposure and dosing errors were not reported, it is likely that the observed effects were chemical

- related.

Table 3

- 4 provides a summary of the developmental toxicity studies available for ace tonitrile.48 Table 3

- 4 Summary of Effect Levels from Key Developmental Toxicity Studies of Acetonitrile Study Information (exposure, duration, test animal, citation) Exposure (ppm or mg/kg/day) Maternal Effects NOAEL, LOAEL Developmental /Reproduction Effects NOAEL, LOAEL

Inhalation , 1 hr, GD8, Syrian golden hamster (Willhite, 1983) 0, 1 , 800, 3 , 800, 5 , 000, 8 , 000 ppm NOAEL: 1 , 800 FEL: 3 , 800 Mortality at doses >=3 , 800 ppm (1/6, 1/6, 3/12) clinical effects in dams at 8 , 000 ppm NOAEL: 3 , 800 LOAEL: 5 , 000 Significant increase in severe fetal abnormalities (exencephaly and encephalocele)

Oral Single dose, GD8, pregnant Syrian golden hamster (Willhite, 1983) 0, 100, 200, 300 , 400 mg/kg NOAEL: 200 FEL: 300 Deaths occurred 1/6 and 4/12 dams at 300 and 400 mg/kg NOAEL: None LOAEL: 100 Decreased fetal body weight

Oral , Single dose GD10, Sprague

Dawley rat (Saillenfait and Sabate, 2000) 0, 2 , 000 mg/kg NOAEL: none LOAEL: 2 , 000 Weight loss, clinical signs (piloerection, prostration, tremors) NOAEL: none LOAEL: 2 , 000 Dysmorphogenic features

Inhalation 6 hrs/day, GDs 6

- 19, Sprague Dawley rat (Mast et al. 1994) 0, 100, 400, 1,200 ppm NOAEL: 100 FEL: 1200 2/33 pregnant rat mortality and 1/10
non
- pregnant rat at 1200 ppm NOAEL: 400 LOAEL: None No effects on fertility, fetal viability or developmental endpoints

Inhalation 6 hrs/day, GDs 6

- 20 Sprague

Dawley rat (Saillenfeit et al., 1993) 0, 1 , 000, 1 , 287, 1 , 592, 1 , 827 ppm NOAEL: 1 , 592 FEL: 1 , 827 Mortality 8/20 at 1 , 827 ppm NOAEL: 1 , 592 LOAEL: 1 , 827 Increase percentage of non

- live implants/litter Oral , GDs 7

- 21 Long

Evans Hooded Rats (Smith et al., 1987) 0, 50, 150, 300, 500 mg/kg

- day NOAEL: 150 FEL: 300 At 300, 11 females moribund, sacrificed NOAEL: 500 LOAEL : None No changes in birth weight, or viability of litters; no changes in pregnancy rates, resorptions or signs of toxicity in offspring. Oral , GDs 6

- 19, Sprague

- Dawley Rat (Johannsen et al, 1986) 0, 125, 190, 275 mg/kg

- day NOAEL: 190 FEL 275 Mortality (2/25 at 275, emaciated appearance 4/25 mg/kg

- day) NOAEL 190 LOAEL: 275 Increase in post

- implantation loss, number of fetuses with unossified sternebrae, decrease in life fetus number per dam
Oral , GD6

- 19, Sprague

- Dawley rats (Johannsen et al (1986) 0, 200, 375, 750, 1,500, 3,000 mg/kg

- day NOAEL: None L OAEL: 200 FEL: 375 NOAEL: 200 LOAEL: 375 Increase in mean number of early resorptions Oral , GDs 6

- 18, New Zealand white rabbit (Argus Research Labs, 1984) 0, 2, 15, 30 mg/kg

- day NOAEL: 15 LOAEL: 30 Mortality ( 5/25 at 30 mg.kg

- day NOAEL: 15 LOAEL: 30 D ecreased number live fetuses/litter49 3.5.1. In Vitro Developmental and Reproductive Studies
. In vitro studies performed by Saillenfait and Sabaté (2000), conceptuses were explanted from uteri of 10

- day pregnant dams and cultured with medium containing al iphatic nitriles with or without hepatic microsomes and NADPH. After 46 hours of culture, viable embryos (embryos with a beating heart) were examined for growth and morphological development. The yolk sac was assessed for its vascularization and circulat ion and its diameter was recorded , as were the crown

- rump length, head length, and number of somite pairs of the developing embryo. All of the nitriles tested, including acetonitrile, produced concentration

- dependent decreases in growth and differentiatio n, including decreased yolk sac diameter, crown

- rump length, head length, and number of somites, and increases in the incidences of morphologically abnormal embryos, including defects in prosencephalon, mesencephalon, eye, maxillary process, auditory syste m, olfactory system, and flexion (Saillenfait and Sabaté, 2000). The addition of the biotransformation system to the culture medium did not change acetonitrile

- induced embryotoxicity, while the percentage of sodium cyanide

- treated embryos with defects dro pped from 100% without the metabolic system to 0% with microsomes and NADPH. Saillenfait et al. (2004) tested the same series of eight nitriles in micromass cultures of rat embryo limb bud cells. The results suggested that acetonitrile at doses of 10

- 300 μg/ml , concentrations below those causing cytotoxicity, was not teratogenic (did not inhibit cell differentiation. The authors cautioned that the agreement was poor between the results of the rat limb bud micromass culture assay and in vivo data on the t eratogenic effects for these particular nitriles.
3.6 GENOTOXICITY Acetonitrile results in gene mutation assays have been predominantly negative, with some notable exceptions. In mutation assays with a variety of Salmonella typhimurium strains, both with and without metabolic activation, acetonitrile has failed to produce a measurable effect (U.S. EPA, 1999). Similarly, acetonitrile did not produce mutations in Saccharomyces cerevisiae D7 (U.S. EPA, 1999). Both sister chromatid exchan ges (SCEs) in Chinese hamster ovary (CHO) cells and gene conversions in S. cerevisiae were increased in the absence of S9 metabolic activation, but not in the presence of S9 (U.S. EPA, 1999). In contrast, chromosomal50 aberrations in CHO cells were increase d in the presence of S9, but not in the absence of S9 (U.S. EPA, 1999). Further, effects in chromosomal aberration assays have demonstrated interference with chromosome segregation both in vivo and in vitro using the Drosophila melanogaster model system ( U.S. EPA, 1999). Increases in micronucleated erythrocytes have been reported in mice treated with acetonitrile via inhalation (NTP, 1996) and intraperitoneal injection (Schlegelmilch et al., 1988); however, the incidence of micronucleated polychromatic er ythrocytes in bone marrow and peripheral blood was not affected in groups of male and female mice (5/sex) given single intraperitoneal injections of 100 or 125 mg/kg acetonitrile in corn oil (Jones et al., 2001). Positive micronucleus assays can indicate either clastogenic activity or interference with chromosome segregation. Sehgal et al. (1990) obtained in vitro evidence that acetonitrile does inhibit microtubule assembly in taxol

