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, 2013
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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 humans 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 an 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.
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TABLE OF CONTENTS

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

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

1.2 CHEMISTRY AND ENVIRONMENTAL FATE ASSESSMENT................................... 1

1.3 HUMAN HEALTH HAZARD ASSESSMENT CONCLUSIONS.................................... 3

1.4 ECOLOGICAL TOXICITY HAZARD ASSESSMENT CONCLUSIONS..................... 4

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

2. CHEMISTRY AND ENVIRONMENTAL FATE................................................................ 7

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

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

2.3 PHYSICAL AND CHEMICAL PROPERTIES ................................................................. 9

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

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

2.6 ENVIRONMENTAL FATE ............................................................................................... 12

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

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

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

3.2 KINETICS AND METABOLISM ..................................................................................... 29

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

3.4 EFFECTS OF REPEATED EXPOSURE ......................................................................... 34
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3.5 DEVELOPMENTAL AND REPRODUCTIVE TOXICITY .......................................... 43

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

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

3.8 POTENTIAL MODES OF ACTION................................................................................. 51

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

3.10 HUMAN RELEVANCE OF ANIMAL 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.................................................................................................................... 75
1

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 Community 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 acetonitrile 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 reported in the National Toxicology Program (NTP)

(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 ENVIRONMENTAL FATE ASSESSMENT

Acetonitrile (CAS Registry No. 75-05-8), also known as methyl cyanide, is a colorless, 

volatile, flammable organic liquid (Budavari, et. al., 1996). Its molecular formula is C2H3N, 

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. productions levels reported to EPA were 

between 50 million to less than 100 million pounds per year (U.S. EPA, 2011a). 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, 2011a). In 2009, releases of acetonitrile reported 

to the Toxic Release Inventory (TRI) totaled 17.7 million pounds, of which 17.4 million pounds 
2

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.

Acetonitrile 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 predictions 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, 116.7

and 3.5 percent, respectively (U.S. EPA, 2011b). 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., 2004). Based 

upon its 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.

Once in contact with underground water, acetonitrile will readily disperse throughout the water 

body. 

The most significant degradation process for acetonitrile is biodegradation; neither 

hydrolysis nor photodegradation is expected to be an important degradation process for 

acetonitrile (Brown et al., 1975, Atkinson, 1994). Observed biodegradation rates of acetonitrile

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 sections 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. 
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1.3 HUMAN HEALTH HAZARD ASSESSMENT CONCLUSIONS

In review of the potential human toxicity of acetonitrile, 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 long-term repeated 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 chemicals. 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 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 

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 were 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 acetonitrileinduced 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 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 acetonitrile exposures which 

are sufficient to cause mortality.
4

In addition to the concern for increased mortality 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 chronic 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 1,827 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 organism (U.S. EPA, 

1991). 

1.4 ECOLOGICAL TOXICITY HAZARD ASSESSMENT CONCLUSIONS

Based upon a review of the limited acute and chronic toxicity data available for aquatic 

species (algae, invertebrates, and vertebrates), adverse environmental effects due 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 freshwater fish and aquatic invertebrates ranged from 100 to 3,600 mg/L. The 

96 hour LC50 values for fish and green algae ranged from 100 to 3454 mg/L (Table 4-2). 

Experimental 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. Experimental chronic toxicity studies on earthworms reported 

100% survival at exposure concentrations of 500 micrograms per gram (μg/g) soil and 200 

microliters per gram (μL/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 acetonitrile to terrestrial species are limited to 

experimental chronic toxicity values for earthworms. These studies reported 100% survival at 

exposure concentrations of 500 μg/g soil for 14 days and 200 μL/g dry soil for 15 days (Phillips 

et al., 1993 and Robidoux et al., 2000, respectively). The phytotoxicity of acetonitrile has been 
5

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 toxicity. 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. NTIS 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 uptake 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. Nitriles. 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. Checkai 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. Svendsen, 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.) growth 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 protection 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 biodegradation 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. 2011. 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 2011a. Inventory Update Rule (IU) 2006. 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.

