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EPA/635/R-08/016
www.epa.gov/iris
oEPA
TOXICOLOGICAL REVIEW
OF
HYDROGEN CYANIDE
AND
CYANIDE SALTS
(CAS No. various)
In Support of Summary Information on the
Integrated Risk Information System (IRIS)
August 2009
NOTICE
This document is an External Peer Review draft. This information is distributed solely for the
purpose of pre-dissemination peer review under applicable information quality guidelines. It has
not been formally disseminated by EPA. It does not represent and should not be construed to
represent any Agency determination or policy. It is being circulated for review of its technical
accuracy and science policy implications.
U.S. Environmental Protection Agency
Washington, DC
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DISCLAIMER
This document is a preliminary draft for review purposes only. This information is
distributed solely for the purpose of pre-dissemination peer review under applicable information
quality guidelines. It has not been formally disseminated by EPA. It does not represent and
should not be construed to represent any Agency determination or policy. Mention of trade
names or commercial products does not constitute endorsement or recommendation for use.
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CONTENTS —TOXICOLOGICAL REVIEW OF HYDROGEN CYANIDE AND CYANIDE
SALTS
LIST OF TABLES v
LIST OF FIGURES vi
LIST OF ACRONYMS vii
FOREWORD ix
AUTHORS, CONTRIBUTORS, AND REVIEWERS x
1. INTRODUCTION 1
2. CHEMICAL AND PHYSICAL INFORMATION 3
3. TOXICOKINETICS 7
3.1. ABSORPTION 7
3.2. DISTRIBUTION 8
3.3. METABOLISM 11
3.4. ELIMINATION 16
3.5. PHYSIOLOGICALLY BASED TOXICOKINETIC MODELS 17
4. HAZARD IDENTIFICATION 19
4.1. STUDIES IN HUMANS—EPIDEMIOLOGY, CASE REPORTS, CLINICAL
CONTROLS 19
4.1.1. Acute Oral, Inhalation, and Dermal Studies 19
4.1.2. Subchronic and Chronic Oral Studies 21
4.1.3. Subchronic and Chronic Inhalation Studies 21
4.2. SUBCHRONIC AND CHRONIC STUDIES AND CANCER BIOASSAYS IN
ANIMALS—ORAL AM) INHALATION 27
4.2.1. Oral Studies 27
4.2.2. Inhalation Studies 38
4.3. REPRODUCTIVE/DEVELOPMENTAL STUDIES 39
4.3.1 Oral Studies 39
4.3.2 Inhalation Studies 42
4.4. OTHER DURATION- OR ENDPOINT-SPECIFIC STUDIES 43
4.4.1. Acute Oral Studies 43
4.4.2. Acute Inhalation Studies 45
4.4.3. Neurotoxicity Studies 46
4.4.4. Immune Endpoints 47
4.5. MECHANISTIC DATA AND OTHER STUDIES IN SUPPORT OF THE
MODE OF ACTION 48
4.5.1. Genotoxi city 48
4.5.2. Acute Neurotoxicity 49
4.5.3. Thyroid Disruption 49
4.5.4. Reproductive Effects 50
4.6. SYNTHESIS OF MAJOR NONCANCER EFFECTS AND MODE OF
ACTION 52
4.7. EVALUATION 01 CARCINOGENICITY 58
4.8. SUSCEPTIBLE POPULATIONS 58
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4.8.1. Possible Childhood Susceptibility 58
4.8.2. Possible Gender Differences 60
4.8.3. Other Susceptible Populations 60
5. DOSE RESPONSE ASSESSMENTS 62
5.1. ORAL REFERENCE DOSE (RID) 62
5.1.1. Choice of Principal Study and Critical Effect 62
5.1.2. Method of Analysis 66
5.1.3. RfD Derivation-Including Application of Uncertainty Factors (UFs) 68
5.1.4. RfD Comparison Information 70
5.1.5. Previous RfD Assessment 73
5.2. INHALATION REFERENCE CONCENTRATION (RfC) 73
5.2.1. Choice of Principal Study and Critical Effect 73
5.2.2. Method of Analysis 76
5.2.3. RfC Derivation—Including Application of Uncertainty Factors (UFs) 77
5.2.4. Previous RfC Assessment 78
5.3. UNCERTAINTIES IN THE ORAL REFERENCE DOSE AND
INHALATION REFERENCE CONCENTRATION 78
5.4. CANCER ASSESSMENT 84
6. MAJOR CONCLUSIONS IN CHARACTERIZATION OF HAZARD
AND DOSE RESPONSE 85
6.1. HUMAN HAZARD POTENTIAL 85
6.2. DOSE RESPONSE 87
6.2.1. Noncancer—Oral 87
6.2.2. Noncancer—Inhalation 89
6.2.3. Cancer 90
7. REFERENCES 91
APPENDIX A. SUMMARY OF EXTERNAL PEER REVIEW AND PUBLIC
COMMENTS AND DISPOSITION A-l
APPENDIX B. BENCHMARK DOSE MODELING RESULTS B-l
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LIST OF TABLES
Table 2-1. Physical and chemical properties of cyanide compounds 5
Table 4-1. Thyroid uptake of 13II in electroplating workers 23
Table 4-2. Thyroid parameters in former silver-reclaiming workers 25
Table 4-3. Thyroid parameters in HCN-exposed and unexposed electroplating workers 26
Table 4-4. Reproductive effects in male rats administered NaCN in drinking water for 13
weeks 30
Table 4-5. Reproductive effects in mice administered NaCN in drinking water for 13
weeks 32
Table 4-6. Summary of subchronic and chronic oral toxicity studies for cyanide in
animals 53
Table 4-7. Summary of inhalation toxicity studies for cyanide in humans 55
Table 5-1. Reproductive endpoints in male rats and mice observed following
administration of NaCN in drinking water for 13 weeks 66
Table 5-2. BMD modeling results for observed reproductive endpoints 67
Table 5-3. Alternate PODs with applied UFs and resulting potential RfVs 71
Table B-l. Decreased cauda epididymis weight in F344 rats following administration of
NaCN in drinking water for 13 weeks B-l
Table B-2. BMD modeling results for decreased cauda epididymis weight in rats B-l
Table B-3. Decreased epididymis weight in F344 rats following administration of NaCN
in drinking water for 13 weeks B-5
Table B-4. BMD modeling results for decreased epididymis weight in rats B-5
Table B-5. Decreased testis weight in F344 rats, following administration of NaCN in
drinking water for 13 weeks B-9
Table B-6. BMD modeling results for decreased testis weight in rats B-9
Table B-7. Decreased testicular spermatid concentration in F344 rats following
administration of NaCN in drinking water for 13 weeks B-13
Table B-8. BMD modeling results for decreased testicular spermatid concentration in
rats B-13
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LIST OF FIGURES
Figure 3-1. Cyanide primary metabolic pathways 12
Figure 4-1. Urinary SCN- of exposed workers plotted against individual breathing
concentrations of HCN 23
Figure 5-1. Potential RfV comparison array for alternate PODs 72
Figure B-l. Observed and predicted decrease in cauda epididymis weight in F344 rats
following administration of NaCN in drinking water for 13 weeks B-2
Figure B-2. Observed and predicted decrease in epididymis weight in F344 rats
following administration of NaCN in drinking water for 13 weeks B-6
Figure B-3. Observed and predicted decrease in epididymis weight in F344 rats
following administration of NaCN in drinking water for 13 weeks B-9
Figure B-4. Observed and predicted decrease in testicular spermatid concentration in
F344 rats following administration of NaCN in drinking water for 13 weeks B-13
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LIST OF ACRONYMS
ACH
acetone cyanohydrin
ADP
adenosine diphosphate
AIC
Akaike's Information Criterion
ALP
alkaline phosphatase
ALT
alanine aminotransferase
ANOVA
analysis of variance
AST
aspartate aminotransferase
ATP
adenosine triphosphate
ATSDR
Agency for Toxic Substances and Disease Registry
AUC
area under the curve
BMD
benchmark dose
BMDL
95% lower confidence limit on the benchmark dose
BMDS
benchmark dose software
BMR
benchmark response
CAP
compound action potential
CASRN
Chemical Abstracts Service Registry Number
CI
confidence interval
CN-
cyanide ion
(CN)2
cyanogen
CNS
central nervous system
EGL
external granular layer
FEV
forced expiratory volume
FVC
forced vital capacity
GD
gestation day
GFAP
glial fibrillary acid protein
GGT
y-glutamyl transferase
HCN
hydrogen cyanide
i.p.
intraperitoneal or intraperitoneally
IPCS
International Programme on Chemical Safety
IRIS
Integrated Risk Information System
i.v.
intravenous or intravenously
KCN
potassium cyanide
KSCN
potassium thiocyanate
LD50
median lethal dose
LDH
lactate dehydrogenase
LOAEL
lowest-observed-adverse-effect level
ML
molecular layer
MP ST
mercaptopyruvate sulfotransferase
NaCN
sodium cyanide
NIS
sodium-iodide symporter
NOAEL
no-ob served-adverse-effect level
NRC
National Research Council
NTP
National Toxicology Program
OOT
cyanate
OSOT
hypothiocyanate
PND
postnatal day
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POD
point of departure
RBC
red blood cell
RfC
reference concentration
RfD
reference dose
RfV
reference value
RR
relative risk
SCN~
thiocyanate
SD
standard deviation
SEM
standard error of the mean
t3
triiodothyronine
t4
thyroxine
TSH
thyroid-stimulating hormone
TWA
time-weighted average
UF
uncertainty factor
U.S. EPA
U.S. Environmental Protection Agency
V
* max
maximum velocity
WBC
white blood cell
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FOREWORD
The purpose of this Toxicological Review is to provide scientific support and rationale
for the hazard and dose-response assessment in IRIS pertaining to chronic exposure to hydrogen
cyanide and cyanide salts. It is not intended to be a comprehensive treatise on the chemical or
toxicological nature of hydrogen cyanide and cyanide salts.
The intent of Section 6, Major Conclusions in the Characterization of Hazard and Dose
Response, is to present the major conclusions reached in the derivation of the reference dose,
reference concentration and cancer assessment, where applicable, and to characterize the overall
confidence in the quantitative and qualitative aspects of hazard and dose response by addressing
the quality of data and related uncertainties. The discussion is intended to convey the limitations
of the assessment and to aid and guide the risk assessor in the ensuing steps of the risk
assessment process.
For other general information about this assessment or other questions relating to IRIS,
the reader is referred to EPA's IRIS Hotline at (202) 566-1676 (phone), (202) 566-1749 (fax), or
hotline.iris@epa.gov (email address).
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AUTHORS, CONTRIBUTORS, AND REVIEWERS
CHEMICAL MANAGER
Kathleen Newhouse, M.S.
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Washington, DC
AUTHORS
Kathleen Newhouse, M.S.
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Washington, DC
Nancy Chiu, Ph.D.
Office of Water/HECD
U.S. Environmental Protection Agency
Washington, DC
Lynne Haber, Ph.D., DABT
Toxicology Excellence for Risk Assessment
Joan Strawson, M.S.
Toxicology Excellence for Risk Assessment
Bonnie Ransom Stern, Ph.D., M.P.H.
Consulting in Health Sciences and Risk Assessment
BR Stern and Associates, ICF Consulting, Inc.
REVIEWERS
This document has been reviewed by EPA scientists and interagency reviewers from
other federal agencies.
INTERNAL EPA REVIEWERS
Jamie Strong, Ph.D.
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Washington, DC
Lynn Flowers, Ph.D.
National Center for Environmental Assessment
U.S. Environmental Protection Agency
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Washington, DC
Ted Berner, M.S.
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Washington, DC
Glinda Cooper, Ph.D.
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Washington, DC
Ralph Cooper, Ph.D.
National Health and Environmental Effects Research Laboratory
Office of Research and Development
Susan Griffin, Ph.D., DABT
Region VIII
Sally Perreault Darney, Ph.D.
National Health and Environmental Effects Research Laboratory
Office of Research and Development
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1. INTRODUCTION
This document presents background information and justification for the Integrated Risk
Information System (IRIS) Summary of the hazard and dose-response assessment of hydrogen
cyanide and cyanide salts. IRIS Summaries may include oral reference dose (RfD) and
inhalation reference concentration (RfC) values for chronic and other exposure durations, and a
carcinogenicity assessment.
The RfD and RfC, if derived, provide quantitative information for use in risk assessments
for health effects known or assumed to be produced through a nonlinear (presumed threshold)
mode of action. The RfD (expressed in units of mg/kg-day) is defined as an estimate (with
uncertainty spanning perhaps an order of magnitude) of a daily exposure to the human
population (including sensitive subgroups) that is likely to be without an appreciable risk of
deleterious effects during a lifetime. The inhalation RfC (expressed in units of mg/m3) is
analogous to the oral RfD, but provides a continuous inhalation exposure estimate. The
inhalation RfC considers toxic effects for both the respiratory system (portal of entry) and for
effects peripheral to the respiratory system (extrarespiratory or systemic effects). Reference
values are generally derived for chronic exposures (up to a lifetime), but may also be derived for
acute (< 24 hours), short-term (>24 hours up to 30 days), and subchronic (>30 days up to 10% of
lifetime) exposure durations, all of which are derived based on an assumption of continuous
exposure throughout the duration specified. Unless specified otherwise, the RfD and RfC are
derived for chronic exposure duration.
The carcinogenicity assessment provides information on the carcinogenic hazard
potential of the substance in question and quantitative estimates of risk from oral and inhalation
exposure may be derived. The information includes a weight-of-evidence judgment of the
likelihood that the agent is a human carcinogen and the conditions under which the carcinogenic
effects may be expressed. Quantitative risk estimates may be derived from the application of a
low-dose extrapolation procedure. If derived, the oral slope factor is a plausible upper bound on
the estimate of risk per mg/kg-day of oral exposure. Similarly, an inhalation unit risk is a
plausible upper bound on the estimate of risk per |ig/m3 air breathed.
Development of these hazard identification and dose-response assessments for hydrogen
cyanide and cyanide salts has followed the general guidelines for risk assessment as set forth by
the National Research Council (1983). EPA Guidelines and Risk Assessment Forum Technical
Panel Reports that may have been used in the development of this assessment include the
following: Guidelines for the Health Risk Assessment of Chemical Mixtures (U.S. EPA, 1986a),
Guidelines for Mutagenicity Risk Assessment (U.S. EPA, 1986b), Guidelines for Developmental
Toxicity Risk Assessment (U.S. EPA, 1991), Guidelines for Reproductive Toxicity Risk
Assessment {U.S. EPA, 1996), Guidelines for Neurotoxicity Risk Assessment (U .S. EPA, 1998a),
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Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a), Supplemental Guidance for
Assessing Susceptibility from Early-Life Exposure to Carcinogens (U.S. EPA, 2005b),
Recommendations for and Documentation of Biological Values for Use in Risk Assessment (U.S.
EPA, 1988), (proposed) Interim Policy for Particle Size and Limit Concentration Issues in
Inhalation Toxicity (U.S. EPA, 1994a), Methods for Derivation of Inhalation Reference
Concentrations and Application of Inhalation Dosimetry (U.S. EPA, 1994b), Use of the
Benchmark Dose Approach in Health Risk Assessment (U.S. EPA, 1995), Science Policy Council
Handbook: Peer Review (U.S. EPA, 2006a), Science Policy Council Handbook: Risk
Characterization (U.S. EPA, 2000b), Benchmark Dose Technical Guidance Document (U.S.
EPA, 2000c), Supplementary Guidance for Conducting Health Risk Assessment of Chemical
Mixtures (U.S. EPA, 2000d), A Review of the Reference Dose and Reference Concentration
Processes (U.S. EPA, 2002),and A Framework for Assessing Health Risks of Environmental
Exposures to Children (U.S. EPA, 2006b).
The literature search strategy employed for this compound was based on the Chemical
Abstracts Service Registry Numbers (CASRNs) and at least one common name. Any pertinent
scientific information submitted by the public to the IRIS Submission Desk was also considered
in the development of this document. The relevant literature was reviewed through
December 2008.
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2. CHEMICAL AND PHYSICAL INFORMATION
The term cyanide refers to any compound that contains the cyanide ion (CIST), consisting
of a carbon atom triple bonded to a nitrogen atom. Hydrogen cyanide (HCN) is a colorless or
pale blue liquid or gas with a faint bitter almond-like odor, while sodium cyanide (NaCN) and
potassium cyanide (KCN) are white crystalline powders. HCN is a weak acid with a pKa of 9.2;
therefore, HCN and CN can interconvert based on pH and temperature. In solution under
physiological conditions, the majority of HCN is present in the undissociated form. The simple
cyanide salts, KCN and NaCN, are very soluble in water and mildly soluble in ethanol. These
compounds readily dissociate in water, and so exposure to any of these compounds in aqueous
media results in exposure to CIST. For the sake of comparability, doses in this review are given
as cyanide (GST) unless stated otherwise. Physical properties for HCN and other simple cyanide
salts are summarized in Table 2-1.
The dissociation constants of metallocyanides vary significantly depending on oxidation
states, pH, temperature, and photodegradation (Beck, 1987). As noted above, some, such as
NaCN and KCN, dissociate completely when dissolved in water, whereas others do not.
Other inorganic and organic compounds containing the GST group include the nitriles, in
which the GST group is covalently bound to the rest of the molecule (e.g., acetone cyanohydrin
[ACH]), the cyanogens (i.e., compounds of the form NC-CN or X-CN, where X is a halogen),
such as cyanogen chloride (CNC1), and "cyanogenic" substances in some plant-based foods.
These cyanogenic compounds contain cyanogen glycosides that can undergo hydrolysis
following ingestion to produce HCN and other cyanide-containing compounds.
Anthropogenic sources are the main origin of cyanide in the environment, but cyanide is
also released from biomass burning, volcanoes, and natural biogenic processes from higher
plants, bacteria, and fungi (Agency for Toxic Substances and Disease Registry [ATSDR], 2006).