- purified Drosophila or mouse microtubules. Acetonitrile induced aneuploi dy in treated mature oocytes of D. melanogaster females exposed either as larvae or as adults (Osgood et al., 1991a,b). Further, acetonitrile induced mitotic aneuploidy in S. cerevisiae exposed to 5% acetonitrile (Zimmermann et al., 1985). There are no d ata on the capacity of acetonitrile to produce heritable genetic mutations in humans or animals. The overall genotoxicity data are inconsistent, and indicate that acetonitrile is not a point mutagen, but does interfere with chromosomal segregation (U.S. E PA, 1999). Genotoxicity following exposure to acetonitrile is not considered to be among the most sensitive endpoints.
3.7 NEUROTOXICITY Neurotoxicity after subchronic exposure to acetonitrile was observed in rhesus monkeys exposed to acetonitrile via inhalation of 330, 660, or 2 , 510 ppm for 7 hours/day, 5 days/week ( calculates to approximately 28, 55, and 209 mg/kg

- day) (Pozzani et al., 1959). Interpretation of this study is limited by the small number of treated animals, the absence of a control grou p of monkeys, and the use of animals that had been used in previous experimentation. Of four exposed rhesus monkeys, the one monkey (female) exposed to 2 , 510 ppm died on the second day, the two female monkeys exposed at 660 ppm died by day 51, and the one male monkey exposed to 330 ppm was sacrificed at day 99 of exposure. At that time, it exhibited considerable excitability. One monkey in each of the 2 , 510, 660, and 330 ppm groups was found to have51 dural and subdural hemorrhages. While the hemorrhages were initially considered to be related to acetonitrile exposure, an independent review of the brain pathology slides by Dr. Karl Jensen of U.S. EPA’s National Health and Environmental Effects Research Laboratory (NHEERL) concluded that there was no eviden ce of treatment

- related neurotoxicity (U.S. EPA, 2000).
3.8 POTENTIAL MODES OF ACTION The available data indicate that the observed mortality following acute exposure to acetonitrile is mediated through its metabolism to cyanide (Hartung , 1982; WHO, 1993; U.S. EPA 1999). Cyanide inhibits a number of enzyme systems, the main target of which is cytochrome c oxidase, the terminal oxidase in the mitochondrial respiratory chain (Ballantyne, 1987). Cyanide has long been known to react with tr ivalent iron of cytochrome c oxidase in mitochondria and block the reduction of oxygen needed for cellular respiration, thus leading to cytotoxic anoxia (Albaum et al., 1946). The acute toxicity of acetonitrile is believed to be mediated, in part, through this mechanism. If the conversion of acetonitrile to cyanide leads to systemic cyanide concentrations that bind the cytochrome c oxidase for a critical period, it may lead to death. Although acute oral doses of cyanide cause cardiovascular, respiratory, and neurophysiological changes, the brain appears to be the organ that is most sensitive to acute cyanide toxicity (WHO, 2004). Death from acute exposure to cyanide is believed to result from central nervous system depression, subsequent to inhibition of brain cytochrome c oxidase activity (Way, 1984). The role of metabolism may be crucial in determining the acute toxicity of acetonitrile via cyanide production. Mortality may result from the absorption and metabolism of high doses of acetonitrile into cyanide at a rate that exceeds the body’s capacity for detoxification. In contrast, slow absorption of the same or greater total load may allow for complete metabolism and not cause mortality due to acute exposure. Acetonitrile is metabolized to inorgani c cyanide but the conversion occurs slowly compared to other nitriles (Hartung, 1982), which may explain the delay in onset of acute symptoms (U.S. EPA, 1999). The mode of action for acetonitrile toxicity following chronic exposure is unknown, although th e metabolism of acetonitrile to its reactive metabolites is likely required for toxicity.52 For chronic acetonitrile exposure to cause mortality through the inhibition of cytochrome c oxidase by cyanide, an accumulation of systemic cyanide reaching a critic al level may be necessary to bind enough cytochrome c oxidase to cause inhibition of cellular respiration. Currently, evidence in humans or animals is not available demonstrating, either qualitatively or quantitatively, that systemic cyanide accumulates o ver time due to chronic acetonitrile exposure. In addition, there is currently not enough data in the scientific literature to quantitatively construct a physiologically

- based pharmacokinetic (PBPK) model of acetonitrile toxicokinetics in humans.

It is also plausible that other biological systems and enzymes, besides cytochrome oxidase, may be inhibited by cyanide (Way, 1984), leading to the accumulation of cellular damage over time and the delayed acetonitrile

- induced mortality observed in the NTP (1 996) mouse study. Lethality from adverse effects of repeat chemical exposures can also result from cumulative, irreversible damage to a critical molecular/cellular target such that the summation of damage eventually exceeds a threshold for non

- viability o f the organism (Klaassen, 2008). However, data demonstrating an accumulation of cellular damage with subsequent cell death following exposure to acetonitrile or cyanide is not available.
3.9 T HYROID TOXICITY Thiocyanate does not demonstrate high acute toxicity, but has been shown to be a potent inhibitor of iodide transport into the thyroid (Klaassen, 2008). Long

- term ingestion of elevated levels of cyanide can result in sustained serum levels of thiocyanate, leading to inhibition of iodide uptake by t he sodium

- iodide symporter in the thyroid gland, reduced production of thyroid hormones T4 and T3, and subsequent thyrotoxicity (Klaassen, 2008). Whether acetonitrile exposure and the subsequent metabolism to thiocyanate is likely to produce this inhibiti on of iodide uptake, and thus decreas ing thyroid hormone synthesis, will depend on the systemic concentrations at which the steady