.
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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 C2H3N 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 pharmaceuticals, 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, 2011a). Of the millions of pounds of 

acetonitrile produced each year, a commercial market exists for about half of the inventory. In 

2009, releases of acetonitrile reported to the TRI totaled 17.7 million pounds, of which 17.4

million 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) will 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 (Koc) value of 5, acetonitrile may demonstrate 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 inoculum in an aqueous media with an initial 

100 ppm concentration of acetonitrile reported 40% degradation after10 days (Urano and Kato 

(1986a,b). The potential for anaerobic biodegradation, has been indicated by studies conducted 

by Li et al., (2007) although 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, based 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 photodegradation 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, cyanoMethanecarbonitrile

Methyl cyanide (MeCN)

NCI-C60822

NSC 7593

2.3 PHYSICAL AND CHEMICAL PROPERTIES

Table 2-1, Physical and Chemical Properties of Acetonitrile (CAS Reg. No. 75-05-8)

Property Valuea Conditions Reference

Molecular Formulae C2H3N

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) 1x106 25 °C Riddick et al., 1986

Log Kow -0.34 - Hansch et al., 1995
10

Henry’s Law constant (atmm3

/mol)

3.45x10-5 25 °C Gaffney et al., 1987

Dissociation constant, pKa -4.30 - Perrin, 1965

Hydroxyl radical reaction 

rate constant (cm3/molecsec)

2.63x10-14 25 °C Atkinson, 1994

Density (g/cm3) 0.786 20 °C McKinney and DeVito, 

1998

Vapor density 1.42 Air = 1 Clayton and Clayton, 1982

Koc 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:

 

Acetonitrile and hydrogen cyanide are the principal by-products of this reaction. 

Approximately 30-40 kilograms (kg) of acetonitrile and 140-180 kg of hydrogen cyanide are 

generated per 1000 kg 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 and DeVito, 1998). Some acrylonitrile producers 

H2C CH CH3 + NH3 + 3/2O2

catalyst H2C CH C N + 3H2O

byproducts: H3C C N + HCN

propylene acrylonitrile

acetonitrile
11

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, 2011). 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, 2011a). Of the millions of pounds of acetonitrile produced 

each year, a commercial market exists for about half of the inventory. In 2009, releases of 

acetonitrile reported to the TRI 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 presence 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., 2004).

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 C4 hydrocarbons by extractive distillation. It also has 

been proposed 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 used 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 extraction 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 (McKinney 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 molecular 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 chloride (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 released 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-m3

/mol (Lyman et al., 1990; Gaffney et al., 1987) using the EPIWIN estimation 

software (U.S. EPA2011b), the volatilization half-lives of acetonitrile are calculated to be 

approximately 0.5 days for a model river (assuming 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 atmosphere 
13

(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., 

2004). 

A soil partitioning coefficient (Koc) value of 5 has been calculated for acetonitrile using 

its experimental log Kow value of -0.34 (Hansch et al., 1995) and a regression-derived equation 

(HSDB, 2002; Lyman et al., 1990). Based on this Koc 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 sources 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 relatively 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 water body is dependent upon multiple 

factors including available air head space above the groundwater body, air exchange rates, and 

groundwater pH, temperature and velocity. Acetonitrile 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 instances where acetonitrile contaminated groundwater makes it way 

to surface water, it potentially 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. Acetonitrile does not undergo direct photodegradation (HSDB, 

2002). The 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 Bank (HSDB, 2002) reports an estimated half-life of 620 

days based upon an experimental rate constant for the reaction of acetonitrile with hydroxyl 

radicals of 2.63E-14 cubic centimeter (cm3)/molecule-sec and an atmospheric concentration of 

5x10+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 hydrocarbon 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 5.8E-3/M-hr and a corresponding half-life at pH 7 of 

150,000 years.