Cyanide compounds are used in a number of industrial processes, including mining, metallurgy,
manufacturing, and photography, due to their ability to form stable complexes with a range of
metals. Cyanide has been employed extensively in electroplating, in which a solid metal object
is immersed in a plating bath containing a solution of another metal with which it is to be coated,
in order to improve the durability, electrical resistance, and/or conductivity of the solid. HCN
has also been used in gas chamber executions and in chemical warfare. NaCN and KCN are also
used as rodenticides. Conversion factors for HCN air concentrations are 1 mg/m3 = 0.90 ppm
and 1 ppm =1.11 mg/m3.
Cyanide or cyanogenic compounds are found in many foods. Cyanide compounds occur
naturally as part of sugars or other naturally occurring compounds in certain plant-derived foods,
including almonds, millet sprouts, lima beans, soy, spinach, bamboo shoots, sorghum, and
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cassava roots. The parts of these plants that are eaten in the United States, however, contain
relatively low amounts of cyanide (ATSDR, 2006).
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Table 2-1. Physical and chemical properties of cyanide compounds
Characteristic
Hydrogen
cyanide
Sodium cyanide
Potassium
cyanide
Calcium cyanide
Potassium silver
cyanide
Cyanogen
CASRN #
74-90-8
143-33-9
151-50-8
592-01-8
506-61-6
460-19-5
Synonyms
Prussic acid,
hydrocyanic acid,
Cyclone B
Not found
Not found
Calcyanide,
calcyan
Potassium
dicyanoargentate
Dicyanogen,
ethanedinitrile,
oxalonitrile
Molecular weight
27
49
65
92
199
52
Form
Colorless gas or
liquid
White crystalline
powder
White lumps or
crystals
White powder
White crystals
Colorless gas
Chemical formula
HCN
NaCN
KCN
Ca(CN)2
AgK(CN)2
(CN)2
Boiling point (°C)
25.7
1496
1625
N/Aa
Not found
-21.17
Melting point (°C)
-13.4
563.7
634.5
640
Not found
-27.9
Density (g/mL)/
specific gravity (unitless)
0.6884
1.6
1.52
1.85
2.36
0.9537
Solubility
Ethanol, ether
Water, ethanol
Water, ethanol
Water, ethanol,
weak acid
Water, ethanol
Water, ethanol
"N/A = not applicable.
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Around the world, cassava is a vital staple for about 500 million people. Cassava is a
major source of carbohydrate in parts of Africa, South America, and Southeast Asia. Its starchy
roots produce more food energy per unit of land than any other staple crop, and it can be dried
and ground into flour. Its leaves, commonly eaten as a vegetable in tropical regions, provide
vitamins and protein. However, the storage root of the cassava plant contains linamarin, a
cyanogenic glycoside that is easily hydrolyzed by the enzyme linamarase (a P-glucosidase) to
release HCN. Although HCN can be readily removed during processing of cassava, cyanide
liberated from residual linamarin is associated with goiter in iodine-deficient populations with
chronic intake of cassava-based food products (Teles, 2002; Abuye et al., 1998).
Information on the concentration of cyanide in drinking water is available from the
National Drinking Water Contaminant Occurrence Database (U.S. Environmental Protection
Agency [U.S. EPA], 2003). In this database, a cross-sectional study of 16 states was used to
develop a statistical estimation indicative of the national occurrence of contaminants in drinking
water. Based on these data, the overall mean cyanide concentration in treated surface water and
groundwater systems was 2 |ig/L and 8 |ig/L, respectively. Cyanide was detected infrequently;
the average among the public water systems that detected cyanide was 60 |ig/L (parts per billion),
although some systems had levels in the parts per million range.
Although concentrations of HCN in foods are expected to be low, one author (Fiksel et
al., 1981) estimated that HCN intake from inhalation of air and ingestion of drinking water
would be less than the intake from food. Estimates of the HCN concentration in the total diet
were not located in the available literature. ATSDR (2006) estimated an atmospheric
concentration of 170 ppt (188 ng/m3), corresponding to an inhalation exposure to the general U.S.
nonurban, nonsmoking population of 3.8 |ig HCN/day, corresponding to 54 ng/kg-day HCN.
Smokers and those exposed to second-hand tobacco smoke make up a subset of the
general population that may be exposed to elevated levels of HCN. Smokers could be exposed
to 10 to 400 |ig HCN per cigarette, whereas nonsmokers exposed to sidestream smoke could be
exposed to 0.06 to 108 |ig HCN per cigarette (ATSDR, 2006).
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3. TOXICOKINETICS
3.1. ABSORPTION
The available data show that cyanide is rapidly and extensively absorbed via the oral,
inhalation, and dermal routes, although quantitative data on the percent or extent of absorption
are limited. Oral absorption has been reported as being lower at lethal doses. Some cyanide
salts, including KCN and NaCN rapidly dissociate in water. Because HCN is a weak acid (pKa
of 9.2), the acidic environment in the stomach favors the nonionized form (HCN) (U.S. EPA,
1992). The nonionized form is also favored under neutral conditions. Thus, HCN and the
dissociated sodium and potassium salts are predominantly present as HCN at the acidic pH levels
of the stomach and lower gastrointestinal tract. Accordingly, these compounds are presumably
absorbed by passive diffusion across the lipid matrix of the intestinal microvilli. The moderate
lipid solubility and small size of the HCN molecule also indicate that HCN crosses mucous
membranes rapidly. HCN is absorbed very rapidly after inhalation, and it penetrates the
epidermis. KCN and NaCN are corrosive to the skin, which can increase dermal absorption. In
the absence of such corrosion, however, these ionic forms of cyanide are absorbed less
completely than HCN via the dermal route.
Limited data are available on oral absorption of cyanide in humans. In a case report
(Liebowitz and Schwartz, 1948) of a suicide attempt by an 80 kg male who ingested an estimated
15-25 mg/kg CN as KCN, the authors estimated that, 2 hours after ingestion, the patient had 2.5
g GST in the body, of which 1.2 g was in the blood, based on a concentration of 200 mg HCN/L
in the blood at that time. This study does not provide any information on the disposition of the
remaining cyanide.
Gettler and Baine (1938) reported data indicating that, at doses above historical lethal
doses, absorption decreases with increasing dose levels. Absorption was estimated at 19.5, 18.1,
and 15.7% in people estimated to have ingested 297, 557, and 1,450 mg HCN in suicide attempts.
The low absorption at the highest dose may have been due to death occurring before absorption
was complete. The assumption that absorption is lower at higher lethal doses is supported by a
case report where absorption was approximately 82% in an individual, estimated as having
ingested 30 mg HCN, who died more than 3 hours after exposure.
Only limited data are available on absorption of inhaled cyanide by humans. Landahl
and Herrmann (1950) measured the pulmonary retention of HCN in 10 volunteers exposed to
concentrations of 0.0005-0.02 mg/L (0.5-20 mg/m3) for up to 3 minutes. All subjects breathed
through their mouths. The percent retained in the lung (and, presumably, the percent absorbed)
was approximately 60% and ranged from 58 to 77% among people who were breathing normally.
Rapid, shallow breathing appeared to decrease absorption.
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Dermal absorption of HCN gas has also been observed in humans. Drinker (1932)
reported that three workers who entered an atmosphere containing 2% HCN (20,000 ppm
[22,100 mg/m3]) became dizzy and weak and were on the verge of unconsciousness, despite
wearing gas masks providing respiratory protection. The observed effects were attributed to
dermal absorption of the gas. Potter (1950) reported on a worker, wearing respiratory protection
and protective clothing, who was accidentally exposed to liquid HCN. Within 5 minutes, the
worker became dizzy, had difficulty breathing, and fell unconscious.
Animal data also indicate extensive absorption. In male Sprague-Dawley rats treated by
gavage with 1 mg/kg KCN, a peak blood level of 6.2 nmol/mL CN~ (160 |ig/L) was observed 2
minutes after treatment, indicating rapid absorption (Leuschner et al., 1991). As described above
for oral poisonings in humans, absorption by dogs decreased as the dose increased well above
lethal levels (Gettler and Baine, 1938). Oral absorption was essentially comparable (16.6% and
15.7%) in dogs that ingested 100 and 50 mg HCN, respectively. However, absorption increased
to 72%) in a dog that ingested 20 mg HCN (1.5 mg/kg).
No animal studies were located that quantitatively evaluated the rate or extent of
absorption via the inhalation or dermal routes. However, Walton and Witherspoon (1926)
reported toxic effects and death in guinea pigs exposed by holding the open end of a test tube
containing liquid HCN against their shaved stomachs and concentration-related signs of toxicity
(including death) in dogs given whole-body exposure to HCN (excluding the head). These
results support the human data, demonstrating that absorption via the dermal route occurs in
animal species and can produce toxic effects, including death.
3.2. DISTRIBUTION
Cyanide distributes rapidly and uniformly throughout the body following absorption.
HCN enters the systemic circulation when inhaled or dermally absorbed (Yamamoto et al., 1982;
Potter, 1950; Drinker, 1932). Limited qualitative and quantitative data are available regarding
the tissue distribution of cyanide in humans from inhalation exposure studies to high doses of
cyanide. For example, cyanide was found in the lung, heart, blood, kidneys, and brain of humans
who died following cyanide inhalation (Gettler and Baine, 1938). In addition, Knowles and Bain
(1968) evaluated the relationship between blood concentrations of cyanide and short-term
accidental exposures to lethal levels of HCN in human case reports. Air concentrations of
>300 ppm (333 mg/m3 HCN), >200 ppm (222 mg/m3HCN), >100 ppm (111 mg/m3HCN), and
>50 ppm (55 mg/m3 HCN) corresponded to blood concentrations of >10, >8-10, >3-8, and >2-4
mg/L, respectively. The authors noted that there is considerable variability in this relationship,
presumably reflecting both interindividual variability and uncertainty of exposure duration and
concentrations estimated or measured retrospectively.
Limited data on distribution of cyanide in humans, following oral exposure, are available.
Immediately following oral cyanide exposure, the stomach contents appear to contain the highest
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concentration of cyanide. Other tissues containing cyanide included the liver, brain, spleen,
blood, kidneys, and lungs (Ansell and Lewis, 1970; Gettler and Baine, 1938).
Several animal studies are available that demonstrate the tissue distribution of cyanide
following both inhalation and oral exposures. In dogs exposed to lethal concentrations of
cyanide by inhalation, the highest concentrations of cyanide were found in the lungs, blood, and
heart (Gettler and Baine, 1938), while in rabbits exposed to 2,714 ppm HCN (3,000 mg/m3) for
5 minutes by inhalation, the highest tissue concentrations were detected in the heart, lung, and
brain, with lower levels in the spleen and kidneys (Ballantyne, 1983).
In rats and rabbits exposed by the oral route, the highest tissue concentrations of cyanide
were in the liver, lung, blood, spleen, and brain (Ballantyne, 1983; Ahmed and Farooqui, 1982;
Yamamoto et al., 1982). Yamamoto et al. (1982) compared cyanide distribution in rats
following oral gavage with NaCN (7 or 21 mg/kg CN) or inhalation exposure to HCN (average
of 356 or 1,180 ppm, equivalent to 393 or 1,303 mg/m3 HCN, respectively). These exposure
levels resulted in death within 10 minutes or less. Elevated concentrations were found in all
tissues evaluated following exposure via either route, but the relative concentrations were route
dependent. There was also some dose dependence, which may have been related to the faster
time to death at higher exposures for each route (approximately 10 minutes at the lower exposure
levels vs. 3-5 minutes at the higher levels). Focusing on the lower oral dose, the highest tissue
concentration of cyanide following exposure was in the liver, followed by the blood and lungs
and then the spleen and brain. After inhalation exposure, the highest concentration was in the
lungs, followed by the blood and liver and then spleen and brain. The kidney was not evaluated
for either exposure route. The route-specific difference may be related to first-pass metabolism
in the liver, following oral dosing, and initial deposition at the portal of entry, following
inhalation exposure.
Okoh and Pitt (1982) investigated the tissue distribution of cyanide in rats fed KCN in the
diet at 77 |imol/day (approximately 5.5 mg/kg-day CN) for 3 weeks and then injected
intraperitoneally (i.p.) with radiolabeled NaCN. Radioactivity was widely distributed, with the
highest concentrations in the gastrointestinal tract, blood, kidneys, lungs, spleen, and liver.
Radioactivity appeared in the stomach as early as 10 minutes after injection, with 18% of the
injected dose found in the stomach contents within 60 minutes of dosing. More than 80% of the
radioactivity in the stomach was present as thiocyanate, with small portions present as cyanide
and radiolabeled carbon dioxide.
In a subchronic study, male Sprague-Dawley rats (26-40/group) received KCN in their
drinking water at doses of 0, 40, 80, or 140/160 mg/kg-day for 13 weeks (Leuschner et al., 1991).
Blood was collected every 2 weeks for analysis of CN~ and SCN levels; both were found to be
dose related. Within each dose group, however, the levels of both cyanide and thiocyanate
remained fairly constant over the 13-week exposure period. Cyanide levels in the blood were
16-25 nmol/mL CN~ (420-650 |ig/L); thiocyanate levels were 341-877 nmol/mL SCN~ (20-51
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mg/L). Small amounts of thiocyanate were also detected in the control animals at concentrations
of 11-53 nmol/mL SCN- (0.64-3.1 mg/L) in plasma.
Howard and Hanzal (1955) exposed rats to HCN in the diet at an average daily dose of up
to 10.8 mg/kg-day CN for 2 years and found virtually no cyanide in the plasma or kidneys.
Cyanide was detectable in the red blood cells (RBCs) of less than half of the rats at an average
concentration of 2 |ig/100 g tissue.
McMillan and Svoboda (1982) incubated cyanide with washed erythrocytes, resuspended
in phosphate-buffered saline containing glucose and bovine serum albumin, and found that
cyanide concentrated in the RBCs. The major portion of cyanide in blood is sequestered in the
erythrocytes, and a relatively small proportion is transported via the plasma to target organs
(International Programme on Chemical Safety [IPCS], 2004). The RBC:plasma ratio for cyanide
is approximately 200:1 (IPCS, 1992). Binding to RBCs is primarily due to cyanide reacting with
ferric iron (Fe3 ) in methemoglobin to form the nontoxic complex cyanomethemoglobin (Chen
and Rose, 1952).
In rats treated orally with K14CN (5 mg/kg), radioactivity levels in plasma and whole
blood were initially (at 3 hours) much higher than levels in RBCs (Farooqui and Ahmed, 1982).
Levels in plasma and whole blood decreased rapidly, and red cell levels increased slightly, so
that at 24 hours plasma levels were only slightly higher than whole blood levels. Most of the
radioactivity in the red cells was in the heme fraction of hemoglobin rather than the membranes.
The reason for the finding of higher levels in the plasma than blood in this study is not clear, but
it may have been due to differences in sampling times. Approximately 60% of the cyanide in
plasma is bound to protein (IPCS, 1992).
Limited information indicates that cyanide can cross the placenta. Pettigrew et al. (1977)
compared cyanide and urinary thiocyanate levels in a small group of smoking and nonsmoking
pregnant women matched for age, height, parity and social class. Maternal plasma and urinary
thiocyanate levels were statistically significantly increased in smokers during gestation at weeks
28, 32, and 36; at delivery, only plasma thiocyanate was measured and was also statistically
higher in mothers who smoked. Mean urinary thiocyanate levels of neonates of smoking
mothers were elevated compared to those of nonsmokers (40.6 (amol/L compared to 23.1 (amol/L,
respectively), although the difference was not statistically significant, probably due to the small
sample size (n = 10).
Lactational transfer of cyanide and thiocyanate has been shown to occur in goats. Soto-
Blanco and Gorniak (2003) dosed lactating goats with 0, 1.0, 2.0, or 3.0 mg/kg-day KCN
(equivalent to 0, 0.4, 0.8, or 1.2 mg/kg-day GST) from lactation days 0 to 90 and measured whole
blood cyanide and thiocyanate concentrations on lactation days 30, 60, and 90. Both whole
blood cyanide and plasma thiocyanate concentrations were increased in a dose-dependent
manner in treated mothers, with mixed results regarding time dependence. In the offspring, both
blood cyanide and plasma thiocyanate increased with increasing maternal cyanide dose at
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lactation day 30 and decreased with lactation time. By lactation day 90, the concentration of
these compounds in the blood/plasma of the offspring was very low or undetectable. The study
authors attributed these findings to a decrease in milk consumption, accompanied by a
concomitant increase in solid food (grass and feed) during the latter part of lactation.
Small levels of cyanide are normally present in blood plasma at 0-140 |ig/L and in other
tissues at <0.5 mg/kg CN (ATSDR, 2006; Feldstein and Klendshoj, 1954). Chandra et al.
(1980) found that nonsmokers with no occupational exposure to cyanide had an average of
3.2 |ig/100 mL CIST (32 |ig/L) in blood; smokers had average blood cyanide levels of
4.8 |ig/100 mL (48 |ig/L). The background level is attributed to exposure to cyanogenic food,
vitamin Bi2, and passive tobacco smoke. Cyanide preferentially binds to hemoglobin in RBCs
but does not appear to accumulate in tissues after chronic oral exposure to inorganic cyanides
(Leuschner et al., 1991; Chen and Rose, 1952).
3.3. METABOLISM
The major metabolic pathway for cyanide is conversion to the less acutely toxic
compound thiocyanate, primarily by rhodanese, with some conversion occurring via 3-
mercaptopyruvate sulfur transferase. Conversion to thiocyanate accounts for 60-80% of a
cyanide dose. Minor pathways include incorporation into a 1-carbon metabolic pool or
conversion to 2-aminothiazoline-4-carboxylic acid (ATSDR, 2006). Conversion to 2-
aminothiazoline-4-carboxylic acid via reaction with cystine accounted for approximately 15% of
an injected dose of cyanide in rats (Wood and Cooley, 1956). Small amounts are also converted
to carbon dioxide in exhaled air or excreted unchanged as HCN in exhaled air. These pathways
are shown in Figure 3-1.