- state establishes itself with continued treatment. The nature of iodine uptake into the thyroid gland and thyroid hormone pr oduction may contribute to the lack of overt signs of thyrotoxicity, such as alterations in systemic thyroid stimulating hormone, thyroxine, and triiodothyronine levels, as demonstrated in the NTP (1996) studies. Since acetonitrile exposure was administer ed 6 hours/day, 5 days/week, temporary53 activation of the sodium

- iodide transporter may have occurred because of thiocyanate clearance during periods of non

- exposure. In addition, the thyroid gland stores thyroglobulin as colloid within the gland (Klaassen , 2008), and this reservoir has the potential to serve as a reserve for the production of thyroxine, thus maintaining circulating levels. Neuronal pathology that would suggest hypothyroidism was not observed in the NTP (1996) studies. This suggests adeq uate thyroid hormone (T 3 and T 4 ) during development. It should be noted that the animals used in the NTP (1996) 13

- week and 2

- year mouse and rat studies were approximately 6 weeks old when they began exposure. Current literature suggests rat neuronal dev elopment is sensitive to thyroid hormone action from shortly after birth through post natal day 20

- 30 (Oppenheimer and Schwartz, 1997). The exposure periods in the NTP (1996) studies, therefore, appear to occur outside the proposed critical window for neu ronal thryroid hormone sensitivity. In addition to direct effects on the thyroid, a decrease in thyroid hormones following exposure to cyanide, mediated through the cyanide metabolite thiocyanate, may result in reproductive toxicity effects (U.S. EPA, 2010 ). As stated in the Toxicological Review of Hydrogen Cyanide and Cyanide Salts (U.S. EPA, 2010), “research exists to suggest that reproductive tissues in developing and adult animals are responsive to alterations in thyroid hormone levels.”
3.10 HUMAN RE LEV ANCE OF ANIMAL DATA Acetonitrile has been shown to cause death in rabbits, mice, rats, and humans (U

. S

. EPA, 1999). The molecular mechanisms underlying the cause of death for the various species are thought to be the same across species, and the clinical symptoms of exposure in cases of accidental or occupational human exposure are similar to those observed in studies using laboratory animals (WHO, 1993; U.S. EPA 1999). These symptoms of acetonitrile poisoning include neurotoxic effects (i.e. ataxia, tremors and convulsions) commonly associated with cyanide poisoning. This evidence strongly supports equiva lent modes of action for acetonitrile lethality upon acute exposure for test animals and humans, and by extension suggests relevance54 of acetonitrile lethality due to chronic exposures in humans, as seen in animals from the NTP (1996) subchronic mouse inhal ation study. Death was observed in rabbits at dose (30 mg/kg

- day) much lower than that which caused mortality in rats, mice, and hamsters. Species differences in sensitivity are due to variability in toxicokinetic and toxicodynamic parameters. The toxicokinetic parameters include binding affinity (K m ) and metabolic rate (V max ) for cyanide metabolism. These toxicokinetic parameters dictate the ability of species to detoxify acetonitrile and its metabolites, particularly cyanide moieties. Variabilit y between species is further compounded by the tissue specific inducible nature of the enzyme responsible for converting acetonitrile to hydrogen cyanide and/or formaldehyde (cytochrome P450) and the enzyme responsible for converting hydrogen cyanide to th iocyanate (rhodanese). A study comparing the rhodanese activity of rats and humans observed that the respiratory epithelium of human nasal tissue had a lower affinity (K m ) for cyanide and a lower rate (V max ) for cyanide metabolism than rhodanese from rat nasal tissue (Lewis et al., 1991). This suggests that humans may be more sensitive to cyanide than rats because cyanide may be detoxified to thiocyanate more rapidly in rat nasal tissue than in animal species with lower amounts or bioactivity of the enzym e (U.S. EPA, 1999) . No estimation of effects on sensitive human populations is currently available (e.g. fetal effects, genetic polymorphisms, individual variation). In the absence of human data associating chronic exposure of humans with increased risk

o f death, the toxicity observed in animals following repeated exposures to acetonitrile is assumed to be relevant to humans. Case studies reporting mortality in humans after acute accidental poisoning describe effects consistent with those observed after e xposure to inorganic cyanide and other saturated aliphatic nitriles exposures. Animal study findings suggest that the toxic effects of acetonitrile may be correlated with the metabolic release of cyanide and/or formaldehyde, a process that would also occu r in humans.
3.11 CONCLUSION The NTP (1996) 13

- week subchronic inhalation study was identified as the principal study for deriving the Reference Concentration (RfC) in the Toxicological Review of acetonitrile55 (U

. S

. EPA, 1999). The use of mortality as the critical effect for derivation of the chronic Reference Concentration (RfC) for the Integrated Risk Information System (IRIS) is due to the steep dose

- response curve for acetonitrile, which is consistent with the exposure

- response data for other cyani de

- containing chemicals. In addition, the Toxicological Review identified the 400 ppm dose in the NTP (1996) mouse study as a FEL given the death of a mouse at week 2 at 400 ppm and the increased mortality at 800 ppm.
Examination of the dose

- response relationship for acetonitrile

- induced mortality at various time intervals in the NTP 13

- week study indicates that, for mice and rats, subchronic exposure resulted in mortality at lower concentrations, 800 ppm and 1 , 600 ppm respectively, than that causing d eath in acute toxicity studies (e.g ., LC 50 values of 2 , 300 ppm for mice and 12 , 000 ppm for rats ) as cited by the World Health Organization (WHO, 1993 ). It should also be noted that mortality was observed in pregnant rabbits, rats, and hamsters at oral doses as low as 30 mg/kg

- day in rabbits, and inhalation exposure levels as low as 1 , 200 ppm in rats which are similar or lower than the 800 to 1,600 ppm exposure levels that resulted in mortality following subchronic exposure in rats and mice, and may indi cate that pregnancy is a susceptible lifestage

. Under the Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005), the available data provide inadequate information to assess the carcinogenic potential of acetonitrile. The data from a NTP 2

- year cancer bioassay indicate that the carcinogenic potential of acetonitrile in the rat and mouse is low (NTP, 1996). However, because of uncertainty regarding the significance of results due to dose selection in the NTP 2