As stated earlier in this report, in 2009 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, 

2011). 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 membranebound organelles), and eukaryotes (organisms whose cells are internally organized into complex 
15

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 conditions 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 Mechanisms 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 unknown. 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

H3C C N

H2O

Nitrilase

H3C

OH

O

NH3 +

H3C C N

H2O

Nitrile hydratase

H3C

NH2

O H2O

Amidase

H3C

OH

O

NH3 +
16

Since acetamide is not formed 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 upon several factors including media, temperature, available 

oxygen, acetonitrile concentration, specific microorganisms involved and the degree of 

acclimation. For acetonitrile, the 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, was 

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 

proceeds 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 sludge. 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 initial 
17

concentration of 100 ppm was observed in 28 days using an activated sludge inoculum (OECD 

Test Method 301C).

The biodegradability of acetonitrile was also observed with river water; the 12-day ThOD 

(theoretical oxygen demand) test measured the generation of CO2 from Ohio River water spiked 

with acetonitrile at concentrations ranging from 0.1-25 mg/L. After 12 days, a total generation 

of CO2 was observed equivalent 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 concentrations

of acetonitrile by a mixed culture of microorganisms obtained from raw activated sewage sludge. 

It 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 used was a synthetic mineral 

salt solution. The microorganisms used in the study were acclimated with acetonitrile. As 

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% degradation after 186 hours (Li et 

al., 2007).

Concerning media type and its impact on acetonitrile degradation, based upon studies 

conducted by Takemoto et al., (1981) the 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. It 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 
18

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 aerobic 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 biodegradation rates were generally slower than aerobic 

biodegradation rates, with the exception of some chemical classes (e.g., nitroaromatic, 

halogenated hydrocarbons, but not nitriles). At 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). Acetamide, acetic acid, and ammonia were 

detected during the anaerobic biodegradation process which suggests that these products may be 

possible 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 solution. The 83% reduction of 

acetonitrile was observed after a 90-day incubation period at 37o

C using an initial acetonitrile 

concentration of 50 mg/L.

2.6.3.4 Biodegradation Organisms. Several microorganisms have been identified as 

demonstrating the ability 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 capable 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 Candida 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 concentrations 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 Distribution Assessment. Table 2-2 presents a model (U.S. EPA, 2011b) 

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 981.2, 

18.2 and 0.08 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 TRI by 

U.S. manufacturers in 2009 (U.S. EPA 2011). The model estimate (EPI Suite version 4.0, (U.S. 

EPA 2011b)) 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 environmental fate processes. These 

half-life values were also generated by the model (U.S. EPA, 2011b). 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 1.54x103

Water 16.7 360

Soil 3.5 720

Sediment 0.0329 3.24x103

Source: EPI Suite, (U.S. EPA, 2011b)

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 TRI (U.S. EPA, 2011), 17.7 million

pounds were released in 2009, of which 17.4 million pounds were disposed of on-site in 

underground injection wells. Based upon the 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. 

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 generated 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

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 5.8E-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 of Medicine’s Hazardous Substance Data Bank (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) using 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, reported acetamide 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 onsite for energy recovery by the manufacturers (U.S.EPA 2011). 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 
22

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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27

CHAPTER THREE: HUMAN HEALTH HAZARD ASSESSMENT

3.1 SUMMARY

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 acetonitrile 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 body 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 observed.

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 505 mg/kg-day) and in 

mice at 800 ppm (calculates to approximately 402 mg/kg-day) (NTP, 1996). Effects were not 

observed in rats or mice following chronic exposure to 400 ppm (calculates to approximately 126 
28

mg/kg-day) acetonitrile in rats and 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 carcinogenicity 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 result 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 mortality 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 

(Johannsen et al., 1986). An 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 maternal 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 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). 

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 KINETICS AND METABOLISM

3.2.1. Absorption. Human 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/m3

) for 4 hours. Resulting blood cyanide levels peaked 

and reached 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 acetonitrile radiolabeled with 14C 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). 