Rhodanese, a mitochondrial enzyme that converts cyanide to thiocyanate, facilitates
transfer of a sulfur atom to cyanide from a sulfane-sulfur donor. Because these donors must
contain an S-S bond, glutathione, thiosulfate, and cystine are sulfur donors for rhodanese,
whereas the thiols, cysteine, and reduced glutathione are not donors. Rhodanese is widely
distributed throughout the body. Using immunohistochemical staining techniques, rhodanese in
rabbits has been located in the liver, where it is most abundant in the hepatocytes near blood
vessels (Sylvester and Sander, 1990). It was also found in the lung, localized in epithelial cells
that formed the barrier between inhaled air and blood vessels. In the kidneys, rhodanese was
present in tubules closest to the glomeruli. The authors concluded that sites with the greatest
abundance of rhodanese are located to maximize conversion of cyanide to thiocyanate following
both oral and inhalation exposure. The presence of rhodanese in these tissues also indicates the
importance of first-pass metabolism in determining the toxicity of inhaled and ingested cyanide.
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CN-
Major Path (80%) .
Cyanide -
J • (SQST)
Mnor Path
2-Aminothiazo
ine-4-carboxylic acid &
2-Lninothiazolidine-4-carboxylic acid
HCN
Hydrogen Cyanide
(in expired air)
HCN
(Pool)
Urinary Excretion
CNO"
Cyanate
C02
Carbon dioxide
HCOOH
Formic Acid
Some
excreted
in urine
Metabolism of one-
carbon compounds
Formates
Figure 3-1. Cyanide primary metabolic pathways.
Source: Adapted from Ansell and Lewis (1970).
Metabolism of cyanide by rhodanese exhibits zero-order kinetics relative to cyanide; the
concentration of sulfur-containing donor molecules is the rate-limiting factor. The primary
endogenous sulfur donor is thiosulfate; others include glutathione and cystine. Schulz et al.
(1982) evaluated the metabolism of cyanide in humans continuously infused with the
hypotensive drug sodium nitroprusside, Na2[Fe(CN)5N0]-2H20, which completely releases the
CN in the blood. The authors estimated that the detoxification rate of cyanide in humans
(in the absence of antidotes) is about 1 |ig/kg-minute. McNamara (1976) estimated the
detoxification rate in humans as 17 |ig/kg-minute based on a study in men injected intravenously
(i.v.) with HCN. Lawrence (1947, reported in an extended abstract) found that continuous i.v.
infusion of NaCN into dogs at a rate of 0.013 mg/kg-minute (apparently as milligrams of CN but
not explicitly stated) was "tolerated over 37 hours" and speculated that this rate of infusion could
be tolerated indefinitely. Infusion rates of 0.028 mg/kg-minute or higher resulted in lethality.
Based on these findings, the whole-body rate of cyanide detoxification in dogs can be estimated
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to be approximately 13 |ig/kg-minute. The actual rate may be lower because some of the
cyanide at this dose may not have been detoxified but also may have been insufficient to cause
lethality. These findings suggest that the rate of cyanide detoxification in humans and dogs may
be similar.
Devlin et al. (1989a) evaluated rhodanese activity in rat liver and skeletal muscle. Using
histochemical staining techniques, the authors determined that only low levels of rhodanese
activity were present in the blood vessels. In contrast, high levels of rhodanese activity were
detected in the liver and skeletal muscle. Although the activity of rhodanese in muscle was
lower than in the liver, the authors concluded that the total skeletal muscle mass makes a
significant contribution to whole-body metabolism of cyanide. In a follow-up study in perfused
liver and hind-limb muscle, Devlin et al. (1989b) observed that the liver cleared 80% of the
available cyanide compared to 18% for the hind limbs. However, when the hind-limb data were
extrapolated to total muscle mass, muscle cleared cyanide 2.6-fold faster than did liver in the
absence of exogenous thiosulfate. When thiosulfate was included in the perfusion medium, liver
clearance was dependent on flow rate, but muscle tissue clearance was unaffected. Westley
(1981) found that purified bovine liver rhodanese has a high turnover rate of almost
20,000/minute in vitro (i.e., 1 mol of rhodanese could convert 20,000 mols of cyanide to
thiocyanate in 1 minute). This high turnover rate, coupled with the basal amount of rhodanese in
the liver and other tissues, means that the rate of cyanide metabolism should not depend critically
on the enzyme content of the tissue. Therefore, the limiting factor for cyanide metabolism is the
availability of the sulfur donor rather than the rhodanese metabolic capacity. Similarly,
differences between muscle and liver in ability to detoxify cyanide appear to be related to the
availability of sulfur donors.
Lewis et al. (1991) observed the presence of rhodanese in the epithelium of human nasal
tissue. Rhodanese activity in human nasal epithelium was higher in nonsmokers than smokers.
Individual enzyme kinetic data (Vmax and Km) suggested that decreased activity in smokers may
be due to decreased affinity. In kinetic studies with adequate sulfur present, rhodanese in human
nasal tissue exhibited a higher affinity (lower Km) for cyanide and a lower maximum velocity
(lower Vmax), compared to rhodanese in human liver. Human rhodanese exhibited a higher Km
and lower Vmax than did rat rhodanese. Dahl (1989) also found that rat nasal tissue exhibited
high levels of rhodanese activity, particularly in the olfactory region, which had almost sevenfold
more activity on a per milligram mitochondrial protein basis than did rhodanese in rat liver.
Rhodanese activity was also observed in the respiratory tracts of dogs, particularly in the nasal
cavity (Aminlari et al., 1994).
The tissue distribution and activity of rhodanese is highly variable among species. In
dogs, Himwich and Saunders (1948) observed that the highest activity of rhodanese was
observed in the adrenal glands, followed by the liver. The brain, spinal cord, kidneys, and testes
also had large amounts of rhodanese. In general, monkeys, rats, and rabbits had much higher
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rhodanese activity (milligrams CN converted to thiocyanate [SCN ] per gram of tissue) than
dogs, with the liver and kidneys containing the highest activity. Drawbaugh and Marrs (1987)
also studied the tissue distribution of rhodanese in several species, including marmosets, rats,
hamsters, rabbits, guinea pigs, dogs, and pigeons. The highest rhodanese activities were found in
rats, hamsters, and guinea pigs; the lowest were found in pigeons, marmosets, and dogs. Except
for rabbits, rhodanese activity was higher in the liver than in the kidneys of the species studied.
However, the authors noted that the biological significance of species differences in rhodanese
activity is unclear.
The study by Drawbaugh and Marrs (1987) has been used to suggest that the dog is not
an appropriate animal model for cyanide toxicity in humans, due to significantly lower levels of
rhodanese in this species as compared with humans. However, as noted above, the amount of
sulfur donor, not the amount of rhodanese itself, is the rate-limiting factor for detoxification of
cyanide by rhodanese, even at bolus doses resulting in high acute toxicity. Schulz (1984)
reported that the rate of cyanide detoxification in humans is slower than the rate in rodents or
dogs despite the higher levels of rhodanese in humans. Other data (McNamara, 1976) suggest
that the cyanide detoxification rate in humans is slightly higher than in dogs. Furthermore,
urinary concentrations of thiocyanate have been shown to be higher in dogs than rats (National
Toxicology Program [NTP], 1993; Kamalu, 1993). Lower thiocyanate levels would be expected
if metabolism via the rhodanese pathway were limited in this species. Although this analysis did
not normalize by urine specific gravity or other factors, it suggests that cyanide was metabolized
to a similar degree in dogs and rats.
Chronic exposure to cyanide resulted in increased rhodanese levels in rabbits, suggesting
that rhodanese is inducible, at least in this species (Okolie and Osagie, 1999). Alternatively, the
increased levels could be due to other factors, such as increased protein stability. Data were not
located regarding whether chronic exposure increases rhodanese levels in other species.
Several polymorphisms in rhodanese have been identified in human populations, though
only a minimal effect on cyanide detoxification was detected (Billaut-Laden et al., 2006). A
second enzyme that converts cyanide to thiocyanate is mercaptopyruvate sulfurtransferase
(MPST). This enzyme differs from rhodanese in that it catalyzes the transfer of sulfur from an
organic thiol to cyanide (Wing and Baskin, 1992). Therefore, this enzyme breaks a carbon-
sulfur bond to facilitate transfer of sulfur to cyanide, whereas rhodanese breaks a sulfur-sulfur
bond. MPST is most active at pH 9.5, while rhodanese is most active at pH 8.6, which is closer
to physiological pH. MPST also appears to have a different tissue distribution from that of
rhodanese; this enzyme has been reported as being located in the RBCs and kidneys. MPST is
located in both the mitochondria and the cytosol, making it more accessible for conversion of
cyanide than rhodanese, which occurs only in the mitochondria (Wing and Baskin, 1992).
Support for the role of MPST in cyanide detoxification was provided in in vitro studies by Huang
et al. (1998). These authors demonstrated that addition of L- or D-cysteine to hepatocytes in cell
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culture prevented cyanide cytotoxicity and enhanced the formation of thiocyanate.
Mercaptopyruvate and thiocystine, metabolites of L- and D-cysteine, are substrates of MPST.
Huang et al. (1998) observed that, when formation of these metabolites in isolated hepatocytes
was prevented, the formation of thiocyanate was also inhibited. However, it is not clear whether
MPST directly transfers sulfur to cyanide or whether it acts indirectly by transferring sulfur to
albumin in the liver. The modified albumin could then be excreted to form a sulfane-sulfur pool
that is available to react with cyanide via rhodanese (Wing and Baskin, 1992).
Although the reaction of rhodanese with cyanide is irreversible, thiocyanate can be
converted back to cyanide and sulfate by the action of thiocyanate oxidase located in the RBCs,
lymphocytes, mammary gland, and thyroid (Wood, 1975). Thiocyanate oxidase has been found
in the erythrocytes of humans, dogs, rabbits, and rats (Goldstein and Rieders, 1953).
This enzyme catalyzes the reaction of hydrogen peroxide and thiocyanate to form cyanide and
sulfate. In addition, these enzymes produce an intermediate oxidation product of thiocyanate, the
OSCN ion known as hypothiocyanate, which reacts with cyanide to form cyanate (OCN ),
which is then hydrolyzed to ammonia and carbon dioxide (Wood, 1975).
The minor pathway shown in Figure 3-1 involves the spontaneous reaction of cyanide
with cystine to yield 2-aminothiazoline-4-carboxylic acid, which tautomerizes to 2-imino-4-
thiazolidinecarboxylic acid. This pathway accounted for approximately 15% of the cyanide dose
in a female rat receiving daily i.p. injections of NaCN for 8 days (Wood and Cooley, 1956). In
another experiment in the same publication, the percentage metabolism via this pathway was
higher when rats were injected i.v. with labeled cystine and subsequently were administered
NaCN subcutaneously.
The mean blood thiocyanate level in smokers with untreated tobacco amblyopia (a
condition causing visual defects that has been attributed to cyanide exposure) was significantly
lower than the concentration in smokers overall, suggesting that people with this condition have
a decreased ability to convert cyanide to thiocyanate (Pettigrew and Fell, 1973). Blood cyanide
levels were low in both smokers and nonsmokers in this study, with no significant effect of
smoking or tobacco amblyopia, perhaps because approximately 1-2 hours had elapsed between
the time the last cigarette was smoked and when cyanide levels were measured. The authors
suggested that the excess cyanide was bound up as cyanocobalamin (one form of vitamin Bi2),
but they did not investigate this hypothesis. Cyanide also reacts with methemoglobin
(hemoglobin that has been oxidized either by normal metabolism or by xenobiotic oxidant
stressors) in RBCs to form cyanomethemoglobin. Schulz (1984) noted that, theoretically, 1 g of
methemoglobin can bind approximately 60 jamol of HCN and that 1 L of erythrocytes should be
able to bind approximately 50-200 |imol (1.4-5.4 mg) HCN at physiological levels of
methemoglobin (0.25-1%). This readily reversible reaction is considered to be a naturally
occurring detoxification pathway for low levels of cyanide in the blood (Lundquist et al., 1985)
and forms the basis of the first phase of treatment for acute cyanide poisoning, which consists of
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administering amyl nitrite or sodium nitrite to cyanide-poisoned individuals (Klaassen, 2001).
Amyl nitrite and sodium nitrite are oxidants that increase the conversion of hemoglobin to
methemoglobin, thus providing a sink for CN away from tissue cytochrome c oxidase. Under
this treatment approach, cyanide is then detoxified by slow release from cyanomethemoglobin
and cytochrome c oxidase and subsequent conversion by the enzyme rhodanese to SCNT, which
has much lower acute toxicity than cyanide. Sodium thiosulfate administered as the second
phase of treatment for acute cyanide poisoning accelerates detoxification by supplying a sulfur
substrate for the reaction. Other substances used to detoxify cyanide include hydroxycobalamin
(vitamin Bi2a), an antidote used outside the U.S. that binds cyanide to form cyanocobalamin
(vitamin Bi2), and cobalt edetate, which is used as an antidote in some countries due to the high
affinity of cobalt for cyanide (Klaassen, 2001).
3.4. ELIMINATION
Data in humans and animals indicate cyanide is primarily excreted in the urine as
thiocyanate following both inhalation and oral exposure. Smaller amounts are excreted as
urinary cyanide or as HCN or carbon dioxide in exhaled air. Following occupational exposure to
0.19-0.75 ppm HCN, urinary thiocyanate levels in nonsmoking exposed workers were
approximately seven times the levels in nonsmoking controls (Chandra et al., 1980). Urinary
cyanide levels were also elevated in the exposed workers, but they were approximately two
orders of magnitude lower than thiocyanate levels.
Following a single subcutaneous injection of rats with [14C] KCN, 89% of the excreted
radioactivity was detected in urine within 24 hours; about 4% of the excreted radioactivity was
expired in air, primarily as carbon dioxide (Okoh, 1983). The authors found that 71-79% of the
urinary activity was in the form of thiocyanate. The excretion pattern was not affected by prior
exposure to cyanide in diet for 6 weeks.
In a related study of rats injected i.p. with radiolabeled NaCN after being fed KCN in the
diet at 77 |imol/day (approximately 5.5 mg/kg-day CN) for 3 weeks (Okoh and Pitt, 1982), 86%
of the radioactivity in the expired air was present as carbon dioxide and 14% was present as
HCN. Boxer and Rickards (1952) also observed that exhaled air contained both radiolabeled
HCN and carbon dioxide after dogs were injected subcutaneously with radiolabeled NaCN.
However, the primary path of excretion was still in the form of urinary thiocyanate, though small
amounts of cyanide and cyanocobalamin were also found in the urine. Sylvester et al. (1983)
treated dogs with i.v. NaCN and found less than 1% of the total cyanide dose was eliminated
through exhaled air.
Leuschner et al. (1991) evaluated the elimination of KCN following both acute and
subchronic exposure. For the acute study, three male Sprague-Dawley rats were treated by
gavage with 1 mg/kg KCN. Blood was collected at regular intervals for up to 1 hour following
administration. A peak blood level of 6.2 nmol/mL (160 |ig/L) CN~ was observed 2 minutes
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after treatment; by 60 minutes, the blood levels had dropped to the analytic detection limit. The
authors calculated an elimination half-life of 14 minutes.
In the subchronic portion of Leuschner et al. (1991), male Sprague-Dawley rats (26-
40/group) received KCN in their drinking water at doses of 0, 40, 80, or 160 mg/kg-day for
13 weeks. Blood was collected every 2 weeks for analysis of cyanide and thiocyanate levels.
Urine was collected over a 16-hour period during weeks 6 and 13 of the study to determine
cyanide and thiocyanate levels. Similar patterns of excretion were observed for both urinary
levels of cyanide and thiocyanate. A dose-response relationship was observed for the
concentration of both cyanide and thiocyanate in urine, and a small amount of thiocyanate was
observed in the urine of the controls. The levels of cyanide in the urine were much lower than
the thiocyanate levels; the ratio of cyanide to thiocyanate was about 1 to 1,000. Approximately
11% of the administered cyanide was eliminated per day as urinary thiocyanate during the dosing
period, while only about 0.003% was excreted per day unchanged. The study authors did not
report how they estimated the percent of total dose eliminated; radiolabeled material was not
used. Elimination half-life was not calculated for the subchronic study. Blood levels of cyanide
and thiocyanate were fairly consistent with time. Some elimination may have occurred as
exhaled HCN or carbon dioxide, but data indicate that this route accounts for <10% of a dose of
cyanide following acute dosing (Okoh, 1983). The study authors (Leuschner et al., 1991) also
noted that the percent of administered cyanide excreted via the urine was unchanged between
weeks 6 and 13, indicating that detoxification pathways were not saturated and the mode of
cyanide excretion was not affected over this period.
In cynomolgus monkeys exposed via inhalation for up to 30 minutes to approximately
100-170 mg/m3 HCN, levels of cyanide in the blood remained nearly constant after
approximately the first 10-15 minutes of exposure and for 60 minutes following termination of
exposure (Purser et al., 1984). Thus, the half-life under these conditions was longer than in the
Leuschner et al. (1991) gavage study in rats. The difference may have been due to saturation of
metabolism, first-pass metabolism following oral exposure, or species-related differences.
3.5. PHYSIOLOGICALLY BASED TOXICOKINETIC MODELS
A pharmacokinetic analysis of the distribution and metabolism of cyanide was conducted
in dogs following a single i.v. dose (Sylvester et al., 1983). Dogs (six/group) were administered
i.v. saline, NaCN (20.4 |imol/kg), or sodium thiocyanate (12.3 |imol/kg); cyanide concentration
was determined in whole blood, and thiocyanate concentration was determined in plasma. Blood
levels of CN and SCN measured after administration were used to develop a pharmacokinetic
model in dogs. The conversion of cyanide to thiocyanate was found to follow first order kinetics
Three hours following i.v. dosing with cyanide, 90% of the total dose had been converted to
thiocyanate. The half-life of thiocyanate was determined to be 29 hours. The authors also found
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that less than 8% of the cyanide dose was eliminated through non-thiocyanate routes and only
1% of the total cyanide dose was eliminated through exhalation.