- year bioassay, the carcinogenic potential of acetonitrile for humans following inhalation, oral or dermal exposure cannot be determined. Based on the available data, and given the severity of the effect, mortality, EPA concludes that there is sufficient evidence to support a concern for moderately high toxicity from exposure to acetonitrile. In assessing mortality following acetonitrile exposure, the patterns in the timing of death across exposures demonstrates the chronic nature of the effect . In the NTP (1996) 13

- week mouse inhalation study, mo rtality was observed in the 800 and 1 , 600 ppm treatment groups. The first incidence of mortality in the 800 ppm treatment group was not observed until day 20 , and single deaths continued on days 21, 45, 69 and 89 of the 13

- week study. This pattern of mor tality is dissimilar to that observed in the 1 , 600 ppm treatment group56 of the NTP (1996) 13

- week mouse inhalation study, where initial deaths were observed in the first week and all mice died by day 21. Based on the observed pattern of death in the 800 pp m treatment group of the NTP (1996) 13

- week inhalation study, beginning at the end of the third week and extending through the termination of the study, it can be reasonably anticipated that additional acetonitrile

- induced mortality would have continued be yond the termination of the study and the sacrifice of surviving animals. Because the mortalities extended from the third week of the study to study termination, the data indicates that the mortality observed in the 800 ppm treatment groups is not due to a single acute exposure to sufficiently high acetonitrile concentrations, but rather is best explained as being the result of long

- term repeated exposures. The observed exposure
- response relationship for acetonitrile demonstrates that a threshold exists at which acetonitrile exposures are sufficient enough to cause mortality, and, as such, mortality would not necessarily be expected following chronic exposure at the doses tested in the NTP (1996) 2

- year study because the acetonitrile exposure levels in the study design were not sufficient to cause mortality. In addition to the concern for increased mortality following repeated dose exposures, immunotoxicity and developme ntal toxicity are also of concerns. Evidence of immunotoxicity was observed in mice following a short

- term exposure to 100 mg/kg

- day, but was not observed in rats following subchronic and chronic exposures. Multiple studies demonstrated toxicity in the d eveloping fetus at oral doses ranging from approximately 30 mg/kg

- day in rabbits and 100

- 275 mg/kg

- day in rats and inhalation concentrations of 1 , 827 ppm in rats

. It should be noted that these doses were also maternally toxic; however, the observation of maternal toxicity does not abrogate all concern with the observations of toxicity in the developing fetus. Figure 3

- 2 provides an exposure

- response array for the toxicological studies of acetonitrile (U.S. EPA, 1991)

.57 Figure 3
- 2. Exposure
- response array for toxicological studies of acetonitrile.60 3.12 REFERENCES Ahmed, A.E., J.P. Loh, B. Ghanayem et al. 1992. Studies on the mechanism of acetonitrile toxicity: I. Whole body autoradiograph ic distribution and macromolecular interaction of 214C

acetonitrile in mice. Pharmacol. Toxicol. 70:322

- 330, as cited in U.S. EPA, 1999. Albaum, H.G., J. Tepperman and O. Bodansky. 1946. The in vivo inactivation by cyanide of cytochrome oxidase and its effects on high energy phosphorous compounds in the brain. J. Biol. Chem. 64:45

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- 633. Argus Research Laboratories, Inc. 1984. Embryofetal t oxicity and teratogenicity study of acetonitrile in New Zealand White rabbits (Segment II evaluation). Washington, DC: Office of Toxic Substances submission. Microfiche No. OTS 507279. Ballantyne, B. 1983. Artifacts in the definition of toxicity by cya nides and cyanogens. Fundam. Appl. Toxicol. 3:400

- 408. Ballantyne, B. and T. Marrs. 1987. Toxicology of Cyanides. In: Clinical and experimental toxicology of cyanides. Bristol: Wright. pp. 41

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- 835 , as cited in U.S. EPA, 1999.
Dalhamn , T; Edfors, ML; Rylander, R. (1968b) Retention of cigarette smoke components in human beings. Arch Environ Health 17:746

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- 450 IIE1 and catalase in the oxidatio n of acetonitrile to cyanide. Chem. Res. Toxicol. 2:359

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- 622. Freeman, J.J. and E.P. Hayes. 1988. Microsomal metabolism of acetonitrile to cyanide. Biochem. Pharmacol. 37:1153

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- 106 , as cited in U.S. EPA, 1999. Hartung, R. 1982. Cyanides and nitriles. In: Patty’s Industrial Hygiene and Toxicology, 3rd Rev. Ed. Patty, F.A., G.D. Clayton, F.E. Clayton et al., eds. New York: Wiley. pp. 4845

- 4900.61 Hazelton Labor atories. 1983a. 90

- Day subchronic toxicity study of acetonitrile in B6C3F1 mice. Final Report (Revised). Prepared for the National Toxicology Program (NTP).
Hazelton Laboratories. 1983b. 90

- Day subchronic toxicity study of acetonitrile in Fischer 344 rats. Final Report (Revised). Prepared for the National Toxicology Program (NTP), as cited in U.S. EPA, 1999. ImmuQuest Laboratories, Inc. 1984. Limited toxicity of inhaled acetonitrile on the immune system of mice. OTS FYI submission. Microfiche No . FYI

- AX

- 0284

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- metabolizing enzyme rhodanese in human nasal respiratory mucosa. Toxicol. Appl. Pharmacol. 108:114

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- butadiene in B6C3F1 mice following 65 weeks of exposure. Environ Health Perspect 86:93

- 98 , as cited in U.S. EPA, 1999. McMahon, T.F. and L.S. Birnbaum. 1990. Age

- related changes in toxicity and biotransform ation of potassium cyanide in male C57B1/6N mice. Toxicol. Appl. Pharmacol. 105:305

- 314, as cited in U.S. EPA, 1999. Michaelis, H.C., C. Clemens, H. Kijewski et al . 1991. Acetonitrile serum concentrations and cyanide blood levels in a case of suicidal oral acetonitrile ingestion. Clin. Toxicol. 29:447

- 458, as cited in U.S. EPA, 1999.
Moore , N.P., R.J. Hilaaski, T.D. Morris et al. 2000. Acute and subacute toxicological evaluation of acetonitrile. Int. J. Toxicol. 19:363