Additionally, 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 pathway 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 NADPHdependent, and mediated 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 is 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 sulfurcontaining donor molecules, such as thiosulphate, is the rate-limiting factor for the conversion of 

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 acid (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 cavity (U.S. EPA, 1999). The nasal cavity of the rat, in particular 

CH3CN CYP2E1 CH2OHCN

CH2O

HCN SCNHCOOH

CNacetonitrile methyl cyanohydrin

formaldehyde

sulfur donor

(thiosulfate)

hydrogen cyanide

rhodanese

thiocyanate

cyanate

formic acid

2-aminothiolazine-4-carboxylic acid
32

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, 

Lewis et al. (1991) showed that the rhodanese in human nasal tissue had a lower affinity (higher 

Km) for cyanide and a lower rate (lower Vmax) for cyanide metabolism than rhodanese from rat 

nasal tissue. K(m) is the concentration of substrate which permits the 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, such 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 contrast 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 Km and Vmax 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 chronic 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. Approximately 80% of an ingested dose of inorganic cyanide is converted 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, 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. It should be noted that this may 
33

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 

detected 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 occupational 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 congestion (WHO, 1993). Gastric erosion has been reported in individuals who ingested 

acetonitrile (Ballantyne, 1983; Way, 1981).

In animals, oral LD50 values have been reported for the mouse (269-453 mg/kg) and the 

rat (1,730-4,050 mg/kg) and inhalation LC50 values 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 LC50 in mice for several 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 LD50 of 617 mg/kg for Crl:CD-1(ICR)BR mice and an inhalation LC50 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, 

hematology, 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-day2. Statistically significant decreases in red and white blood cell 

counts, segmented neutrophil and monocyte counts, hematocrit, 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 (inhalation 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 (m3

/day) 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 m3

/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 m3 / day x (6hr / 24hr) x 

(5 days / 7days) ÷ 0.020 kg = 49.9 mg/kg-day
36

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, 1983a, 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 were 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. An increase in 

liver/brain 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 subchronic 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 505 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 erythrocytes 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, including, 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 mice
38

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 

6-13a

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 

female 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” (NTP 1996). The NTP (1996) report 

states that, “the dose response for 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 observed 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 mice 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 statistically 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 carcinogenicity 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 acetonitrile (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/kgday) and B6C3F1 mice exposed to concentrations 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 acetonitrile. 

Histopathological examination did not reveal neoplastic or nonneoplastic lesions in any organs of 
41

exposed females. In male rats, a statistically significant increase in the incidence of basophilic 

hepatic foci was observed in the 200 and 400 ppm groups (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 rats, 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 the 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 observed, 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 groups, 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). Combined 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 Table 3-3). Although there was a statistically significant 

positive trend in the incidences of hepatocellular adenomas, carcinomas, 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 neoplastic finding in rats was a marginal increase in the incidence of hepatocellular 

adenoma or carcinoma (combined) in male rats exposed to 400 ppm. Statistical significance was 
42

achieved (P=0.045) for a dose-related trend, but pairwise comparisons of the exposed group 

incidences versus the control incidence were not significant. In 400 ppm males, the incidence of 

hepatocellular 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

Eosinopilic 

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 only 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 spontaneous 

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 respiratory 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 acetonitrile 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 

identified 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 

toxicity 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. The 
45

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 (wholebody) 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 included 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 and 

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 mononitriles, 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 

1,827ppm (calculates to approximately 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 

highest 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 1,827ppm, along with a decrease (not significant) 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 maternal 
46

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 rats (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 midgestation, and at 500 mg/kg/day, 16 females died prior to termination (>50% 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 offspring 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 maternal 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 ≥750 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 developmental 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 non-viable fetuses, number of early and late resorptions, number of total 

implantations and number of corpora lutea. At the highest dose of 275 mg/kg-day, two females 
47

died by the end of the study (8% mortality) and four others were described in the study report 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, 15, 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 reported 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 gastrointestinal 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 dose 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 acetonitrile.
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 /ReproductionEffects

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, SpragueDawley 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 SpragueDawley 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 LongEvans 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