Additionally, some data exist on the comparative toxicokinetics of cyanide and
thiocyanate in several species (Sousa et al., 2003). Rats (n = 42), pigs (n = 6), and goats (n = 6)
were studied up to 24 hours after a single gavage dose of 3.0 mg/kg KCN. Cyanide was quickly
absorbed in all species. The peak plasma concentration of cyanide was highest in goats,
followed by rats and pigs. Goats also had the highest volume of distribution, highest area under
the curve (AUC), and slowest elimination compared with the other two species. The similarities
in absorption data between species indicated that pH differences between the monogastric
stomachs of rats and pigs (pH 1-2) and ruminant stomachs (pH 6.8) did not noticeably impact
absorption of cyanide. Toxicokinetic parameters for thiocyanate indicated the peak plasma
concentrations and AUC to be greatest in rats, followed by goats and pigs. Blood levels of
cyanide in each species indicate rapid decreases in cyanide blood concentration by 3 hours
following dosing, with the half-life of elimination for thiocyanate for all species about 9-11
times longer.
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4. HAZARD IDENTIFICATION
4.1. STUDIES IN HUMANS—EPIDEMIOLOGY, CASE REPORTS, CLINICAL
CONTROLS
4.1.1. Acute Oral, Inhalation, and Dermal Studies
The effects of acute, high-level exposure to cyanide are well characterized (reviewed in
AT SDR [2006], IPCS [2004], U.S. EPA [1992], and Hall and Rumack [1990]). Although acute
oral doses of cyanide cause cardiovascular, respiratory, and neurophysiological changes, the
brain appears to be the organ most sensitive to acute cyanide toxicity (IPCS, 2004). Several
studies of the acute effects of cyanide in humans, following suicide attempts or accidental
poisoning by the oral and inhalation routes, provide additional details, although most such
studies include only a limited characterization of exposure. Symptoms of severe cyanide
poisoning include vomiting, nausea, weakness, confusion, lethargy, cyanosis, weak and ataxic
movements, increased respiratory and heart rates, progressing to coma with respiratory
depression, seizures, cardiovascular collapse, and death. The principal feature of the acute
toxicity profile for cyanide includes lethality by all routes of administration, with a steep rate-
dependent dose-response curve. Death from cyanide poisoning is believed to result from central
nervous system (CNS) depression, subsequent to inhibition of brain cytochrome oxidase activity
(Way, 1984). The toxic effects and lethality associated with acute exposure to CN in humans
and animals are generally similar and are believed to result from inactivation of cytochrome
oxidase and inhibition of cellular respiration during the terminal reaction of the electron transport
chain. This inhibition prevents the formation of adenosine triphosphate (ATP) via oxidative
phosphorylation. IPCS (2004) has reported that in humans the lowest reported oral lethal dose is
0.54 mg/kg body weight; the average absorbed dose at the time of death was estimated at
1.4 mg/kg body weight (calculated as HCN). Rapid recovery from relatively low, short-term
inhalation exposures often occurs once the exposed individual is moved to fresh air. Individuals
suffering from higher oral and inhalation exposures may benefit from supplemental oxygen and
the use of antidotes. Oral exposure results in slower absorption, passage to the liver, and faster
detoxification. If the patient responds to treatment and survives, recovery is usually prompt and
complete; however, delayed neurological symptoms, including neuropsychiatric manifestations
and Parkinson-type disease, can occur (IPCS, 2004). Exposure to lower levels can cause
flushing, light-headedness, dizziness, headache, and other symptoms indicative of hypoxia
(Wolfsie and Shaffer, 1959).
Liebowitz and Schwartz (1948) reported a case of cyanide poisoning following ingestion
of an estimated 38-63 mg/kg of KCN (15-25 mg/kg CIST). The patient was comatose with
muscular rigidity and a thready pulse on admission. By 8 hours after admission, the patient was
alert, and the symptoms had begun to subside (with the exception of weakness, nausea, and an
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enlarged heart). The authors suggested that the reason that the patient recovered from exposure
to a dose that was about 30 times the estimated lethal dose may have been because he was a
chemist who frequently immersed his hands in thiosulfate. Therefore, exposure to thiosulfate
may have had an antidotal effect.
Several authors (Grandas et al., 1989; Rosenberg et al., 1989; Carella et al., 1988; Uitti et
al., 1985) reported the development of symptoms of parkinsonism in patients who recovered
from ingestion of a single dose of cyanide. The four cases included an 18-year-old male who
ingested 5.6-7.6 mg/kg cyanide in a suicide attempt (Uitti et al., 1985), a 46-year-old woman
who ingested an unreported amount of cyanide by accidental poisoning (Carella et al., 1988), a
46-year-old man who ingested 8.6 mg/kg cyanide in a suicide attempt (Rosenberg et al., 1989),
and a 39-year-old man who ingested an unknown amount of cyanide in a suicide attempt
(Grandas et al., 1989). In all cases, the patients recovered from the acute symptoms of cyanide
poisoning with treatment, and a neurologic examination immediately following the poisoning
was normal. Follow-up neurologic examination at times of 3 weeks (Rosenberg et al., 1989),
4 months (Uitti et al., 1985), or 1 year (Grandas et al., 1989; Carella et al., 1988), however,
revealed that the patients had developed symptoms of parkinsonism, including generalized
rigidity, bradykinesia, tremors of tongue and eyelids, slow-shuffling gait, and a weak dysphonic
voice. A computerized tomography scan or magnetic resonance imaging showed lesions in the
putamen and globus pallidus regions of the brain (Grandas et al., 1989; Rosenberg et al., 1989;
Carella et al., 1988; Uitti et al., 1985). Similarly, Lam and Lau (2000) reported mild impairment
of recent memory and concentration, which was confirmed by neurological testing, a year after a
19-year-old woman experienced an episode of acute inhalation exposure to cyanide.
Potter (1950) reported a case history of a worker accidentally exposed via inhalation to an
undetermined concentration of HCN. Early symptoms were dizziness, dyspnea, and weakness of
the legs followed by a period of deep unconsciousness accompanied by absent reflexes,
stertorous respiration, rapid pulse, fixed and unreactive pupils, and convulsions. The subject
recovered with treatment, and no after-effects were reported. In another case report, a worker
was found unconscious, lying in tank sludge after working without protective gear in a plating
tank containing silver-cyanide sludge (Singh et al., 1989). The duration of exposure was
unknown, but the tank air was later measured to contain 200 ppm HCN (220 mg/m3 HCN). He
had dermal evidence of chemical burns and was not breathing, with a rapid pulse, fixed and
dilated pupils, and no recorded blood pressure or response to pain. Blood cyanide was
804 |imol/L within V2 hour of hospital arrival and decreased to 15 |imol/L at 18 hours after
arrival and following detoxification efforts. The patient died within 3 days despite extensive
treatment.
Cyanide is readily absorbed through the skin (Lam and Lau, 2000; Potter, 1950);
therefore, systemic toxicity can readily result from dermal exposure to cyanide fumes or direct
dermal contact with HCN. A case report of a worker accidentally exposed to a brief stream of
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liquid HCN on his hand (amount not specified) reported that the worker became deeply
unconscious within 5 minutes of exposure (Potter, 1950). Breathing was hoarse, his face was
flushed, and reflexes were absent. The subject recovered with sodium nitrite and sodium
thiosulphate treatment.
Dizziness, weakness, and a throbbing pulse were reported when three workers wearing
gas masks entered an atmosphere containing 2% HCN gas (Drinker, 1932). These effects were
attributed to the dermal absorption of the gas. The men developed symptoms of dizziness,
weakness, and throbbing pulse after about 10 minutes of exposure and eventually became
unconscious. The symptoms persisted for several hours following exposure.
Relatively mild symptoms of cyanide poisoning (flushing, dizziness, headache, throat
discomfort, chest tightness, skin itchiness, and eye irritation) were reported in firefighters who
were wearing self-contained breathing apparatus when they responded to a HCN gas release
(Lam et al., 2000). The effects were attributed to dermal cyanide absorption and direct contact
with skin and eyes.
4.1.2. Subchronic and Chronic Oral Studies
No subchronic or chronic studies of human exposure to cyanide by the oral route were
located. However, a number of studies have examined populations exposed to cyanogenic
compounds in foods, particularly cassava root, which can be dried and ground into flour and is a
primary source of carbohydrate in many parts of Africa and Southeast Asia (Bonmarin et al.,
2002; Okafor et al., 2002; Makene and Wilson, 1972). Due to concern about effects of
cyanogenic compounds, many of the recent studies of chronic cyanide toxicity have been
conducted in the developing world, where exposure to cyanogenic compounds in food is a
significant public health and agricultural (for livestock) concern. Symptoms reported in these
populations include ataxic tropic neuropathy, spastic paraparesis (paralysis, particularly of the
lower extremities), optic atrophy, and decreased nerve conduction velocity. Effects seen in these
studies are often confounded by dietary deficiencies; particularly low dietary intake of protein,
vitamin Bi2, and/or iodine; and overall malnutrition. In addition, animal studies comparing the
effects of cassava ingestion and ingestion of cyanide indicate that some of the observed effects
are due to compounds besides cyanide in cassava (Banea-Mayambu et al., 1997; Kamalu, 1993;
Olusi et al., 1979). Because of the confounding factors of dietary deficiencies in the studied
populations and the presence of other potentially toxic compounds in cassava, human and animal
studies of cassava are of limited use for the hazard assessment of cyanide.
4.1.3. Subchronic and Chronic Inhalation Studies
Several reports of occupationally exposed workers indicate that chronic exposure to low
concentrations of cyanide can cause alterations of thyroid function and neurological symptoms
(Baneijee et al., 1997; Blanc et al., 1985; El Ghawabi et al., 1975). Occupational exposure to
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cyanide occurs primarily via inhalation, although dermal and limited oral exposure also can
occur. HCN is noted for its systemic toxicity, which would be expected to occur at
concentrations below those at which any direct respiratory tract effects would be anticipated.
El Ghawabi et al. (1975) evaluated the effects of long-term occupational exposure to
cyanide in 36 male workers employed in the electroplating sections of three factories in Egypt.
Cyanide exposure was from a plating bath that contained 3% copper cyanide, 3% NaCN, and 1%
sodium carbonate. Individual breathing zone air samples were taken to determine the levels of
airborne cyanide to which the men were exposed. Fifteen-minute air samples were collected by
using a Midget impinger. Twenty male volunteers of the same age group and socioeconomic
status who had no occupational exposure to cyanide were chosen as controls. Information on
how or from where the controls were recruited was not provided. None of the exposed or control
workers were cigarette smokers at the time of the study. Participants were prohibited from
ingesting cyanide-containing foods during the course of the investigation. Cyanide-exposed
workers and controls were given medical examinations (with special focus on thyroid
abnormalities), interviewed regarding medical history, and questioned regarding symptoms
experienced. Thyroid function (as measured by uptake of radiolabeled iodide) was assayed, and
urinary levels of SCN were recorded over a two month period and reported as average daily
excretion (in milligrams). No investigation of thyroid hormone levels was reported. Of the 36
workers, 14 had been exposed for 5 years, 14 for 5-10 years, 7 for 10-15 years, and 1 for greater
than 15 years. The mean and median exposure times for the worker population were not
reported, but, based on the categorical data provided, the median duration can be estimated to be
approximately 5 years. The mean cyanide air concentrations in the breathing zones of workers at
each of the three plants were 10.4, 6.4, and 8.1 ppm (11.5, 7.1, and 8.9 mg/m3) HCN,
respectively, with a range of 4.2-12.4 ppm (4.6-13.7 mg/m3) HCN. The authors reported
workers were exposed to other chemicals during the electroplating process (e.g., gasoline, alkali,
and acid), though concentrations to these other chemicals were not quantified. Urinary SCN~
concentrations from exposed workers were measured during two successive months.
Graphically presented data of mean individual urinary SCN~ levels plotted against the
concentration of HCN in the air indicated a strong positive linear relationship between urinary
SCN~ and HCN concentration in the air (Figure 4-1).
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10
CO
"D
O)
E
O
CO
8-
6-
4-
0-
0
• •
~ ~
>
¦ ^
(0
yf ~ ~
• Factory A
¦i 2"
¦ Factory B
=>
~ Factory C
i 1 1 1 1 1 1 1 1 1 1 1 1 r
5 10
HCN in air (ppm)
15
Figure 4-1. Urinary SCN- of exposed workers plotted against individual
breathing concentrations of HCN.
Source: El Ghawabi et al. (1975).1
Twenty of the 36 exposed workers (56%) had thyroid enlargement rated as being mild to
moderate; however, there was no correlation between duration of exposure and either incidence
or magnitude of enlargement. The authors reported that none of the workers showed clinical
symptoms of either hypo- or hyperthyroidism, although the basis of this assessment was not
provided.. Radioactive iodine uptake measured following a 2-day break in HCN exposure
indicated statistically significantly elevated iodide uptake after 4 hours (38.7% compared to
22.4%) and 24 hours (49.3% compared to 39.9%) as compared with controls (Table 4-1). In a
separate assay, blood samples taken 72 hours after administration of the iodide tracer indicated
protein bound I131 (PBI131) in the blood to be similar between controls and workers (0.11 ± 0.041
compared to 0.12 ± 0.039).
Table 4-1. Thyroid uptake of 1311 in electroplating workers
Percent 131I thyroid uptake (mean ± standard deviation)
After 4 hours
After 24 hours
Controls (n = 20)
22.42 ±7.21
39.95 ±4.80
Exposed (n = 36)
38.722 ±6.63a
49.33 ± 10.6la
1 Reproduced by EPA from a graph published in El Ghawabi et al. (1975) by estimation of original data points and
regeneration of graph. EPA's reproduced linear regression line had a slope of 0.63 compared to the slope of 0.65
reported in the published study and a statistically significant Pearson correlation coefficient of 0.5 (p < 0.001).
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"Significant difference (p < 0.001) by Student's t-test.
Source: El Ghawabi et al. (1975).
The authors noted that the radioactive iodide uptake test was conducted after the workers
had been away from work for 2 days, allowing time for thiocyanate, with a 3-day elimination
half-life (as established by Schulz et al. [1979]), to be partially cleared from their systems. The
authors suggested that the sudden cessation of cyanide exposure may have caused the thyroid
gland to rapidly accumulate iodine. Increased 24 hour uptake of radioactive iodide has been
reported to occur in hyperthyroidism, iodine deficiency, and goiter (Ravel 2005; NLM 2008a).
Other findings noted in this study included significantly higher hemoglobin (14.8 vs. 13.4
g/dL) and lymphocyte counts (42 vs. 30%) and punctate basophilia of erythrocytes in 78% of
workers. The authors indicated that the observed punctuate basophilia was not characteristic of
HCN exposure and may be related to other concomitant chemical exposures. Symptoms
reported more frequently in the exposed workers than in the controls included (in decreasing
order of frequency) headache, weakness, and changes in the senses of taste and smell.
Incidences of symptoms at individual plants were not reported, and no evaluation of symptoms
by exposure concentration was presented. Based on the observed thyroid effects, the lowest
mean concentration recorded in the three factories of 6.4 ppm (7.1 mg/m3) HCN was designated
as a lowest-observed-adverse-effect level (LOAEL) for this review.
A group of 36 male workers who had been exposed to HCN fumes in a silver-reclaiming
facility in Illinois were retrospectively studied by Blanc et al. (1985), following the death of one
employee from cyanide overexposure and the closure of the factory due to health and safety
violations. The authors attempted to recruit all previous employees. In this study, data
collection for the former workers included physical examinations (including examination of
neurological effects), serum biochemistry, hematology, urinalysis, serum enzymes (aspartate
aminotransferase [AST] and alanine aminotransferase [ALT]), and thyroid hormone analysis
(thyroid-stimulating hormone [TSH] and thyroxine [T4]), and a questionnaire designed to
determine exposure, symptoms during employment, and current symptoms. Workers were
qualitatively categorized into low-, moderate-, or high-exposure groups based on their primary
job activities. The median time elapsed since last employment at the facility was 10.5 months;
the median duration of employment was only 8.5 months. Environmental monitoring conducted
the day after the plant was shut down found that the 24-hour time-weighted average (TWA)
exposure was 15 ppm (16.6 mg/m3) HCN. None of the former workers were found to have
palpable thyroid gland abnormalities, mucosal erosion, or focal neurological deficits. Clinical
tests revealed decreases in the absorption of vitamin Bi2 (a possible factor protecting against
cyanide toxicity) and decreased folate levels that may have been secondary to the decrease in
vitamin Bi2. Serum TSH levels were also reported as being significantly elevated in workers
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relative to laboratory controls (concurrent controls were not used in this study design).
Triiodothyronine (T3) uptake in the highest exposed workers (n = 9) was statistically
significantly elevated compared to that in laboratory controls (Table 4-2). The authors reported
that this elevation may reflect a postinhibitory response.
Table 4-2. Thyroid parameters in former silver-reclaiming workers
Population
Percent T3 uptake"
TSHa (jiU/mL)
Laboratory controls (n = 100)
30.0 ±2.8
1.7 ± 1.2
All workers (n = 33)
30.9 ±2.6
2.2 ± 1.6b
Low-exposure workers (n = 24)
30.4 ±2.3
2.2 ± 1.7
High-exposure workers (n = 9)
32.4 ± 2.4°
2.4 ± 1.3
aValues are mean ± standard deviation.
Statistically significant by Student's t-test atp< 0.05 compared to laboratory controls.
Statistically significant by Student's t-test atp< 0.01 compared to laboratory controls.
Source: Blanc et al. (1985).
A statistically significant positive trend for self-reported weight loss was demonstrated
against the exposure index, supporting an exposure-response relationship. A statistically
significant trend was also found between the incidence of symptoms reported during active
employment (headache, dizziness, nausea, and bitter almond taste), as well as those reported at
the time of the survey (after adjustment for time elapsed since exposure) and the qualitative
index of exposure, providing evidence of another exposure-response relationship. Some of the
symptoms were reported as persisting for 7 or more months following exposure termination.
The reported central nervous system effects suggest the occurrence of neurotoxicity associated
with exposure to cyanide or its metabolite, thiocyanate. Dermal exposure to cyanide was
reported by half of the workers, and additional exposure by ingestion was likely due to poor
general hygiene in the factory in addition to inadequate personal protective equipment and
worker training. Because there were multiple possible routes of cyanide exposure, including
dermal exposure and contamination of food, data do not support for the selection of a reliable
LOAEL for inhalation. This study does demonstrate, however, the occurrence of non-transient
effects of thyroid function (as measured by percentage of T3 uptake) from occupational exposure
to HCN.