- 364.
Mueller M and C. Borlan d. 1996. Delayed cyanide poisoning following acetonitrile ingestion. Postgrad. Med. J. 73(859):299

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- 05

- 8) in F344/N rats and B6C3F1 mice (i nhalation studies). NTP Technical Report Series 447. Oppenheimer, JH; Schwartz, HL. (1997) Molecular basis of thyroid hormone

- dependent brain development. Endocr Rev 18(4):462

- 75. Osgood, C., M. Bloomfield and S. Zimmering. 1991a. Aneuploidy in Drosophila , IV. Inhalation studies on the induction of aneuploidy by nitriles. Mutat. Res. 259:165

- 176, as cited in U.S. EPA, 1999. Osgood, C., S. Zimmering and J.M. Mason. 1991b. Aneuploidy in Drosophila , II. Further validation of the FIX and ZESTE genetic test systems employing female Drosophila melanogaster . Mutat. Res. 259:147

- 163, as cited in U.S. EPA, 1999. Pozzani, U.C., C.P. Carpenter, P.E. Palm, C.S. Weil and J.H. Nair. 1959. Mammalian toxicity of acetonitrile. J. Occup. Med. 1:634

- 642

. Saillenfait, A.M., P. Bonnet, J.P. Guenier et al. 1993. Relative developmental toxicities of inhaled aliphatic mononitriles in rats. Fundam. Appl. Toxicol. 20:365

- 375 as cited in U.S. EPA, 1999. Saillenfait, A.M. and J.P. Sabaté. 2000. Comparativ e developmental toxicities of aliphatic nitriles: In vivo and in vitro observations. Toxicol. Appl. Pharmacol. 163:149

- 163. Saillenfait, A.M., J.P. Sabaté and C. Gaspard. 2004. Effects of aliphatic nitriles in micromass cultures of rat embryo limb bud cells. Toxicol. In Vitro . 18(3):311

- 318. Schlegelmilch, R., A. Krug and H.U. Wolf. 1988. Mutagenic activity of acetonitrile and fumaronitrile in three short term assays with special reference to autoinduction. J. Appl. Toxicol. 8:201

- 209, as cited in U.S. EPA, 1999. Sehgal, A., C. Osgood and S. Zimmering. 1990. Aneuploidy in Drosophila . III: Aneuploidogens inhibit in vitro assembly of taxol

- purified Drosophila microtubules. Environ. Mol. Mutagen. 16:217

- 224, as cited in U.S. EPA, 1999. Smith, M.K., George, E.L., Zenick, H., Manson, J.M., and Stober, J.A. 1987. Developmental Toxicity of Halogenated Acetonitriles: Drinking Water By

- Products of Chlorine Disinection. Toxicology, 46:83

- 93. Swenne, I., Eriksson, U.J., Christoffersson, R., et al. 1996. Cyanide detoxification in rats exposed to acetonitrile and fed a low protein diet. Fundam Appl Toxicol 32:66

- 71 , as cited in U.S. EPA, 1999.63 Sylvester M; Sander C. (1990) Immunohistochemical localization of rhodanese. Histochem J. Apr 22(4):19 7

- 200. Taskinen, H., P. Kyyronen, K. Hemminki et al. 1994. Laboratory work and pregnancy outcome. J. Occup. Med. 36:311

- 319. U.S. EPA. 1988. Recommendations for and Documentation of Biological Values for use in Risk Assessment. Office of Research and Development, Cincinnati, OH, EPA/600/6

- 87/008 Feburary 1988. U.S. EPA. 1991. Guidelines for Developmental Toxicity Risk Assessment. Risk Assessment Forum, Washington, DC. EPA/600/FR

- 91/001.
U.S. EPA. 1994. Methods for Derivatio n of Inhalation Reference Concentrations and Application of Inhalation Dosimetry. Office of Research and Development, Washington D.C., EPA/600/8

- 90/066F October 1994. U.S. EPA. 1999. Toxicological Review of Acetonitrile. Office of Research and Developm ent. Washington, DC. January, 1999. Available at http://www.epa.gov/iris/toxreviews/0205

- tr.pdf

. U.S. EPA. 2000. OPPT/RAD Decision on Neurotoxicity Endpoint for Acetonitrile. Office of Pollut ion Prevention and Toxics, Washington, DC. U.S.EPA. 2005. Guidelines for Carcinogenic Risk Assessment. U.S. Environmental Protection Agency, Risk Assessment Forum, Washington, DC. EPA/630/P

- 03/001B. Available at http://www.thecre.com/pdf/20050404_cancer.pdf

. U.S. EPA. 2010. Toxicological Review of Hydrogen Cyanide and Cyanide Salts. Office of Research and Development. Washington, DC. Available at http://www.epa.gov/iris/toxreviews/0060tr.pdf Way, J.L. 1981. Pharmacologic aspects of cyanide and its antagonism. In: Cyanide in Biology. Vennesland, B., E.E. Conn, C.J. Knowles et al., eds. New York, NY: Academic Press. pp. 29

- 49. Way, J.L. 1984. Cyanide intoxication and its mechanism of antagonism. Ann. Rev. Pharma. Toxicol. 24, 451

– 459. WHO (World Health Organization). 1993. Environmental Health Criteria 154: Acetonitrile. International Programme on Chemical Safety, Geneva, Switze rland. Available at http://www.inchem.org/documents/ehc/ehc/ehc154.htm

. WHO. 2004. Concise International Chemical Assessment Document (CICAD): 61. Hydrogen Cyanide and Cyanides: Human Health Aspects. Geneva, Switzerland.64 Willhite, C.C. 1981. Inhalation toxicology of acute exposure to aliphatic nitriles. Clin. Toxicol. 18(8) :991

- 1003. Willhite, C.C. 1983. Developmental toxicology of acetonitrile in the Syrian golden hamster. Teratology. 27:313

- 325. Willhite, C.C. and R.P. Smith. 1981. The role of cyanide liberation in the acute toxicity of aliphatic nitriles. Toxicol. Appl. Pharmacol. 59:559

- 602. Zimmermann, F.K., V.W. Mayer and I. Scheel. 1985. Acetone, methyl ethyl ketone, ethyl acetate, acetonitrile and other polar aprotic solvents are strong inducers of aneuploidy in Saccharomyces cerevisiae . Mutat. Res. 149:33 9

- 351.65 CHAPTER FOUR: ECOLOGICAL HAZARD ASSESSMENT

4.1. SUMMARY Based upon a review of the limited acute and chronic toxicity data available for aquatic species (algae, invertebrates, and vertebrates) , adverse environmental effects d ue to acetonitrile can only be anticipated at relatively high exposure concentrations. This conclusion is based on observed acute toxicity values for aquatic organisms that are greater than or equal to 100 mg/L , indicating relatively low acute aquatic toxic ity (Table 4

- 2) 1 (see for example: http://www.epa.gov/oppefed1/ecorisk_ders/toera_analysis_eco.htm#Ecotox)

. Acute toxicity values for freshwater fish and aquatic invertebrates ranged from 100 to 3 , 600 mg/L. The 96 hour LC 50 values for fish and green algae ranged from 100 to 3 , 454 mg/L.