LOAEL: 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

Decreased number live fetuses/litter
49

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 aliphatic 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 circulation 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 differentiation, 

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 system, 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 dropped 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 teratogenic 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 exchanges (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, chromosomal 
50

aberrations in CHO cells were increased 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 erythrocytes 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 aneuploidy 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 data 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. EPA, 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 group 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 have 
51

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 evidence 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 trivalent 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 inorganic 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 the 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 critical 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 over 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 (1996) 

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 of 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 THYROID 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 the 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 

inhibition of iodide uptake, and thus decreasing 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 production 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 administered 6 hours/day, 5 days/week, temporary 
53

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 adequate thyroid hormone (T3 and T4) 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 development 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 neuronal 

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 RELEVANCE 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 equivalent modes of action for acetonitrile 

lethality upon acute exposure for test animals and humans, and by extension suggests relevance 
54

of acetonitrile lethality due to chronic exposures in humans, as seen in animals from the NTP 

(1996) subchronic mouse inhalation 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 (Km) and metabolic rate (Vmax) for cyanide metabolism. These toxicokinetic parameters 

dictate the ability of species to detoxify acetonitrile and its metabolites, particularly cyanide 

moieties. Variability 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 thiocyanate (rhodanese). A study comparing the rhodanese activity of rats and humans 

observed that the respiratory epithelium of human nasal tissue had a lower affinity (Km) for 

cyanide and a lower rate (Vmax) 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 enzyme (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 

of 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 exposure 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 occur 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 acetonitrile
55

(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 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 death in acute toxicity studies (e.g., LC50 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 indicate 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, mortality 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 mortality is dissimilar to that observed in the 1,600 ppm treatment group 
56

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 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 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 developmental 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 developing 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

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cigarette smoke. Arch Environ Health 16:831-835, 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 17:746-748, as cited in U.S. EPA, 1999.

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oxidation of acetonitrile to cyanide. Chem. Res. Toxicol. 2:359-66, [as cited in U.S. EPA, 1999]

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J. Toxicol. Environ. Health. 15:609-622.

Freeman, J.J. and E.P. Hayes. 1988. Microsomal metabolism of acetonitrile to cyanide. 

Biochem. Pharmacol. 37:1153-1159.

Haquenoer, JM; Dequidt, J; Jacquemont, MC. (1975) Intoxications experimentales 

parl’acetonitrile 2eme note: intoxications aiques par voie pulmonaire. Eur J. Toxicol 8:102-106,

as cited in U.S. EPA, 1999.

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Rev. Ed. Patty, F.A., G.D. Clayton, F.E. Clayton et al., eds. New York: Wiley. pp. 4845-4900.
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Hazelton Laboratories. 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-0284-0292.

Johannsen, F.R., G.J. Levinskas, P.E. Berteau et al. 1986. Evaluation of the teratogenic 

potential of three aliphatic nitriles in the rat. Fundam. Appl. Toxicol. 7:33-40.

Jones, E.V., V. Fox, B.M. Elliot et al. 2001. The mutagenic potential of acetonitrile in the bone 

marrow and peripheral blood of the mouse. Mutagen. 16:151-154.

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McGraw Hill, seventh edition, pp. 73-94, 829.

Lewis, J.L., C.E. Rhoades, P-G. Gervasi et al. 1991. The cyanide-metabolizing enzyme 

rhodanese in human nasal respiratory mucosa. Toxicol. Appl. Pharmacol. 108:114-120.

Lu Y. and A.I. Cederbaum. 2008. CYP2E1 and oxidative liver injury by alcohol. Free Radic Biol 

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studies: Acetonitrile in rats. Battelle Laboratory for NIEHS, NTP. PNL-940. 

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carcinogenicity of 1,3-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 

biotransformation 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.

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62

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 

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development. Endocr Rev 18(4):462-75.

Osgood, C., M. Bloomfield and S. Zimmering. 1991a. Aneuploidy in Drosophila, IV. 

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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.