In a study of electroplating workers in a factory in India, Banerjee et al. (1997) compared
levels of the thyroid hormones T3, T4, and TSH in 35 male workers who had been exposed to
cyanide via inhalation for more than 5 consecutive years to a randomly selected control group of
35 unexposed male workers matched for age and dietary habits. None of the subjects used
tobacco products or had a prior history of thyroid disease. No environmental monitoring data
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were provided on HCN levels in the factory. However, serum SCN levels, a measure of internal
dose, were reported in both workers and controls. The average serum thiocyanate level in the
exposed workers was 316 (amol/L compared with 90.8 |imol/L in the controls, a difference that
was statistically significant (p < 0.01). The exposed workers had significantly lower levels of T3
and T4 (48 and 37% lower, respectively) and significantly higher levels of TSH (142%)
compared with controls (Table 4-3). In addition, there was a significant negative correlation
between serum T4and thiocyanate concentrations (r = -0.363,p< 0.05) and a significant positive
correlation between TSH and thiocyanate concentrations (r = 0.354,p< 0.05). There was also
an apparent negative correlation between T3 and thiocyanate (r = -0.245), but this difference was
not statistically significant. Levels of T3 and T4in exposed workers were outside the reported
normal range for these endpoints, indicating a potentially clinically relevant alteration of thyroid
hormone levels.
Table 4-3. Thyroid parameters in HCN-exposed and unexposed electroplating
workers
SCN (jimol/L)a
T4 (jig/dL)a
T3 (jig/dL)a
TSH (jiU/mL)a
Controls (n = 35)
90.8 ± 9.02
6.09 ±0.601
111.0 ± 9.3
1.2 ±0.301
Exposed (n = 35)
316 ± 15.0b
3.81 ±0.3181c
87.2 ± 8.1°
2.91 ± 0.201°
aValues are mean ± standard deviation.
Statistically significant by Student's t-test atp< 0.01 compared to laboratory controls.
Statistically significant by Student's t-test atp< 0.05 compared to laboratory controls.
Source: Baneijee et al. (1997).
As part of a report on excretion of cyanide and its metabolites, Chandra et al. (1980)
reported on a group of 23 electroplating workers chronically exposed to HCN fumes at 0.2-
0.8 mg/m3, with a mean value of 0.45 mg/m3. The title of this report indicated that workers were
exposed chronically, though exposure durations were not provided. The concentration in the
breathing zone was reported as 0.1-0.2 mg/m3, with a mean of 0.15 mg/m3. The authors noted
that the workers complained of symptoms typical of cyanide poisoning but provided no
additional information on specific symptoms or further analysis. In the absence of further
information, no independent assessment of this study is possible.
Chatgtopadhyay et al. (2000) investigated the effect of exposure to cyanide fumes on
pulmonary function in workers at a metal-tempering plant. The authors evaluated 24 workers in
an initial assessment and conducted a follow-up study on 17 of these workers 2 years later. The
control group for the initial study consisted of 14 unexposed workers matched for socioeconomic
status and race. The follow-up study did not include a concurrent control group; data from the
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control group in the initial assessment were used for comparison. No information was provided
on cyanide concentrations in the air. The mean duration of exposure was 21.0 ± 5.03 years at the
time of the first assessment. In the initial study, there were statistically significant decreases in
pulmonary function as assessed by reduced peak expiratory flow rate, forced vital capacity
(FVC), and forced expiratory volume (FEV) in 1 second as a percentage of FVC (FEVi%);
decreases in other pulmonary function parameters were not statistically significant. In the
follow-up study, statistically significant decreases were observed in all measured pulmonary
function parameters. However, concurrent controls were not utilized in the follow-up study.
Furthermore, adjustments for smoking and other co-occurring chemical exposures as sources of
potential confounding were not conducted.
Population based studies examining inhalation of cyanide at ambient levels and potential
health outcomes are limited. However, studies examining smokers, a subgroup with higher
inhalation exposure to cyanide through tobacco smoke, have indicated an association between
smoking and thyroid disorders. Specifically, a meta-analysis of eight studies found a statistically
significant association between smoking and the development of goiter in women (OR= 1.29
95% CI 1.01-1.65) (Vestergaard 2002). A more recent epidemiological study of a population in
an industrialized area of Germany with relatively low intake of iodine has indicated that SCN"
urinary excretion is a cofactor or indicator for goiter in non-smokers as well as smokers. These
same authors found urinary ratios of iodide to SCN" to be predictive of increased risk for
development of goiter as compared to iodide status alone (Brauer et al., 2006). This study
population was believe to have high exposure to HCN due to industrial activities in the area,
though exposure levels of HCN were not presented.
4.2. SUBCHRONIC AND CHRONIC STUDIES AND CANCER BIOASSAYS IN
ANIMALS—ORAL AND INHALATION
4.2.1. Oral Studies
NTP (1993) reported the results of a subchronic bioassay of NaCN administered in
drinking water to rats and mice. F344 rats were administered NaCN in drinking water at
concentrations of 0, 3, 10, 30, 100, or 300 ppm for 13 weeks. These concentrations are
equivalent to the following doses, estimated by the study authors and based on measured body
weights and water consumption (converted to CN equivalents for this assessment): 0, 0.16, 0.48,
1.4, 4.5, or 12.5 mg/kg-day CN~ in male rats and 0, 0.16, 0.53, 1.7, 4.9, or 12.5 mg/kg-day CN~
in female rats. The parameters evaluated included body weight, clinical signs, water
consumption, clinical chemistry, hematology, urinalysis, extensive histopathology, selected
organ weights (heart, kidneys, liver, lungs, thymus gland, testes, epididymis, cauda epididymis),
testicular sperm measures (spermatid count and spermatid heads), epididymal sperm measures
(spermatozoa count and motility), and vaginal cytology. Thyroid weight or levels of thyroid
hormones were not evaluated in this study.
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In rats, no treatment-related effects on mortality or clinical signs of toxicity were seen in
either males or females. Body weight was statistically significantly decreased by 6% in high-
dose males, but this was not considered to be biologically significant by the study authors. No
body weight changes were observed in females. There was a dose-related decrease in water
consumption that was greater than 10% in both sexes exposed to 100 or 300 ppm NaCN,
compared with controls. Decreased urine volume and increased urine specific gravity were
observed in the high-dose male rats and were attributed to decreased water consumption.
Urinalysis data were not reported for females. Urinary thiocyanate concentration was
statistically significantly increased at drinking water concentrations of 30 ppm NaCN and higher
at the end of the study. There were no observed effects on nonreproductive organ weights in
males, but there was a statistically significant increase in absolute (16%) and relative (12%) liver
weights in high-dose females relative to controls. For all examined organs, there were no
histopathologic changes that were attributed to cyanide exposure. In particular, no histologic
effects were observed in either the thyroid or the brain. Female rats in the 100 and 300 ppm dose
groups (4.9 and 12.5 mg/kg-day CN , respectively) spent significantly more time in proestrus
and diestrus compared with controls, but there was no clear dose response and the authors did
not consider these results to be exposure related.
Male reproductive endpoints in the testis and epididymis were evaluated only in rats
exposed to >30 ppm NaCN (>1.4 mg/kg-day CIST). All reproductive parameter measurements
were conducted with the left reproductive organ (Table 4-4). In addition to evaluation of
epididymis weight, the weight of the cauda subsection of the epididymis was also measured.
This section of the epididymis functions as a site of sperm maturation and storage. Because the
cauda is part of the epididymis, these weights are not independent endpoints. Reproductive
organ weights were reported by NTP (1993) as absolute organ weights. For this review, relative
weights of reproductive organs were also calculated based on the individual animal data
downloaded from NTP's web site (http://ntp-server.niehs.nih.gov/) and were statistically tested
by using analysis of variance (ANOVA) followed by Dunnett's test.
NTP (1993) reported statistically significant decreases in several reproductive parameters
including epididymis weight, cauda epididymis weight, testis weight, number of spermatid heads,
testicular spermatid concentration, and epididymal spermatozoa motility. Absolute and relative
cauda epididymis weights were statistically significantly decreased at all doses examined
(>1.4 mg/kg-day CN~). In contrast, absolute epididymis weight was statistically significantly
depressed only at the highest dose (12.5 mg/kg-day). At the highest dose tested (12.5 mg/kg-
day), cauda epididymis weight (absolute) was decreased 13% below controls. The whole
epididymis weight (absolute) was significantly depressed 7% at this dose, and absolute testis
weight was significantly depressed 8%. Relative epididymis and testis weights were not
significantly different at any dose level. Additionally, standard histopathology did not
demonstrate any morphologic effects in any reproductive organ.
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Testicular spermatid parameters, including spermatid count and spermatid heads per
testis, were statistically significantly depressed at the highest dose tested (12.5 mg/kg-day). No
effect was seen on epididymal spermatozoa concentration; however, spermatozoa motility was
statistically significantly reduced at all tested concentrations (>1.4 mg/kg-day), although motility
did not exhibit a clear dose-related trend. At the lowest and highest dose, the percent of mobile
spermatozoa was reduced 4%, a magnitude of change within the range of historical controls and
not considered by the study authors to be biologically significant. Because fertility may be a
function of total spermatozoa count, rather than the concentration per gram cauda epididymal
tissue, and because decreased cauda epididymis weight can mask changes in spermatozoa
content, the number of total spermatozoa/cauda epididymis were also calculated for this review
(U.S. EPA, 1996). This number did not vary with dose. The unaltered spermatozoa count,
coupled with the decreased cauda epididymal weight, explained the slight dose-related (but not
statistically significant) increase in cauda spermatozoa concentration (see Table 4-4). For the
purpose of this review, a LOAEL of 1.4 mg/kg-day was identified, based on significantly
decreased relative and absolute cauda epididymis weights in male rats.
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Table 4-4. Reproductive effects in male rats administered NaCN in drinking
water for 13 weeks
Study parameter
0 ppm
30 ppm
100 ppm
300 ppm
Number of animals3
10
10
10
10
Dose (mg/kg-day)
0
1.4
4.5
12.5
Weights (g)
Body weight3
338 ±5
335 ±5
338 ±4
319 ± 5°
Epididymis, absolute
0.448 ± 0.006
0.437 ±0.005
0.425 ± 0.007
0.417 ± 0.005 d
(relative weight, 10 ')'
(13.7 ±0.14)
(13.3 ±0.21)
(13.0 ±0.18)
(13.3 ±0.41)
Cauda epididymis, absolute
0.162 ±0.003
0.150 ±0.004c
0.148 ±0.004c
0.141 ± 0.003 d
(relative weight, 10 ')'
(4.92 ± 0.05)
(4.54 ± 0.07)°
(4.51 ±0.12)d
(4.49 ± 0.09)d
Testis, absolute
1.58 ±0.03
1.56 ±0.02
1.52 ±0.02
1.46 ± 0.02d
(relative weight, 10 V'
(48.2 ± 0.64)
(47.3 ±0.58)
(46.4 ± 0.40)
(46.6 ± 0.80)
Testicular spermatid measurements
Spermatid heads (107/g testis)
11.35 ± 0.38
10.88 ±0.53
10.92 ±0.37
10.57 ±0.33
Spermatid heads (107/testis)
17.86 ±0.61
16.94 ±0.81
16.58 ±0.63
15.42 ±0.44c
Spermatid count (mean/10^1 mL
suspension)
89.28 ±3.05
84.68 ±4.03
82.90 ±3.16
77.10 ± 2.20°
Epididymal spermatozoa! measurements
Motility (%)
94.24 ±0.58
90.67 ± 1.25°
92.09 ±0.85c
90.66 ± 1.46c
Concentration (106/g cauda
epididymal tissue)
615 ±42
684 ± 40
699 ±33
709 ± 45
Spermatozoa count
(106/cauda epididymis)13
99.4 ±6.8
102.9 ±7.5
102.8 ±4.9
99.4 ±5.8
aData reported as mean ± standard error of the mean. Statistical significance determined by NTP, using Dunnett's
test (forbody weight only) or Shirley's test.
bCalculated for this assessment based on individual animal data available at
http://ntp-apps.niehs.nih.gov/ntp_tox/index.cfm. Statistical significance tested by one way ANOVA followed by
Dunnett's test.
Statistically different from control atp< 0.05.
Statistically different from control at p < 0.01.
Source: NTP (1993).
NTP (1993) also conducted a subchronic bioassay of NaCN administered in drinking
water in mice. B6C3F1 mice (10/sex/group) were administered NaCN in drinking water at
concentrations of 0, 3, 10, 30, 100, or 300 ppm for 13 weeks. These concentrations are
equivalent to the following doses, estimated by the study authors based on measured body
weights and water consumption (converted to CN equivalents for this assessment): 0, 0.26, 0.96,
2.7, 8.6, or 24.4 mg/kg-day in male mice and 0, 0.32, 1.1, 3.3, 10.1, or 28.8 mg/kg-day in female
mice. The parameters evaluated were identical to rats and included body weight, clinical signs,
water consumption, clinical chemistry, hematology, urinalysis, extensive histopathology,
selected organ weights (heart, kidneys, liver, lungs, thymus gland, testes, epididymis, cauda
epididymis), testicular sperm measures (spermatid count, spermatid heads), epididymal sperm
measures (spermatozoa count and motility), and vaginal cytology. Thyroid weight and level of
thyroid hormones were not evaluated.
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In mice, no significant treatment-related effects on mortality, body weight, or clinical
endpoints were observed. Water consumption in both males and females was decreased in the
mid- and high-dose groups. Absolute and relative liver weights were significantly increased by
18 and 23%, respectively, in the high-dose females, and relative liver weight (but not absolute
liver weight) was significantly increased in high-dose males (12%). However, there was no clear
dose response. No treatment-related effects were observed in clinical chemistry, hematology,
urinalysis, nonreproductive organ weights, or histopathology in any of the assessed organs.
Reproductive effects were only evaluated in mice exposed to the highest three doses (2.7 mg/kg-
day and higher). As in the rat component of this study, relative organ weights were not
originally reported by NTP but were calculated for this review and statistically analyzed by using
ANOVA followed by Dunnett's test.
In male mice, NTP (1993) found the absolute weights of the epididymis and cauda
epididymis to be statistically significantly decreased in the high-dose group (24.3 mg/kg-day)
relative to controls (Table 4-5). At the high dose, absolute epididymis and cauda epididymis
weights were reduced 10 and 18%, respectively. Relative cauda epididymis weight was
significantly decreased (18%) at 8.6 mg/kg-day. Relative epididymal and testis weights (relative
and absolute) were not statistically significantly decreased nor were sperm parameters
(spermatozoa per gram cauda epididymis, total spermatozoa per cauda epididymis, and
spermatozoa motility). No reproductive effects were reported at any of the dose levels tested for
female mice. For this review, a LOAEL of 8.6 mg/kg-day was determined based on a
statistically significant decrease in relative cauda epididymis weight.
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Table 4-5. Reproductive effects in mice administered NaCN in drinking
water for 13 weeks
Study parameter
0 ppm
30 ppm
100 ppm
300 ppm
Number of animals3
9
10
10
9
Dose (mg/kg-day)
0
2.7
8.6
24.3
Weights (g)
Body weight
37± 1.0
39.2 ± 1.3
38.6 ± 1.1
35.5 ±1.1
Epididymis
0.049 ±0.001
0.047 ± 0.002
0.047 ±0.001
0.044 ±0.001c
(relative weight, 10 ')'
(13.5 ±0.54)
(12.1 ±0.52)
(12.1 ±0.42)
(11.8 ±0.29)
Cauda epididymis
0.017 ±0.001
0.016 ±0.000
0.015 ± 0.001
0.014 ± 0.001c
(relative weight, 10 ')'
(4.74 ± 0.24)
(4.12 ±0.15)
(3.88 ± 0.22)°
(3.68 ± 0.17)d
Testis
0.121 ±0.002
0.113 ±0.008
0.117 ±0.002
0.118 ±0.003
(relative weight, 10 ')'
(33.4 ± 1.8)
(29.2 ± 2.2)
(30.3 ±0.90)
(31.7 ±0.92)
Testicular spermatid measurements
Spermatid heads (107/g testis)
18.47 ± 1.13
21.48 ±2.34
17.42 ± 1.34
18.17 ± 1.62
Spermatid heads (107/testis)
2.24 ±0.14
2.26 ±0.14
2.03 ±0.15
2.11 ± 0.16
Spermatid count (mean/10^1 mL
suspension)
69.94 ±4.34
70.80 ±4.25
63.28 ±4.53
66.06 ± 4.87
Epididymal spermatozoa! measurements
Motility (%)
92.38 ±0.81
90.63 ± 1.34
91.43 ±0.55
89.52 ±0.96
Concentration (106/g cauda
epididymal tissue)
1235 ±82
1393 ±70
1386 ±70
1462±101
Spermatozoa count
(106/cauda epididymis)13
21.2 ± 1.2
22.3 ± 1.3
20.5 ± 1.1
19.6 ±0.85
aData reported as mean ± standard error. Statistical significance determined by NTP using Dunnett's test (for body
weight only) or Shirley's test.
bCalculated for this assessment based on individual animal data available
at http://ntp-apps.niehs.nih.gov/ntp tox/index.cfm. Statistical significance tested by one way ANOVA, followed
by Dunnett's test.
Statistically different from control atp< 0.05.
Statistically different from control at p £ 0.01.
Source: NTP (1993).
As part of a study evaluating the effects of cyanogenic compounds in cassava, Kamalu
(1993) evaluated the toxicity of inorganic cyanide administered in a rice diet to male dogs
(six/group) for 14 weeks. The diet was supplemented at feeding time with NaCN at a dose
calculated to release 10.8 mg HCN per kilogram cooked food. Based on a reported daily food
consumption of 0.1 kg/kg body weight, this corresponds to 1.08 mg/kg-day HCN or 1.04 mg/kg-
day CNT. Further information about the study animals was not provided in this study, but animal
selection and study pretreatment were described in an earlier publication by the same author
(Kamalu, 1991), apparently describing the same study. That publication reported that dogs of
mixed breeds were purchased from local African markets at 6 weeks of age; treatment was
initiated when the dogs were approximately 22 weeks old. The authors noted that the dogs were
repeatedly treated for ecto-and endoparasites. It is unclear what impact the compromised health
status and repeated treatment for parasites had on the observed effects in the dogs. The basal diet
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used rice as the carbohydrate source, supplemented with pork, bone meal, and a vitamin and
mineral supplement that included iodine.