E xperimental chronic toxicity values for green and blue

- green algae, freshwater invertebrates, freshwater fish embryos , and duckweed ranged from 160 to 7 , 300 mg/L , indicating relatively low chronic toxicity. Experime ntal chronic toxicity studies on earthworms reported 100% survival at exposure concentrations of 500 μ g/g soil and 200 μ L /g soil. No data were available to assess chronic toxicity to estuarine/marine fish or invertebrates. Acetonitrile is not expected to bioaccumulate in aquatic organisms (BCF is 3.2 based on log K ow :
- 0.34 )

4.2 ECOLOGICAL TOXICITY
Databases used for the identification of ecological toxicity were ECOTOX (2002) (which contains three databases: AQUIRE, PHYTOTOX, and TERRETOX) and the Office of Pesticide Programs (OPP) Environmental Fate and Effects Division (EFED) Pesticide Ecotoxicity Database. In addition, the TOXLINE dat abase was examined to identify additional aquatic toxicity and population effects studies (See TOXLINE citation in References) . No laboratory data on acute estuarine/marine fish toxicity or chronic estuarine/marine invertebrate and fish

1 Many of the test concentrations were reported as nominal and thus table values could potentially underestimate toxicity as th ey are likely higher than the actual ex posure concentrations. The EPIWIN model was therefore used to calculate a volatilization 1⁄2 life for lakes which was then use d to estimate acetonitrile concentrations remaining in the static exposure systems for the nominal 100 mg/L treatments. The mo del estimated 1⁄2 life of 4 d indicates that approximately 60 to 70 mg/L would be expected to remain at the end of the 96 hour static exposures. Se e section 2.6.1 Environmental Transport for the details of the Model Lake estimation.66 toxicity were foun d for acetonitrile

. L imited data on acute estuarine/marine invertebrate toxicity and algal toxicity were available . Quantitative Structural Activity Relationships (QSARs) were considered for some ecological toxicity endpoints but could not be used to predict the chronic toxicity of acetonitrile to estuarine/marine fish or invertebrates. EPA’s Ecological Structural Activity Relationship Program (ECOSAR, 2000) was used to supplement available experimental data. ECOSAR was developed by the EPA Offic e of Pollution Prevention and Toxics and was used to predict the toxicity of acetonitrile based on structural similarities to chemicals for which aquatic toxicity has been measured experimentally

. ECOSAR

- derived acute aquatic toxicity values were generate d for estuarine/marine fish, mysid shrimp , and green algae. The predicted 96

- hour LC 50 and EC 50 values for these organisms ranged from 1 , 630 to 4 , 274 mg/L. The values of the parameters used for predicting the aquatic toxicity (as well as bioconcentration potential and environmental persistence) of acetonitrile are presented in Table 4

- 1. Table 4

- 2 presents the measured or predicted values indicating acute and chronic toxicity of acetonitrile to aquatic organisms and toxicity to terrestrial organisms. P r edicted toxicity values are presented in Table 4

- 2 primarily for those categories of toxicity for which no experimental data were available.
Table 4

- 1

. Values Used in ECOSAR/EPIWIN (U.S.EPA, 2000a) for Predicting the Toxicity, Bioconcentration Potential, and Persistence of Acetonitrile

Parameter Value Source

SMILES notation N#CC EPIWIN v. 3.12, US. EPA, 2000 a
Chemical name Acetonitrile Budavari et al., 1996

CAS number 75

- 05

- 8 Budavari et al., 1996

Molecular formula C 2 H 3 N Budavari et al., 1996 Molecular weight 41.05 Budavari et al., 199667 Parameter Value Source

Log K ow - 0.34 Hansch et al., 1995 [as cited in PHYSPROP , 2006 ]

Melting point

- 45.7 0 C McKinney and DeVito, 1998

Boiling point 81.6 0 C McKinney and DeVito, 1998 Henry ’ s Law constant 3.45x10 - 5 atm m 3 /mol at 25 0 C Gaffney et al., 1987 Vapor pressure 88.8 mm Hg at 25 0 C Boublik et al., 1984
Water solubility 1.0x10 6 mg/L at 25 0 C Riddick et al., 1986 [as cited in PHYSPROP]

Chemical class Neutral organics EPIWIN v. 3.12, US. EPA, 2000 a68 Table 4

- 2, Ecological Toxicity Values for Acetonitrile 2

Species Duration and test endpoint Experiment al type a,b Value (mg/L) c Reference Acute aquatic toxicity Algae Green algae predicted value 96

- hour EC 50 predicted 1 , 630 ECOSAR v. 0.99h, 2000 Freshwater invertebrate Water flea ( Daphnia magna ) 96

- hour LC 50 S, U >100 Ewell et al., 1986
Water flea ( Daphnia magna ) 48

- hour L C 50 R, U g 3 , 562 Tong and Hongjun, 2001 Water flea ( Daphnia magna ) 48

- hour LC 50 R, U g 3 , 600 Tong et al., 1996

Chironomus sp p
. 48

- hour L C 50 S, U g 3 , 165 Tong and Hongjun, 2001 Freshwater fish
Fathead minnow ( Pimephales promelas ) 96

- hour LC 50 F, M o 1 , 640 Brooke et al., 1984

Fathead minnow (Pimephales promelas) 96

- hour LC 50 S, U >100 Ewell et al., 1986
Fathead minnow ( P. promelas ) 96

- hour LC 50 d S, U e 1 , 000 Henderson et al., 1961 Fathead minnow ( P. promelas ) 96