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inhaled aliphatic mononitriles in rats. Fundam. Appl. Toxicol. 20:365-375 as cited in U.S. EPA, 

1999. 

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nitriles: In vivo and in vitro observations. Toxicol. Appl. Pharmacol. 163:149-163.

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fumaronitrile in three short term assays with special reference to autoinduction. J. Appl. Toxicol. 

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63

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Pollution Prevention and Toxics, Washington, DC.

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64

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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 due 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 toxicity (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 

LC50 values for fish and green algae ranged from 100 to 3,454 mg/L. 

Experimental 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. Experimental chronic toxicity studies on earthworms reported 

100% survival at exposure concentrations of 500 μg/g soil and 200 μL/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 Kow: -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 database 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 they 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 model 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. See section 2.6.1 

Environmental Transport for the details of the Model Lake estimation.
66

toxicity were found for acetonitrile. Limited 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 Office 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 generated for estuarine/marine fish, mysid 

shrimp, and green algae. The predicted 96-hour LC50 and EC50 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. 

Predicted 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, 

2000a

Chemical name Acetonitrile Budavari et al., 1996

CAS number 75-05-8 Budavari et al., 1996

Molecular formula C2H3N Budavari et al., 1996

Molecular weight 41.05 Budavari et al., 1996
67

Parameter Value Source

Log Kow -0.34 Hansch et al., 1995 [as cited 

in PHYSPROP, 2006]

Melting point -45.7 0C

McKinney and DeVito, 

1998

Boiling point 81.6 0C

McKinney and DeVito, 

1998

Henry’s Law constant 3.45x10-5 atm m3/mol at 25 0C Gaffney et al., 1987

Vapor pressure 88.8 mm Hg at 25 0C Boublik et al., 1984

Water solubility 1.0x106 mg/L at 25 0C Riddick et al., 1986 [as 

cited in PHYSPROP]

Chemical class

Neutral organics EPIWIN v. 3.12, US. EPA, 

2000a
68

Table 4-2, Ecological Toxicity Values for Acetonitrile2

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 EC50 predicted 1,630 ECOSAR v. 0.99h, 

2000

Freshwater invertebrate

Water flea (Daphnia 

magna)

96-hour LC50 S, U >100 Ewell et al., 1986

Water flea (Daphnia 

magna)

48-hour LC50 R, U g 3,562 Tong and Hongjun, 

2001

Water flea (Daphnia 

magna)

48-hour LC50 R, U g 3,600 Tong et al., 1996

Chironomus spp. 48-hour LC50 S, U g 3,165 Tong and Hongjun, 

2001

Freshwater fish

Fathead minnow 

(Pimephales promelas)

96-hour LC50 F, Mo 1,640 Brooke et al., 1984

Fathead minnow 

(Pimephales promelas)

96-hour LC50 S, U >100 Ewell et al., 1986

Fathead minnow (P. 

promelas)

96-hour LC50

d S, Ue 1,000 Henderson et al., 

1961

Fathead minnow (P. 

promelas)

96-hour LC50

d S, Uf 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 they 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 model estimated 1⁄2 life 

of 7.5 days indicates that approximately 69 mg/L would be expected to remain 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 LC50

d S, Ue 1,850 Henderson et al., 

1961

Guppy (Lebistes)

reticulatus)

96-hour LC50

d S, Ue 1,650 Henderson et al., 

1961

Carp (Cyprinus carpio) 48-hour LC50 NS 730r Nishiuchi, 1981 [as 

cited in IPCS, 1993] 

Carp (C. carpio) 96-hour LC50 R, U g 3,454 Tong and Hongjun, 

2001

Estuarine/marine invertebrate

Brine shrimp larvae 

(Artemia salina)

24-hr LC50

(at 72 hrs of larval 

development)

S, Us 399.65 Barahona-Gomariz et 

al., 1994

Mysid shrimp 96-hour LC50 predicted 4,274 ECOSAR v. 0.99h, 

2000

Estuarine/marine fish

Saltwater fish 96-hour LC50 predicted 259 ECOSAR v. 0.99h, 

2000
70

Chronic aquatic toxicity

Algae

Green algae 

(Scenedesmus 

quadricauda)