Blood was obtained from each dog at study weeks 1,3, and 14; urine was collected at
weeks 1, 3, 5, 7, and 14. Plasma and urinary thiocyanate concentrations were determined for
each collection period. Serum enzymes (including y-glutamyl transferase [GGT], ALT,
isocitrate dehydrogenase, total serum protein, serum albumin, serum globulin, and urinary
protein), as well as sodium (Na), magnesium (Mg), and phosphorus (P), were measured.
Histopathologic evaluation was performed on the liver, kidneys, heart, testes, and adrenal glands
of each dog. The thyroid gland was not evaluated. At all time points evaluated, both plasma and
urinary thiocyanate concentrations were significantly increased in the treated dogs compared
with controls. Relative to controls, treated dogs had significantly increased urinary protein
concentration at weeks 5 and 14. No treatment-related effects were observed in serum enzymes,
total serum protein, albumin, or globulin or in Na, Mg, and P concentrations. No histopathologic
changes were observed in the liver or heart of treated dogs; however, treatment-related effects
were observed in the kidneys, testes, and adrenal glands. Kidneys of the treated dogs had casts
in the lumens of the renal tubules, accompanied by occasional desquamation. In the testes,
specialized reproductive morphologic analysis indicated that the treated dogs had a significantly
decreased percentage of tubules in stage VIII of the spermatogenic cycle (characterized by
elongated spermatids lining the lumen of the seminiferous tubules) as compared with controls
(p < 0.01). This percentage was 1.6 ± 1.07% (mean ± standard error of the mean [SEM]) in the
treated group compared with 14.4 ± 0.94% in the controls. Treated dogs also had an increased
incidence of abnormal cells and sloughing of germ cells in the seminiferous tubules.
Hyperplasia and hypertrophy were observed in the adrenal gland. The adrenal medulla was
unaffected by inorganic cyanide treatment. Although the width of the adrenal cortex did not
differ significantly between the cyanide-treated and control groups, the zona glomerulosa (the
most superficial layer of the adrenal cortex) was significantly wider in treated dogs. The results
of this study indicate that cyanide may be a reproductive toxicant in male dogs. Based on
histopathologic changes in the kidneys, testes, and adrenal glands, the only dose tested
(1.04 mg/kg-day) was considered to be a LOAEL.
An evaluation of the thyroid from this study was presented in Kamalu and Agharanya
(1991). At week 14, serum T3 was significantly decreased by 55%, and thyroid weight was
significantly increased by 23% in the cyanide-exposed group. A histopathologic evaluation of
the thyroid gland found decreased colloid content compared to that of controls.
A 40-week study in New Zealand white rabbits that reported both liver and kidney
lesions supports the kidney as a possible target organ for toxicity, following exposure to cyanide
(Okolie and Osagie, 1999). In this study, groups of six male rabbits were fed a diet of growers'
mash only or mash containing 702 ppm CN~ as KCN for 10 months. Based on daily food
consumption and weekly body weight measurements, the study authors estimated that the
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average CN intake was 0.39 mg/day per rabbit in the control group and 36.5 mg/day per rabbit
in the exposure group. Initial and final body weights were averaged to estimated daily doses of
0.2 and 20 mg/kg-day, respectively. Decreased body weight (33%) and decreased food
efficiency were observed in the high-dose group (33%). At the end of 10 months of treatment,
serum levels of ALT, alkaline phosphatase (ALP), lactate dehydrogenase (LDH), and sorbitol
dehydrogenase were increased. ALP levels were reduced in the lung but not in the heart. LDH
was increased in the liver and kidneys, a finding that the study authors interpreted as indicative
of a shift from aerobic to anaerobic metabolism, thereby increasing the production of lactic acid.
Biochemical evidence of tissue injury in the liver and kidney was supported by histopathologic
findings of focal areas of hepatic necrosis and congestion and renal tubular and glomerular
necrosis. No abnormal histopathology was reported for the pancreas or the heart. Neither a full
list of tissues examined nor additional information on histopathologic changes in other organs
was provided in the study. However, the occurrence of focal pulmonary edema and necrosis in
treated rabbits was reported in a second paper on the same study (Okolie and Osagie, 2000).
Based on necrosis in the liver and kidney, the only dose tested (20 mg/kg-day) is considered for
this review to be a LOAEL.
Manzano et al. (2007) examined the effects of subchronic (70-day) KCN ingestion in
45-day-old Lanrace-Large white pigs. The number and sex of animals used in this study are
unclear since study details in the published report are conflicting, with indications of 6 animals
per group in the materials and methods section but 5-10 animals indicated per group in the tables.
Animals were administered KCN in the diet, twice per day, for total daily doses of 0, 2, 4, or
6 mg/kg-day (0, 1.4, 2.8, or 4.3 mg/kg-day GST). Blood samples were collected prior to the
experimental period and then every week thereafter and analyzed for ALT, glucose, cholesterol,
blood urea nitrogen, creatine, T3, T4, and thiocyanate. At the conclusion of the experiment,
thyroid glands were weighed, and tissues from the CNS, thyroid, pancreas, liver, and kidneys
were examined histologically. Significantly decreased serum ALT was observed at >1.4 mg/kg-
day. Additionally, significantly increased urea and creatinine were observed at doses
>2.8 mg/kg-day. Thyroid weight was significantly increased 24% in animals in the highest dose
group, though significant alterations in thyroid hormones were not observed. Histological
alterations of the thyroid gland, characterized by numerous vacuoles in the colloid of the thyroid
follicles, were observed in all dosed animals. The authors also reported histologic alterations in
the liver, kidney, and CNS. Liver lesions were reported as karyolysis and pyknosis (nuclear
DNA changes denoting cell death) and distortion of the normal lobular architecture.
Degeneration of the renal tubular epithelial cells was reported in the kidney. In the brain,
minimal degeneration of Purkinje cells and loss of cerebellar white matter were reported. All
histologic lesions were reported by the authors to occur in a dose-related fashion, though neither
incidence nor statistical analysis of these findings was presented. A LOAEL of 4.3 mg/kg-day
based on increased thyroid weight was determined for this review.
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Jackson (1988) evaluated the effects of oral administration of KCN on behavior and
thyroid function in miniature pigs. Doses of 0, 0.4, 0.7, or 1.2 mg/kg-day KCN were
administered to three pigs per group via gavage. A total of five females and seven males were
used; each dose group contained both male and female animals (one dose group contained two
females and one male while the others contained two males and one female). The solutions were
administered once daily for 24 weeks, prior to feeding, in order to increase the gastrointestinal
absorption of cyanide. Regularly measured serum thiocyanate levels were positively correlated
with cyanide dose. Serum levels of T3, T4, and glucose were measured every 6 weeks.
Behavioral evaluations were conducted daily. Two categories of behavior were evaluated:
performance measures, including innate behavior, and learning measures, including the
acquisition and retention of new behaviors. No other endpoints were evaluated. Changes in
thyroid hormones were portrayed graphically as means without reporting variance (SEM or
standard deviation [SD]) or individual animal data. Individual animal data were requested from
the study authors by EPA but were not provided. Both T3 and T4 demonstrated a dose-related
decrease (23 and 13%, respectively) that was statistically significant by week 18 of the study.
Thyroid histopathology was not evaluated in this study. A variety of behaviors were
significantly altered in treated animals, including a decrease in dominance behavior (high-dose
group), a decrease in fighting (mid- and high-dose group), an increase in flight response (all
treated groups), a decrease in exploratory behaviors (all groups), and less aggressive feeding
patterns (high-dose group). The authors concluded that the overall pattern of behavioral changes
in the group administered 1.2 mg/kg-day was different from that of the control animals but that
the changes at lower doses were inconsistent. This study supports the large body of evidence
demonstrating that the thyroid is a target organ for cyanide toxicity. Based on reported
behavioral changes and decreased thyroid hormones, the LOAEL and no-observed-adverse-
effect level (NOAEL) identified for this review are 1.2 and 0.7 mg/kg-day CN , respectively.
Philbrick et al. (1979) evaluated the long-term health effects of oral exposure to cyanide
in rats. Male rats from Woodlyn Laboratories (10/group, strain not specified) received diets
(10% casein supplemented with 0.3% methionine, potassium iodide, and vitamin B12) containing
either 0 or 1,500 ppm KCN or an equal molar amount (2,240 ppm) of potassium thiocyanate
(KSCN) for 11.5 months. Parallel studies were conducted with rats provided a diet deficient in
methionine, iodine, and vitamin B12, containing 0 or 1,500 ppm KCN (44 mg/kg-day CN-)2. At
4 and 11 months, plasma T4 levels, T4 secretion rates, and urinary thiocyanate levels were
measured in five animals per group. After sacrifice, brain, heart, liver, and thyroid weights were
recorded. Histopathologic evaluation was conducted on the brain, optic and sciatic nerves, spinal
cord, and thyroid gland. This study design, although limited by the use of only one dose level,
allowed for the comparison of effects mediated directly through cyanide versus the primary
2 Based on the average food intake across rat strains (U.S. EPA, 1988) and adjusting for the molecular weight ratio
of cyanide to potassium cyanide.
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metabolite, thiocyanate. It also allowed for the comparison of cyanide and thiocyanate treatment
in control rats compared to rats fed nutritionally restricted diets.
Body weight gains of the KCN-treated animals were significantly lower than those of
controls, beginning at week 8; administration of KSCN did not affect body weight. Urinary
thiocyanate levels were reported as |ig/g food ingested (instead of |ig/mL urine) and were
significantly higher than in controls in all treated groups. Urinary thiocyanate levels were lower
at 11 months than at 4 months of treatment in KCN- but not KSCN-treated groups. This appears
to indicate reduced metabolism of KCN with chronic exposure. Additionally, though animals
were fed an equal molar amount of KCN and KSCN, at 11 months urinary levels of SCN in the
KCN-treated animals were only approximately one quarter of the urinary excretion of SCN~ in
KSCN-treated animals. This appears to indicate only partial metabolism of cyanide into
thiocyanate. Administration of cyanide altered serum T4 levels at 4 months but not at 11 months;
thiocyanate altered serum T4 levels at both 4 and 11 months. After 4 months of treatment, rats in
the cyanide-exposed groups had significantly decreased plasma T4 levels (53%) and decreased T4
secretion rates (68%) compared to controls; however, after 11 months of cyanide treatment, T4
levels no longer differed from those of controls, though T4 secretion rates were depressed 27%.
KSCN-treated animals also showed a significant reduction (62%) in T4 secretion rates (at
4 months but not at 11 months) and decreased T4 levels (55% at 4 months and 26% at
11 months). At the termination of the study, relative thyroid weights were significantly
increased in both KCN- and KSCN-treated animals by 43 and 33%, respectively. In the
nutritionally restricted control animals, levels of T4 and T4 secretion rates were lower compared
to controls fed a standard diet. However, alterations of T4 levels, T4 secretion rates, and thyroid
weights in animals on the restricted diet treated with KCN and KSCN were of similar magnitude
compared to treated animals on the standard diet. No histopathologic lesions were observed by
light microscopy in the optic or sciatic nerves or thyroid gland of any group. Increased
vacuolation was observed in the spinal cord white matter of treated animals (with both KCN and
KSCN) receiving sufficient or deficient methionine compared to controls, and spinal cord
demyelination induced by methionine deficiency was exacerbated by treatment. No information
on incidence or severity of the observed histologic lesions was reported by the authors. No
measurable differences were reported in spinal cord pathology between cyanide- and
thiocyanate-treated rats. Based on increased thyroid gland weight, decreased thyroid hormone
levels, and histopathologic changes in the spinal cord, the single dose tested (44 mg/kg-day CN )
is considered to be a LOAEL.
Howard and Hanzal (1955) conducted a 2-year dietary study in which 10 Carworth Farms
rats per sex per group were administered food fumigated with HCN. The authors indicated that
only rats surviving to the end of the study were analyzed histologically because the accuracy of
necropsies performed on animals that died early were compromised by autolysis. It appears that
seven, five, and nine males and six, seven, and six females were examined histologically in the 0,
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4.3, and 10.8 mg/kg-day dose groups, respectively. Although special feeding jars were used to
minimize air circulation and evaporation, the study authors noted that it was necessary to
measure the loss of HCN due to evaporation from the chow and to prepare fresh rations every
other day to keep the HCN concentration near the target values of 100 and 300 ppm (milligrams
HCN per kilogram diet). The average daily concentrations were 73 and 183 mg CN per kg diet.
These average concentrations of cyanide in the food were estimated based on Howard and
Hanzal's (1955) data for concentrations at the beginning and end of each food preparation period
and assuming a first-order rate of loss for the intervening period (U.S. EPA, 1992). From the
data reported on food consumption and body weight, estimated doses were 0, 4.3, and 10.8
mg/kg-day. There were no treatment-related effects on growth rate, no gross signs of toxicity, no
hematologic effects, and no histopathologic lesions in the tissues evaluated from an undisclosed
subset of animals (heart, lungs, liver, kidneys, spleen, stomach, small and large intestines,
adrenals, thyroid, testes or uterus and ovaries, cerebrum, and cerebellum). Histopathology of the
spinal cord was not examined. Howard and Hanzal (1955) also reported that there appeared to
be no effect on relative organ weight of the liver, kidneys, spleen, brain, heart, adrenals, testes, or
ovaries. The highest dose tested, 10.8 mg/kg-day, is considered to be the NOAEL.
Two studies by Soto-Blanco et al. (2002a, b) provide evidence of neurological changes
associated with cyanide ingestion in rats and goats. In the first study, Soto-Blanco et al. (2002a)
evaluated the toxicity of KCN administered daily by gavage (vehicle not stated) to male Wistar
rats for 12 weeks. Administered doses of KCN were 0, 0.15, 0.3, or 0.6 mg/kg-day, equivalent
to 0, 0.06, 0.12, or 0.24 mg/kg-day CN~, respectively. The number of animals included in each
group was seven, six, six, and seven for the control, low-, mid-, and high-dose groups,
respectively. Endpoints evaluated included clinical signs of toxicity, body weight, food
consumption, serum cholesterol, glucose, T3, and T4. Histopathologic examination was limited
to the CNS, thyroid gland, and pancreas. No treatment-related effects were reported for clinical
signs of toxicity, body weight gain, food consumption, serum T3 and T4, or serum glucose.
Plasma cholesterol was significantly decreased in the high-dose group. No histopathologic
changes were observed in the thyroid gland or the pancreas. Reported CNS effects in the high-
dose group included neuron loss in the hippocampus, damaged Purkinje cells (further details not
reported) and loss of white matter in the cerebellum, and the occurrence of a dose-related
increase in spheroid bodies on white matter in the spinal cord. EPA was unsuccessful in
obtaining incidence data from the study authors for the observed histologic lesions. Because
quantitative information was not reported on these histologic observations, neither a LOAEL nor
a NOAEL could be determined from this study.
Soto-Blanco et al. (2002b) also evaluated the neurotoxicity of cyanide, administered daily
as KCN, to male Alpine-Saanen goats in milk or water for 5 months at doses of 0, 0.3, 0.6, 1.2,
or 3.0 mg/kg-day (equivalent to 0, 0.12, 0.24, 0.58, and 1.2 mg/kg-day CIST). The test compound
was administered twice daily in milk (one-half of the daily dose per treatment) for the first
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3 months and in water for the remainder of the treatment period. The number of animals per
dose group ranged from six to eight. The CNS was evaluated histologically for the presence of
glial fibrillary acid protein (GFAP), a marker for glial cells. The only clinical signs of toxicity
were transient muscular tremors and ataxia in one high-dose animal on days 121-123 of the
study. Neuropathology, including congestion, hemorrhage, and gliosis in the cerebellum, spinal
cord, and pons, as well as spheroids on the gray matter of the spinal cord, was observed at 0.58
and 1.2 mg/kg-day CIST. Additional findings in the high-dose group included damage and loss of
Purkinje cells in the cerebellum, spongiosis in the pons, and spheroids, axonal swelling, gliosis,
spongiosis, and ghost cells in the medulla oblongata. GFAP immuno-staining confirmed the
gliosis observed by histopathology. While this study confirmed that the CNS is a target organ of
subchronic cyanide administration, no information on the incidence or severity of histologic
findings was reported. Therefore, a NOAEL or LOAEL could not be determined from this study.
4.2.2. Inhalation Studies
No chronic or subchronic animal studies of cyanide inhalation exposure were located.
However, several subchronic inhalation studies of related compounds, including cyanogen (CN)2
and ACH, are available. A 6-month inhalation study in monkeys (5 males/group) and rats (30
males/group) exists for the gas (CN)2. (CN)2 is thought to break down in aqueous solution to
CN and OCN ions (Cotton and Wilkinson, 1980). Lewis et al. (1984) exposed three groups of
male rhesus monkeys or male albino rats (Sprague-Dawley) to 0, 11, or 25 ppm (CN)2
6 hours/day, 5 days/week for 6 months. This would be equivalent to 0, 12, or 28 mg/m3 HCN
(based on the creation of 1 mol CN~ per mol (CN)2 in water). Pathology evaluated for both
monkeys and rats included gross and microscopic examination of heart, liver, kidney, cerebellum,
lungs, thyroid, spleen, and bone marrow. Hemoglobin, hematocrit, T3, and T4 were also
evaluated. Additionally, behavioral tests and electrocardiograms were administered to the
monkeys. No significant changes were seen in monkeys other than decreased lung moisture
content in both dose groups. The only effect noted in rats was significantly depressed body
weight (13%) in the high-dose group.
Inhalation studies of subchronic ACH exposure in male and female rats are available.