- hour LC 50 d S, U f 1 , 000 Henderson et al., 1961

2 Many of the test concentrations were reported as nominal and thus table values could potentially underestimate toxicity as th ey are likely higher than the actual exposure concentrations. The EPIWIN model was therefore used to calculate a volatilization 1⁄2 life for lakes which was then used to estimate acetonitrile concentrations remaining in the static exposure systems for the nominal 100 mg/L treatments. The mo del estimated 1⁄2 life of 7.5 days indicates that approximately 69 mg/L would be expected to remai n at the end of the 96 hour static exposures. See section 2.6.1 Environmental Transport for the details of the Model Lake estimation.69 Species Duration and test endpoint Experiment al type a,b Value (mg/L) c Reference Bluegill ( Lepomis macrochirus ) 96

- hour LC 50 d S, U e 1 , 850 Henderson et al., 1961 Guppy ( Lebistes ) reticulatus ) 96

- hour LC 50 d S, U e 1 , 650 Henderson et al., 1961 Carp ( Cyprinus carpio ) 48

- hour LC 50 NS 730 r Nishiuchi, 1981 [as cited in IPCS, 1993] Carp ( C. carpio ) 96

- hour LC 50 R, U g 3 , 454 Tong and Hongjun, 2001
Estuarine/marine invertebrate Brine shrimp larvae ( Artemia salina ) 24

- hr LC 50 (at 72 hrs of larva l development) S, U s 399.65 Barahona

- Gomariz et al., 1994

Mysid shrimp 96

- hour LC 50 predicted 4 , 274 ECOSAR v. 0.99h, 2000 Estuarine/marine fish
Saltwater fish 96

- hour LC 50 predicted 259 ECOSAR v. 0.99h, 200070 Chronic aquatic toxicity Algae
Green algae ( Scenedesmus quadricauda ) 8

- day LOEC (cell multiplication inhibition) m S, U p 7 , 300 Bringmann and Kuhn, 1978a,b

Blue

- green algae ( Microcystis aeruginosa ) 8

- day LOEC (cell multiplication inhibition) m S, U p 520 Bringmann and Kuhn, 1978a,b Freshwater invertebrate Water flea ( Daphnia magna ) 21

- day LOEC (reproduction) R, U g 320 Tong et al., 1996 n ; Tong and Hongjun, 2001 21

- day NOEC (reproduction) 160

Water flea ( D. magna ) 21

- day LOEC (% survival) R, U g 640 h Tong et al., 1996

21

- day NOEC (% survival) 320 h

Water flea 16

- day EC 50 predicted 59 ECOSAR v. 0.99h, 2000 Freshwater fish Carp embryos ( Cyprinus carpio ) 7

- day NOEC R, NS g 400 Tong, 1999 7

- day LOEC 800 7

- day ChV 556 Freshwater fish 30

- day ChV predicted 291 ECOSAR v. 0.99h, 2000 Estuarine/marine invertebrate

- No data Estuarine/marine fish

- No data71 Terrestrial toxicity and phytotoxicity Avian species

- no data

Terrestrial invertebrate Earthworm (Eisenia foetida ) 14

- day LC 50 U >500 i,j Phillips et al., 1993
Earthworm ( E. andrei ) 15

- day LC 50 M >200 j,k,l Robidoux et al., 2000 Plants Duckweed ( Lemna minor ) 96

- hour NOEC (number of fronds and growth inhibition) t R, U q 1 , 000 Tong and Hongjun, 1997; 2001

Duckweed ( L. minor ) 96

- hour ChV (number of fronds and chlorophyll A concentration) R, U 1 , 342 Tong and Hongjun, 2001

Duckweed ( L. minor ) 96

- hour LOEC (number of fronds and growth inhibition) t R, U q 1 , 800 Tong and Hongjun, 1997; 2001

Duckweed ( L. minor ) 96

- hour L C 50 (growth inhibition) R, NS q 3 , 663, 3 , 685, 368 u Tong and Hongjun, 1997

a pH and water hardness were within accepted ranges (6

- 8.5 and 50

- 250 mg/L as CaCO 3 , respectively) unless noted otherwise

.

b Experiment type: S = static, R = renewal, F = flow

- through, M = measured, U = unmeasured, NS = not specified

.

c mg/L = ppm

.

d In the study report, the endpoint value was reported as a TL m (Median Tolerance Limit), which was defined as the concentration at which 50% of the fish were killed, which is equivalent to an LC 50 .

e Water hardness of the dilution water was reported as 20 ppm.72 f Water hardness of the dilution water was reported as 380 ppm.

g Water hardness of the dilution water was reportedly 1.86 + 0.08 mg/L as CaCO 3 .

h Note: the study (Tong et al., 1996) reported a 21

- day NOEC value of 640 mg/L for survival; however, this appears incorrect. Based on the percent survival data provided in the study report, the value of 640 mg/L (the highest concentration tested) appears to be the 21

- day LOEC and 320 mg/L appears to be the NOEC.

I μ g/g soil

.
j The study report did not explicitly report an LC 0 value, but reported that there was 100% survival at all test concentrations.

k μ L/g dry soil

.

l Note: Although the study report presented results based on nominal concentrations, the study report also indicated that test concentrations were measured over the course of the study using gas chromatography. The study reported that the nominal concentrat ion of 200 μ L/g decreased to 0.90 +/

- 0.26 u L/g after 15 days.

m The study reported a “toxic threshold” concentration (TT), described as “the pollutant concentration causing the onset of cell multiplication inhibition,” which is equivalent to a LOEC.

n The s tudy report by Tong et al. (1996), which reported results of both acute and chronic tests, had several inaccuracies: it reported the test type as both static and renewal; it reported that four pollutants were tested but presented results for only two; it r eported a water hardness outside of the preferred range of 50

- 250 mg/L as CaCO 3 ; standard deviations were not presented and there is no evidence in the study report that a statistical analysis was performed on the results; and the chronic NOEC and LOEC val ues for survival appear incorrectly reported in the study.

o Water hardness was reported as 43 mg/L as CaCO 3.

p Water hardness was not reported.

q Water hardness and pH were not reported in Tong & Hongjun, 1997.