8-day LOEC (cell 

multiplication 

inhibition)m

S, Up 7,300 Bringmann and Kuhn, 

1978a,b

Blue-green algae 

(Microcystis 

aeruginosa)

8-day LOEC (cell 

multiplication 

inhibition)m

S, Up 520

Bringmann and Kuhn, 

1978a,b

Freshwater invertebrate

Water flea (Daphnia 

magna)

21-day LOEC 

(reproduction)

R, U g

320 Tong et al., 1996n

; 

Tong and Hongjun, 

2001

21-day NOEC 

(reproduction)

160

Water flea (D. magna) 21-day LOEC (% 

survival)

R, U g 640h Tong et al., 1996

21-day NOEC (% 

survival)

320h

Water flea 16-day EC50 predicted 59 ECOSAR v. 0.99h, 

2000

Freshwater fish

Carp embryos (Cyprinus 

carpio)

7-day NOEC R, NSg 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 data
71

Terrestrial toxicity and phytotoxicity

Avian species - no data

Terrestrial invertebrate

Earthworm (Eisenia 

foetida)

14-day LC50 U >500i,j Phillips et al., 1993

Earthworm (E. andrei) 15-day LC50 M >200j,k,l Robidoux et al., 2000

Plants

Duckweed (Lemna 

minor)

96-hour NOEC 

(number of fronds 

and growth 

inhibition)t

R, Uq 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, Uq 1,800 Tong and Hongjun, 

1997; 2001

Duckweed (L. minor) 96-hour LC50

(growth 

inhibition)

R, NSq 3,663, 

3,685, 

368u

Tong and Hongjun, 

1997

a pH and water hardness were within accepted ranges (6-8.5 and 50-250 mg/L as CaCO3, 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 TLm (Median Tolerance Limit), which was 

defined as the concentration at which 50% of the fish were killed, which is equivalent to an LC50.

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.

gWater hardness of the dilution water was reportedly 1.86+0.08 mg/L as CaCO3.

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 

μg/g soil.

j

The study report did not explicitly report an LC0 value, but reported that there was 100% survival at all 

test concentrations.

k μL/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 concentration of 200 μL/g decreased to 0.90+/-0.26 uL/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 study 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 reported a water hardness outside of the preferred range of 50-250 

mg/L as CaCO3; 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 values for survival 

appear incorrectly reported in the study.

o Water hardness was reported as 43 mg/L as CaCO3.

p Water hardness was not reported.

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

r

The 48-hour LC50 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 LC50 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 report 96-hour LC50 values of >100 to 3,454 mg/L in the fathead minnow, bluegill, carp, and 

guppy, and a 48-hour LC50 value of 730 mg/L for carp. Experimental studies in freshwater 

invertebrates (Daphnia magna) reported 96- and 48-hour LC50 values of >100 to 3,600 mg/L and 

a study in Chironomus sp. reported a 48-hour EC50 value of 3165 mg/L. ECOSAR predicted a 

96-hour EC50 value of 1,630 mg/L for green algae. A study of an estuarine/marine invertebrate 

(brine shrimp larvae) reported a 24-hour LC50 value of 400 mg/L. ECOSAR predicted 96-hour 

LC50 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 available 

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 EC50 of 59 mg/L for Daphnia sp. Ecotoxicological 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 reported 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, 1,300 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. The study reported water hardness 

outside of the preferred range of 50 to 250 mg/L as CaCO3. Fifth, it appears that nominal 

concentrations were used. Preference is given to actual measured 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 μg/g 

soil for 14 days and 200 μL/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 on duckweed also reported 

96-hour LC50 values of 3,663, 3,685, and 368 mg/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, 2000a). Based on a calculated BCF of 3, 

acetonitrile is not expected to bioaccumulate in the tissue of aquatic organisms.
75

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