ACH is a liquid at room temperature, with a boiling point of 95°C and a vapor pressure of
0.75 mm Hg. In neutral to basic aqueous environments, ACH is reported to dissociate readily to
acetone and cyanide (U.S. EPA, 1985). Sprague-Dawley rats (15/sex/group) were exposed by
inhalation at average concentrations of 10.1, 28.6, or 57.7 mg/m3 ACH 6 hours/day, 5 days/week
for 14 weeks (Monsanto Co., 1985a, b). This dose corresponds to the molecular equivalent of
HCN concentrations of 3.2, 8.8, and 18.2 mg/m3. Endpoints analyzed included hematology,
clinical chemistry (including T3 and T4 levels), and gross and microscopic histopathology on a
wide range of organs and tissues. No effects on mortality, body weight, or behavior were
observed in treated animals. Blood and urine levels of thiocyanate were elevated in a dose-
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dependent manner, although no alterations in T3 or T4 were observed. No significant gross or
microscopic histology was observed in treated animals compared with controls. In summary, the
study authors found no gross signs of toxicity attributable to subchronic inhalation exposure to
ACH in rats. Reproductive and developmental studies by the inhalation route have also been
conducted for ACH (Monsanto Co., 1985a, b; IRDC, 1984) and are described in section 4.3.2.
4.3. REPRODUCTIVE/DEVELOPMENTAL STUDIES
4.3.1 Oral Studies
Imosemi et al. (2005) fed 20 pregnant female Wistar rats 0 or 500 ppm KCN (equivalent
to 20 mg/kg-day) in the diet during gestation and up to postnatal day (PND) 50. Offspring
(five/group) were killed on PNDs 1, 9, 14, 21, 28, and 50, and the cerebellar tissues were
examined grossly. Parameters examined included body weight, brain weight, cerebellar weight,
maximum vermis length (length between cerebellar hemispheres), maximum side-to-side
dimensions of the cerebellum, and maximum thickness (anteroposterior dimension) of the
cerebellum. Aggressive and restless behavior was noted in the exposed dams but not in controls.
Additionally, significantly decreased body weight (6%) and brain weight (19%) were observed
in the treated pups on PNDs 14 and 9, respectively. No significant changes in body weight or
brain weight were found at the additional five time points examined. Cerebellar weight was
significantly reduced on PNDs 14, 21, and 28. The maximum vernal length was significantly
reduced on day 50 and the maximum side-to-side width of the cerebellum was reduced on day 29.
In a separate publication, the authors also reported on microscopic parameters of the cerebellum
(Malomo et al., 2004). A significantly thicker external granular layer (EGL) was seen in the
experimental group on PNDs 14 and 21. Reduced thickness of the molecular layer (ML) was
also observed on PNDs 28 and 50. The density and size of the Purkinje cells were not different
between groups. Additionally, staining of the white matter was similar between groups,
suggesting normal myelination. The authors concluded that maternal consumption of 20 mg/kg-
day CN did not significantly affect microscopic indicators of cerebellar development but caused
mild changes later in postnatal life. The authors suggested that the presence of a thicker EGL
layer in the experimental group suggested delayed maturation and migration of cells in the
cerebellum. A LOAEL of 20 mg/kg-day CN~ was determined for this review based on altered
maturation of the cerebellum.
Soto-Blanco and Gorniak (2004) evaluated effects of gestational exposure to cyanide in
pregnant mixed-breed goats (six per group). Starting on day 24 of pregnancy, goats were
administered, by gavage, 0, 1, 2, or 3 mg/kg-day KCN (equivalent to 0, 0.4, 0.8, or 1.2 mg/kg-
day CIST) until parturition (day 150). Blood samples were collected every other week and
analyzed for plasma glucose, cholesterol, and thiocyanate. T4 and T3 concentrations in plasma
were measured from the offspring at birth and at 1 week old. One control dam and one dam
from the highest dose group were sacrificed at day 120. Three months after birth the male
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offspring and one dam from each group were sacrificed and the pancreas, thyroid, and entire
CNS (including spinal cord) were collected for histologic examination. Two dams from the
highest dose group experienced clinical signs of cyanide intoxication, specifically ataxia and
convulsions. Cyanide treatment did not significantly alter the number of live offspring or the
length of gestation, though average length of gestation in all treated groups was about 2.5 days
shorter than controls. T3 levels in dams and offspring tested at birth were significantly elevated
over controls in the highest dose group, while T4 levels did not appear to be different. The dam
sacrificed at day 120 of pregnancy revealed increased reabsorption vacuoles in the thyroid
follicular colloid and severe spongiosis of the cerebral, internal capsule, and cerebellar peduncle
white tracts, suggestive of myelin edema of the white matter. The histopathologic study of dams
and offspring 3 months after birth revealed no lesions. Due to the lack of incidence and severity
data for the observed histologic effects, a LOAEL could not be determined from this study.
Soto-Blanco and Gorniak (2003) dosed mixed-breed, lactating goats (six per group) with
0, 1.0, 2.0, or 3.0 mg/kg-day KCN (equivalent to 0, 0.4, 0.8, or 1.2 mg/kg-day CIST) by gavage
(in water) from lactation days 0 to 90 and measured whole blood cyanide and thiocyanate
concentrations in dams and offspring on lactation days 30, 60, and 90. On the 90th day, glucose,
cholesterol, plasma urea nitrogen, creatinine, T4, T3, AST, ALT, and GGT were also determined.
After 90 days, one dam from each dose group and every male goat from all litters was killed.
The pancreas, thyroid glands, liver, kidneys, and the whole CNS were collected for histologic
examination. No clinical signs of toxicity were seen in any group, although one dam in the
highest dose group died on the 55th day of lactation. Both whole blood cyanide and plasma
thiocyanate concentrations were increased in a dose-dependent manner in treated dams. In the
offspring, both blood cyanide and plasma thiocyanate increased with increasing maternal cyanide
dose, peaking at lactation day 30, and decreased with lactation time. Plasma parameters in all
groups of dams and offspring appeared to be unaffected by KCN treatment, except for the level
of T4 in dams, which was significantly elevated (20%) over controls (p < 0,01 by a t-test
conducted for this review). T3 and T4 levels also appeared elevated in the high-dose animals,
although these differences were not significant. In the thyroid, histopathologic changes,
characterized by an increased number of reabsorption vacuoles on the colloid of the thyroidal
follicles were observed in dams and offspring. Additionally, histologic changes in the liver and
kidney were noted, characterized by hepatocellular vacuolization and degeneration and mild
vacuolization of tubular epithelial cells, in dams and offspring. The authors noted that observed
histologic lesions were most intense in the highest KCN dose group. No histologic lesions were
noted in other examined tissues. In the absence of incidence data or statistical analysis on any
histologic changes, a LOAEL was not determined for this study.
Tewe and Maner (1981) fed female rats (20 per group, strain not specified) either a basal
diet prepared from low-HCN cassava meal or the basal diet supplemented with 500 ppm of KCN
throughout mating, gestation, and lactation. In addition, two female weanling rats per litter were
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maintained on each diet for 28 days following weaning. Adult rats on the basal diet alone
received a dose of 1.2 mg/kg-day (based on a dietary HCN concentration of 12 mg/kg and
average food intake among female rats of 0.102 kg/kg body weight). Adult rats on the basal diet
plus 500 ppm KCN (high-cyanide diet) received a total CN dose of 21.6 mg/kg-day, including
the 1.2 mg/kg-day from the basal diet and 20.4 mg/kg-day as KCN. For the weanling rats, the
corresponding doses were approximately 1.9 mg/kg-day for the basal diet and 34.3 mg/kg-day
for the basal + KCN diet, based on average food intake (0.162 kg/kg body weight) for female
weanling rats (U.S. EPA, 1988). As compared with controls, the high-cyanide diet had no effect
on body weight of pregnant rats, food consumption, maternal liver or kidney weights, litter size,
birth weight of pups, or pup mortality. In the weanling rats, the high-cyanide diet resulted in
significant decreases in food consumption and growth rate and an increase in the ratio of food
consumption to body weight gain, indicating that the decreased weight gain was not due solely to
poor palatability. The high-cyanide diet also resulted in a significant increase in serum
thiocyanate in both dams and weanlings compared with animals on the basal diet alone. The
activity of rhodanese, the enzyme that metabolizes cyanide to thiocyanate, in the liver and
kidneys was comparable in all groups. A LOAEL of 34.3 mg/kg-day was identified from this
study, based on decreased daily weight gain in weanlings; a NOAEL of 1.2 mg/kg-day (in
adults) and 1.9 mg/kg-day (in weanlings) was determined (for this review) based on cyanide
content of the basal cassava diet.
Teratogenicity has been reported in some nontraditional developmental studies that
administered cyanide or cyanogenic foods to animals. Doherty et al. (1982) administered NaCN
to hamsters subcutaneously via osmotic minipumps at doses around 0.13 mmol/kg-hour
(approximately 80 mg/kg-day) from gestation days (GDs) 6-9 and observed fetotoxic effects,
including significantly increased fetal resorptions and malformations and decreased crown-rump
length. Doses utilized in this study covered a narrow range from about 78-81 mg/kg-day. At
these doses, clinical signs of toxicity were evident in the dams, including weight loss, ataxia, and
dyspnea. At the lowest dose tested (78 mg/kg-day), there were 63% resorptions as compared to
10% in controls. Additionally, at this dose, 62% of fetuses were malformed vs. 5% of controls.
The majority of malformations in treated groups were characterized as neural tube defects.
Coadministration of cyanide with the cyanide poisoning antidote sodium thiosulfate, which
serves as a sulfur donor in the conversion of cyanide to thiocyanate by the enzyme rhodanese,
protected against maternal toxicity and teratogenic effects. Another developmental study by
Frakes et al. (1986) observed reduced ossification and decreased body weight in offspring of
hamsters administered a protein-deficient cassava diet containing low- or high-cyanide levels
during days 3-14 of gestation. The developmental effects of dietary cyanide were not evaluated
in animals fed a protein-sufficient diet. Low (mean: 0.65 mmol CN/kg food) and high (mean: 8
mmol CN/kg food) cassava-containing diets equaled daily doses of approximately 1.3 or
14 mg/kg-day cyanide, respectively, and averaged only 4% protein (the standard laboratory diet
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contained 25% protein). Body weight in the cassava-fed dams was approximately 30% lower
than in control dams fed a standard, protein-sufficient diet, regardless of whether they were in
the high- or low-cyanide group. The numbers of implantations, resorptions, live fetuses, and
malformed fetuses in cyanide-treated groups were not statistically different from those in
controls. Fetal body weight was significantly decreased by 14 and 8% in low- and high-cyanide
treatment groups, respectively (compared with the low-protein controls). Significantly decreased
ossification centers (28—37%) were observed in portions of the fetal skeletons, including the
sacrocaudal vertebrae, metatarsals, and sternebrae. No dose-related trend was observed in the
decrease in maternal body weight, decrease in fetal body weight, or decrease in ossification
between the low- and high-cyanide dose groups.
A Japanese study (Amo, 1973) reported that 0.05 mg/kg-day of cyanide administered in
drinking water decreased the fertility and survival rate in the F1 generation and produced 100%
mortality in the F2 generation in mice. Although no other studies exist on F2 animals treated
with cyanide, the data presented by Amo (1973) on decreased survival of the F1 generation are
not consistent with the body of available literature for cyanide, which indicate no decrease in
survival of the F1 generation of goats treated gestationally with doses twice as high (Soto-Blanco
and Gorniak, 2004) or in rats treated gestationally with doses >20 mg/kg-day (Imosemi et al.,
2005; Tewe and Maner 1981). Additionally, studies in rats with ACH, which breaks down into
cyanide and acetone following inhalation or oral exposure, have not observed decreases in
reproductive parameters or F1 survival at inhalation exposures equivalent to 66 mg/m3 HCN
(Monsanto Co., 1985a, b). Furthermore, gavage dosing of pregnant Sprague-Dawley rats with
doses of ACH equivalent to 3 mg/kg-day during GDs 6-15 did not decrease survival in offspring
compared with controls. Nor was there any difference in number of viable fetuses,
postimplantation losses, mean fetal body weight, fetal sex distribution, and fetal malformations
between treated animals and controls (IRDC, 1984).
4.3.2 Inhalation Studies
No studies exist on the potential reproductive or developmental toxicity of inhaled
cyanide. However, male and female fertility indices were investigated in rats exposed via
inhalation to the cyanide precursor ACH, which decomposes to acetone and cyanide (IPCS,
2005). At room temperature, ACH is primarily a liquid (boiling point: 95°C); however, in these
inhalational studies, the ratio of target and analytical air concentrations were close to unity,
indicating that ACH was primarily present as a vapor.
In a male fertility study (Monsanto Co., 1985a), Sprague-Dawley rats (15/group) were
exposed by inhalation to ACH at 0, 35, 104, or 209 mg/m3 for 6 hours/day, 5 days/week over a
period of 69 days. These doses are equivalent to 0, 11, 32, or 65 mg/m3 HCN. Following the
exposure period, males were mated with three nonexposed females each. Pregnant females were
sacrificed at mid-gestation (GDs 13-15), and pre- and postimplantation losses were determined.
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Males were sacrificed 3 weeks following cessation of exposure. Histologic analysis of
reproductive organs, including the testis, epididymis, prostate gland, and seminal vesicle, was
conducted; reproductive organ weight and sperm parameters were not evaluated. No treatment-
related differences were seen in mean body weight, clinical chemistry, or histology of treated
males. The mating efficiency, number of live implants, and pre- and postimplantation losses
were not different between treated and control groups.
Female Sprague-Dawley rats (24/group) were exposed by inhalation 6 hours/day,
7 days/week for 21 days to 0, 38, 108, or 207 mg/m3 ACH (0, 12, 33, or 64 mg/m3 HCN) and
then mated with untreated males (Monsanto Co., 1985b). Exposure of the females was
continued until the day of mating, and the females were sacrificed at mid-gestation (GDs 13-15)
to determine pregnancy status, nidations, pre- and postimplantation loss, and histology of the
ovaries and uteri. No clinical signs of toxicity were observed in treated animals except for dose-
related observations of red nasal discharge or encrustation in some animals. No treatment-
related differences were seen in mean body weight, clinical chemistry, or histology. Mating
efficiency, pregnancy rates, number of live implants, and pre- and postimplantation losses in
treated animals were comparable to control values.
4.4. OTHER DURATION- OR ENDPOINT-SPECIFIC STUDIES
4.4.1. Acute Oral Studies
In evaluating the oral toxicity of cyanide, both the total amount administered and the rate
of absorption are important (U.S. EPA, 1992), because toxicity results from exceeding the
body's capacity for detoxification of cyanide, which occurs mainly in the liver. At high doses of
cyanide, the availability of the sulfur donor needed for detoxification by the enzyme rhodanese
can become rate limiting. If absorption of ingested cyanide proceeds too quickly, the capacity of
the liver to form thiocyanate via first-pass metabolism may be exceeded. In contrast, slow
absorption of the same total oral load of cyanide may allow complete metabolism by the liver.
Similarly, an acute cyanide dose is more toxic when administered by inhalation compared with
the same dose administered by ingestion, because the inhalation route bypasses first-pass
metabolism in the liver and directly enters systemic circulation.
The significant impact of absorption on the rate of detoxification of cyanide is
responsible for the observation that lethal dose (LD50) values for NaCN (presented below) are
lower than the acute and chronic LOAEL values. The LD50 values are based on bolus doses that
result in rapid absorption of a large amount of cyanide that overwhelms the detoxification
capacity of the body. In contrast, the acute and chronic LOAEL values are based on cyanide
administration at a lower dose rate over the course of a day. This slower dose rate means that the
body is able to detoxify higher doses of cyanide (on the basis of administered mg/kg) without
being overwhelmed, and thus it can handle a higher total dose load.
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Acute oral LD50 values for cyanide in rats range from 3 mg/kg-day (Ballantyne, 1988) to
8 mg/kg-day (Smyth et al., 1969) for cyanide administered as NaCN. Single daily doses of 4
mg/kg-day in rats and 6 mg/kg-day in mice as KCN resulted in 95% mortality (Ferguson, 1962).
Dermal LD50 values in rabbits range from 4.1 to 8.9 mg/kg. Clinical signs observed following
single dermal doses ranging from 0.9 to 2.5 mg/kg include rapid breathing, dizziness, weakness,
convulsions, and loss of consciousness (ATSDR, 2006). Inhalation LC50 values reported in
animals range from 151 to 579 mg/m3 HCN in various species (ATSDR, 2006).
Palmer and Olson (1979) administered KCN to groups of seven male Sprague-Dawley
rats at 0 or 200 ppm in drinking water or at 0 or 200 ppm in feed for 21 days. Using an average
body weight of 0.12 kg and a water consumption rate of 0.17 L/kg-day as per U.S. EPA (1988),
this assessment estimated the daily CN intake to be approximately 14 mg/kg-day. The only
endpoints evaluated were body weight gain and liver weight. A statistically significant 17%
increase in absolute liver weight was observed relative to controls; thus, the LOAEL was the
single dose tested, 14 mg/kg-day.
The dietary part of this study was inadequate for evaluation of toxicity due to instability
of cyanide concentrations in feed. Although CIST ingestion was estimated at 9 mg/kg-day by
using default assumptions for body weight and food consumption (U.S. EPA, 1988), the study
authors noted that subsequent analysis of cyanide in feed resulted in <20% recovery of the
predicted value compared to 95% recovery for cyanide added to feed immediately prior to
analysis. Therefore, due to uncertainties regarding actual animal dosage, a LOAEL could not be
determined from the dietary part of this study.
Sousa et al. (2002) administered KCN to adult male Wistar rats in drinking water at target
doses ofO, 0.3, 0.9, 3.0, or 9.0 mg/kg-day for 15 days, equivalent to CN~ doses ofO, 0.12, 0.36,
1.2, and 3.6 mg/kg-day. There were 10 rats/group, except for the high-dose group, which
included 6 rats. Weight gain was significantly decreased at the high dose to about a third that of
the controls; however, weight gain was normal in the next lower dose group. There were no
effects on serum levels of T3, T4, and serum levels of ALT, while AST exhibited sporadic
statistically significant changes that were determined not to be dose related. Serum levels of urea
or creatinine were unaffected by treatment. "Moderate" to "severe" congestion and cytoplasmic
vacuolization of the epithelial cells of the proximal tubules were observed in the kidneys of rats
at the two highest doses. Hydropic degeneration of hepatocytes was also noted at the highest
dose. Reabsorption vacuoles were observed in the thyroid gland of animals in all groups,
including controls, and increased in severity with increasing dose. However, quantitative
incidence or severity data for the histopathologic observations were not reported. Based on
moderate kidney vacuolization and congestion, a NOAEL of 0.36 mg/kg-day and a LOAEL of
1.2 mg/kg-day CN~ were identified from this study.