r The 48

- hour LC 50 value of 730 mg/L cited in IPCS (1993) could not be verified because the original study report by Nishiuchi (1981) is in Japanese and no English translation was available.

s The pH of the synthetic seawater was reported as 8.6.

t In Tong and Hongjun , 2001 , the effects measured were number of fronds and chlorophyll concentration.

u Three different 96

- hour L C 50 values were reported in the study report.73 4.2.1. Acute Toxicity. Acute toxicity values are presented in Table 4

- 2. Studies in freshwater fish repo rt 96

- hour LC 50 values of >100 to 3 , 454 mg/L in the fathead minnow, bluegill, carp, and guppy, and a 48

- hour LC 50 value of 730 mg/L for carp. Experimental studies in freshwater invertebrates ( Daphnia magna ) reported 96

- and 48

- hour LC 50 values of >100 to 3 , 600 mg/L and a study in Chironomus sp . reported a 48

- hour EC 50 value of 3165 mg/L. ECOSAR predicted a 9 6

- hour EC 50 value of 1 , 630 mg/L for green algae. A study of an estuarine/marine invertebrate (brine shrimp larvae) reported a 24

- hour LC 50 value of 400 mg/L. ECOSAR predicted 96

- hour LC 50 values of 4 , 274 mg/L for mysid shrimp (an estuarine/marine invertebrate) and 259 mg/L for estuarine/marine fish. 4.2.2. Chronic Toxicity. Chronic toxicity values are presented in Table 4

- 2. The only avai lable chronic freshwater fish study was a 7

- day static

- renewal embryo

- larval study using 36

- hour

- old carp eggs that reported 7

- day NOEC and LOEC values (survival endpoint) of 400 and 800 mg/L respectively. The study also reported a 7

- day chronic value of 556 mg/L. ECOSAR predicted a 30

- day chronic value (survival/growth) of 291 mg/L for freshwater fish. ECOSAR predicted a 16

- day EC 50 of 59 mg/L for Daphnia sp . Ecotox icological 8

- day static threshold concentrations (TTs; equivalent to LOECs) of 7 , 300 mg/ L and 520 mg/L were reported for green and blue

- green algae, respectively. Additionally, a study in freshwater invertebrates ( D. magna ) reported a 21

- day NOEC of 160 mg/L and a 21

- day LOEC of 320 mg/L for reproduction, and a 21

- day NOEC of 320 mg/L and a 21

- day LOEC of 640 mg/L for survival (Tong et al., 1996; Tong and Hongjun, 2001). These data are of questionable reliability for several reasons. First, the NOEC and LOEC values for survival appear to be incorrectly reported in the study. The study re ported 21

- day NOEC and LOEC values of 640 and >1 , 300 mg/L, respectively. However, raw data on percent survival presented in Table 4

- 2 of the study report suggest that the correct 21

- day NOEC and LOEC values for survival are 320 and 640 mg/L. Second, 1 , 30 0 mg/L was not one of the tested concentrations , so it is unclear how the authors determined a 21

- day LOEC of >1 , 300 mg/L for survival. Third, standard deviations were not presented. Fourth, the study reported that four pollutants were tested, but presented results for only two. T he study reported water hardness outside of the preferred range of 50 to 250 mg/L as CaCO 3 . Fifth , it appears that nominal concentrations were used. Pref erence is given to actual m easured concentrations.74 There were no measured or predicted chronic estuarine/marine fish or invertebrate toxicity data available. 4.2.3. Terrestrial Toxicity and Phytotoxicity. Ecological toxicity values are presented in Table 4

- 2. Studies using earthworms reported 100% survival at exposure concentrations of 500 μ g/g soil for 14 days and 200 μ L/g dry soil for 15 days (Phillips et al., 1993 and Robidoux et al., 2000, respectively). Phytotoxicity studies in duckweed reported 96

- hour NOEC and LOEC values (for growth inhibition, chlorophyll A concentration, and number of fronds) of 1 , 000 and 1 , 800 mg/L, respectively; the 96

- hour ChV was 1 , 342 mg/L. A study o n duckweed also reported 96

- hour L C 50 values of 3 , 663, 3 , 685, and 368 m g/L for growth inhibition. 4.2.4. Population Effects. No data on population effects were located.
4.2.5. Bioaccumulation Potential. Bioaccumulation in aquatic organisms can be estimated using the bioconcentration factor (BCF), which describes the degree to which a chemical can be concentrated in the tissues of an organism. Measured BCFs were not available; thus, a BCF was calculated using EPIWIN v. 3.12 (U.S. EPA, 2000 a ). Based on a calculated BCF of 3, acetonitrile is not expected to bioaccumulate in t he tissue of aquatic organisms.75 4.3. REFERENCES
Barahona

- Gomariz, M.V., F. Sanz

- Barrera and S. Sanchez

- Fortun. 1994. Acute toxicity of organic solvents on Artemia salina . Bull. Environ. Contam. Toxicol. 52(5):766

- 771. Boublik, T., V. Fried and E. Hala. 1984. The vapor pressures of pure substances: Selected values of the temperature dependence of the vapor pressures of some pure substances in the normal and low pressure region. Vol. 17. Amsterdam, Netherlands. Elsevier Scientific Publications , as cited in PHYSPROP, 2006] Bringmann, G. and R. Kuhn. 1978a. Threshold values of substances harmful to water for blue algae ( Microcystis aeruginosa ) and green algae ( Scenedesmus quadricauda ) in tests measuring the inhibition of cellular propagation. Vo m. Wasser. 50:45

- 60. (In German) Bringmann, G. and R. Kuhn. 1978b. Testing of substances for their toxicity threshold: Model organisms Microcystis (Diplocystis) aeruginosa and Scenedesmus quadricauda . Mitt. Int. Ver. Theor. Angew. Limnol. 21:275

- 284

. Brooke, L.T., D.J. Call, D.L. Geiger et al. 1984. Acute toxicities of organic chemicals to fathead minnows ( Pimephales promelas ), Vol. 1. Center for Lake Superior Environmental Stud., Univ. of Wisconsin

- Superior, Superior, WI. pp. 414. Budavari, S., M.J. O’Neil, A. Smith et al. (eds). 1996. The Merck Index, 12th edition. Whitehouse NJ, Merck & Co., Inc.
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