Kreutler et al. (1978) evaluated the short-term effects on the thyroid of oral exposure to
cyanide. Male weanling albino rats (strain not specified; 10-24 animals/group) were fed diets
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containing either low protein (2% casein) or normal protein (20% casein) for 2 weeks; treated
rats received the same diets supplemented with 0.2% KCN (equivalent to 99 mg/kg-day CN ,
using an average body weight of 95 g and average food consumption rates [U.S. EPA, 1988]).
Additional groups were administered the low-protein diet with or without KCN and with or
without iodide supplementation. Body weights and food consumption were recorded. Blood
was collected and evaluated for serum TSH levels. Thyroids were removed and weighed. No
difference in body weight was observed among cyanide-treated rats and their respective control
groups. Rats treated with cyanide on the low-protein diet had significantly elevated serum TSH
levels and increased thyroid weights compared to the low-protein control group; supplementation
with iodide in addition to cyanide eliminated these effects. There was no effect on serum TSH or
thyroid weight in the cyanide-treated rats on the normal-protein diet. This study suggests that
severely protein- and iodine-deficient diets are likely to increase the sensitivity of the thyroid
gland to cyanide ingestion.
4.4.2. Acute Inhalation Studies
Relatively few inhalation studies providing quantitative data are available in animals
exposed repeatedly to HCN. Some studies of acute exposure are available in rats, rabbits, and
monkeys (Bhattacharya et al., 1994; Purser et al., 1984; Hugod, 1981). These studies provide
limited information because sample sizes were either very small (Purser et al., 1984) or only a
single organ or endpoint was assessed (Bhattacharya et al., 1994; Hugod, 1981; Valade, 1952).
Purser et al. (1984) exposed cynomolgus monkeys individually to 100, 102, 123, 147, or
156 ppm HCN for up to 30 minutes. These concentrations correspond to 111, 113, 136, 163, or
172 mg/m3 HCN. A single monkey was exposed per concentration, with one monkey exposed to
both 100 ppm and 147 ppm in separate experiments. There was no control group. The time to
incapacitation decreased with increasing exposure levels and ranged from 8 to 19 minutes. The
authors noted that three of the exposures (exposure levels not reported, presumably the three
highest concentrations) were terminated within 30 minutes due to the severity of the symptoms.
The observed symptoms included hyperventilation, decreased and arrhythmic heart rate, loss of
muscle tone and reflexes, and convulsions. Blood cyanide levels reached steady state within
10 minutes. There was no correlation between air concentration and blood cyanide levels.
Bhattacharya et al. (1994) investigated the effects of inhalation of 55 ppm HCN
(61 mg/m3 HCN) for 30 minutes on the pulmonary mechanics of six male Wistar rats.3 In treated
animals, the airflow was increased (20%), accompanied by increased transthoracic pressure
(40%>),and tidal volume (50%>). The respiratory rate, compliance, and minute volume decreased
50, 60, and 25%>, respectively, accompanied by a decrease in pulmonary phospholipids (i.e.,
surfactant) of about 10—30%>. Other effects of cyanide were not evaluated.
3 Data for the investigated parameters were presented graphically and thus the magnitudes of change were estimated
for this review.
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Extensive involvement of the CNS in cyanide toxicity was demonstrated by Valade
(1952), who exposed groups of four dogs to 50 mg/m3 (45 ppm) HCN for a varying number of
30-minute exposure periods conducted at 2-day intervals. Clinical signs included tremors,
stiffness, ataxia, dyspnea, vomiting, and diarrhea. In the longest-term exposure of 36 days
(consisting of 19 exposure periods), two of the four dogs died. Necropsies of the dead and
surviving dogs all showed histopathology in the brain consisting of vasodilation, hemorrhages,
and various cellular lesions.
Myocardial morphology in rabbits was investigated following inhalation of HCN as part
of a study attempting to identify constituents of tobacco smoke responsible for the increased risk
of cardiovascular disease observed in smokers (Hugod, 1981). Male rabbits (22/group) were
exposed to 0 or 0.5 ppm HCN for a period of 4 weeks, after which the animals were sacrificed
and examined for myocardial abnormalities. Following blinded morphologic examination, no
significant effects of cyanide were detected on myocardial ultrastructure.
4.4.3. Neurotoxicity Studies
Crampton et al. (1979) reported on a study in which baboons (7-10/group) were fed a
low cobalamin (vitamin Bi2) diet that was either supplemented with hydroxocobalamin (control)
or unsupplemented. Treated animals received 1 mg/kg KCN subcutaneously 5 days/week. The
body weight of treated animals (with and without hydroxycobalamin supplementation) did not
differ from that of untreated animals. No neurological effects were evident from nerve
conduction measurements or in extensive histopathologic examination of the nervous system,
apparently the only organ system examined.
As described in section 4.1.2, neurological symptoms have been reported in populations
that traditionally consume foods with high concentrations of cyanogenic glycosides, such as
cassava (ATSDR, 2006; Ministry of Health Mozambique, 1984; Osuntokun, 1973; Banea-
Mayambu et al., 1997). Effects include spastic paraparesis, ataxic tropic neuropathy, optic
atrophy, and decreased nerve conduction velocity. Osuntokun (1973) reported that the
neurological effects correlated with blood thiocyanate levels, but other reports found no
correlation between disease severity and thiocyanate level (Ministry of Health, Mozambique,
1984). Several studies (Oluwole et al., 2003; Banea-Mayambu et al., 1997; Kamalu, 1993; Olusi
et al., 1979) indicated that constituents of cassava other than cyanide, such as the parental
cyanogenic glycoside, linamarin, may directly contribute to the characteristic endemic
neurotoxicity observed in these populations. Specifically, an ecological epidemiologic study
conducted in Zaire (Banea-Mayambu et al., 1997) indicated that prevalence of this endemic
neuropathy was more closely correlated with urinary linamarin than urinary thiocyanate.
Fechter et al. (2002) evaluated the effect of HCN exposure on hearing loss and its
interaction with noise-induced hearing loss. Male Long-Evans rats were exposed to HCN for
3.5 hours/day at concentrations of 0, 10, 30, or 50 ppm (equivalent to 0, 11, 33, or 55 mg/m3
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HCN, respectively), to noise alone (i.e., 100 dB volume octave band noise for 2 hours/day
unaccompanied by cyanide exposure), or to noise plus 0, 10, 30, or 50 ppm HCN. Groups of
16 animals were exposed to air alone (0 ppm HCN) or noise alone, and groups of 6-12 animals
were exposed to HCN or HCN plus noise. Hearing loss was assessed 4 weeks after exposure by
evaluating pure tone compound action potential (CAP) thresholds at frequencies ranging from 2
to 64 kHz (i.e., measuring the response at low through high pitches). The CAP threshold is a
measure of change in the electrochemical response of nerve cells in response to auditory
stimulation, a response that is considered to be a measure of cochlear function. This approach
was used in order to evaluate permanent hearing loss rather than the transient loss that occurs
immediately after exposure. Histologic analysis was also conducted on unexposed rats and on
rats exposed to noise alone or in combination with 10 or 30 ppm HCN (three to four rats per
group).
CAP thresholds were not affected by exposure to 10 or 30 ppm HCN. At 50 ppm HCN
(in the absence of noise), CAP thresholds were slightly elevated, but significant differences
among treated groups relative to control were not observed (using ANOVA for repeated
measures). As expected, noise alone did increase the CAP threshold, indicating hearing loss. In
the groups exposed to noise and HCN, there was a concentration-related increase in the CAP
threshold at frequencies of 12-40 kHz, with statistically significant differences at 30 and 50 ppm
as compared with controls. These data indicate that HCN can potentiate noise-induced hearing
loss, but they do not indicate an effect of HCN alone on hearing loss.
In a related study from the same laboratory, i.p. injection of rats with 7 mg/kg KCN
(2.8 mg/kg CN ) caused significant transient hearing loss (Tawackoli et al., 2001). The authors
also found that, in the absence of noise, auditory function recovered as cyanide was eliminated
from the blood. These studies together suggest that hearing loss from cyanide exposure is a
potentially sensitive neurological marker of toxicity. The return of function with the elimination
of cyanide from the blood raises the question of whether a permanent effect would occur under
conditions of high noise exposure and prolonged elevation of blood cyanide levels.
4.4.4. Immune Endpoints
Studies specifically designed to evaluate immune endpoints have not been located in the
HCN database. Additionally, no functional immune measures were identified in the database.
Limited information on immune endpoints exists from human occupational studies and animal
studies. El Ghawabi et al. (1975) found that the percentage of lymphocytes in peripheral blood
was statistically significantly elevated over controls in workers occupationally exposed to HCN
(7-12 mg/m3) for 5-15 years. The percentage of lymphocytes in exposed workers was 42%
(range 32-50) compared to 30% in controls (range 26-40). The total number of leucocytes did
not differ between groups. The biological significance of this magnitude of change in the
relative percentage of lymphocytes is unclear, as is the impact of other chemicals to which the
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workers were concomitantly exposed. Another occupational study (Blanc et al., 1985),
examined workers exposed by inhalation to average concentrations of 17 mg/m3 (15 ppm) for an
average of 11 months. Analyses included a complete blood count and differential with no
significant findings reported for these endpoints.
There are no animal inhalation studies that evaluate the immunotoxicity of HCN, but
inhalation studies on related compounds are available. These studies have evaluated limited
immune-relevant endpoints and are mostly negative. A 3-month inhalation study of ACH (HCN
exposure equivalent of up to 18 mg/m3 or 16 ppm) in rats examined spleen weight and gross and
microscopic histopathology of spleen, lymph nodes, and thymus. In addition, hematology was
examined, including white blood cell (WBC) and differential WBC counts. No changes were
seen in these endpoints (Monsanto Co., 1985a, b). Six-month inhalation studies of (CN)2
inhalation in male rats and monkeys exist (HCN exposure equivalent of 28 mg/m3 or 25 ppm).
Gross necropsy was performed on the spleen and bone marrow. No changes were seen in these
endpoints in either species (Lewis et al., 1984).
Oral studies of cyanide have examined limited immune endpoints. Three-month drinking
water studies in rats and mice (NTP, 1993) with doses up to 12.5 mg/kg-day in rats and
24 mg/kg-day in mice examined immune organs (spleen, thymus, bone marrow, and lymph
nodes) and conducted hematology, including WBC and differential WBC counts. NTP (1993)
did not demonstrate significant changes in any of these endpoints. Additionally, a 2-year oral
study in rats with doses up to 10.8 mg/kg-day examined spleen, thymus, and hematology
endpoints and did not note immunological effects (Howard and Hanzal, 1955).
4.5. MECHANISTIC DATA AND OTHER STUDIES IN SUPPORT OF THE MODE OF
ACTION
4.5.1. Genotoxicity
KCN was not mutagenic in Salmonella typhimurium strains TA82, TA97, TA102, TA98,
TA100, TA1535, TA1537, or TA1538 in the reverse mutation assay with or without metabolic
activation (De Flora et al., 1984; De Flora, 1981). NaCN was not mutagenic in S. typhimurium
strains TA97, TA98, TA100, or TA1535 with or without metabolic activation (NTP, 1993). A
positive response was reported, however, for HCN in S. typhimurium strain TA100 without
metabolic activation; adding metabolic activation reduced the magnitude of the positive response
to 40% of what it had been without metabolic activation (Kushi et al., 1983). Negative results
were obtained in the DNA-repair test in Escherichia coli strains WP67, CM871, and WP2 (De
Flora et al., 1984) and in a test for inhibition of DNA synthesis in HeLa cells (Painter and
Howard, 1982).
Overall, cyanide has tested negative in bacterial mutagenicity studies with and without S9
activation (NTP, 1993; De Flora et al., 1984; De Flora, 1981), although a positive result was
obtained in S. typhimurium strain TA100 with and without S9 activation (Kushi et al., 1983).
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Neither standard chromosome aberration assays nor mammalian gene mutation studies of
cyanide are available. Cyanide was negative in assays for the production of DNA damage and
repair (De Flora et al., 1984; Painter and Howard, 1982).
4.5.2. Acute Neurotoxicity
The mode of action for the acute toxicity of cyanide is well understood (Klaassen, 2001;
Hall and Rumack, 1990). Cyanide is considered a chemical asphyxiant because it impairs
aerobic metabolism without affecting oxygen delivery to the tissues. It has a high affinity for
iron in the ferric state, resulting in binding to and inactivation of tissue cytochrome c oxidase.
Since cytochrome c oxidase normally accepts oxygen from the blood and functions as an
electron acceptor in cellular energy production, this inactivation inhibits cellular respiration. As
anaerobic metabolism proceeds, blood levels of pyruvic acid, lactic acid, and NADPH rise; the
ATP/adenosine diphosphate (ADP) ratio decreases. The earliest effects of acute cyanide toxicity
occur in organs with high aerobic energy demands, particularly the brain and heart. The
inhibition of oxygen use by cells causes oxygen tension to rise in the peripheral tissues, which
results in a decrease in the unloading gradient for oxyhemoglobin. Thus, oxyhemoglobin is
present in the venous blood. In addition to cytochrome c oxidase, cyanide binds to other
metalloproteins and other cellular molecules, including catalase, peroxidase, methemoglobin,
and hydroxycobalamin; this binding also contributes to the symptoms of acute cyanide toxicity.
Cyanide also stimulates the release of secondary neurotransmitters and catecholamines
from the adrenal glands and adrenergic nerves (Kiuchi et al., 1992; Kanthasamy et al., 1991).
Thus, the cardiac effects and the peripheral autonomic responses observed following cyanide
exposure appear to be due to the increase of plasma catecholamine levels. CNS necrosis and
demyelination caused by cyanide may be due to vasoconstriction and low blood flow in the brain,
resulting from low carbon dioxide levels (Brierley et al., 1976). Alternatively, the decreased
ATP/ADP ratio may alter energy-dependent calcium homeostasis in nerve cells (Johnson et al.,
1986). Thus, the acute effects of cyanide result primarily from the interruption of aerobic
metabolism and from the release of secondary neurotransmitters and catecholamines; these
effects include altered respiration, vomiting, nausea, and weakness and ultimately convulsions,
coma, and death.
4.5.3. Thyroid Disruption
The primary cyanide metabolite SCN~ has the same ionic charge and is of similar size as
iodide. SCN competitively inhibits iodide uptake in the thyroid by the sodium-iodide (Na+/F)
symporter (NIS). Iodine is essential for the normal production of the thyroid hormones T3 and
T4. The NIS is a transmembrane protein that actively transports iodide from the bloodstream,
against an electrical and chemical gradient, and concentrates it in the thyroid gland. The human
NIS has been reported to have greater affinity for thiocyanate than for iodide (De Groef et al.,
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2006; Tonacchera et al., 2004; Wolff, 1998). In addition to reducing iodide uptake by the
thyroid, thiocyanate may also cause iodide already accumulated in the thyroid to be discharged
(Wolff, 1998). Additional compounds that have a similar mode of action (i.e.competitive
inhibition of the NIS) include perchlorate, nitrate, chlorate, and fluoroborate; however, each of
these compounds has differing affinities for the NIS and thus different potencies of iodine uptake
inhibition (Tonacchera et al., 2004; Van Sande et al., 2003; Greer et al., 1966). For example,
perchlorate has been estimated to be 15-20 times more potent than thiocyanate in terms of iodine
uptake inhibition (Tonacchera et al., 2004; Greer et al., 1966).
The effect of thiocyanate on the thyroid gland is dose dependent and controlled by
homeostatic processes that tightly regulate and control the synthesis of essential thyroid
hormones in order to ensure a constant systemic supply to meet physiological needs (National
Research Council [NRC], 2005). If thiocyanate interference with iodide uptake is of sufficient
magnitude to decrease the production and secretion rate of thyroid hormones (T4 and T3),
circulating levels of these hormones decrease. Homeostatic mechanisms mediated mainly via
the hypothalamo-pituitary-thyroid feedback axis are rapidly activated to modulate thyroid
hormone synthesis (NRC, 2005; Hill et al., 1989). As the blood levels of these hormones drop,
the hypothalamus, through the release of thyrotropin-releasing hormone, stimulates the pituitary
gland to produce TSH. TSH stimulates the thyroid gland to increase the rate at which it produces
and secretes thyroid hormones. Elevated TSH levels stimulate histologic changes meant to
increase thyroid secretion, such as increased size and number of thyroid cells (Guyton and Hall,
2000). Clinically, this increased size and number of thyroid cells manifests as an enlarged
thyroid gland (goiter). It is only when thiocyanate intake levels are sustained and high enough to
overwhelm homeostatic processes that decreased synthesis and secretion of thyroid hormones
would be expected to occur and thus result in hypothyroidism and effects secondary to
hypothyroidism. This mode of action may be relevant to the thyroid effects observed in both
humans and animals, including thyroid gland enlargement, decreased thyroid hormones, and
increased TSH (Manzano et al., 2007; Baneijee et al., 1997; Kamalu and Agharanya, 1991;
Jackson, 1988; Blanc et al., 1985; Philbrick et al., 1979; El Ghawabi et al., 1975).
4.5.4. Reproductive Effects
The NTP (1993) observed a suite of reproductive effects in rats and mice, including
decreased epididymis weight, testis weight, and testicular spermatid count in rats and mice
treated for 3 months with NaCN in drinking water. The mode of action of these reproductive
effects is not well established. However, some data exist in hypothyroid animals, suggesting that
disruptions in thyroid hormone levels may affect the male reproductive system. Studies in
humans and animals have demonstrated that cyanide exposure can result in decreased thyroid
hormone levels (Manzano et al., 2007; Jackson, 1988; Philbrick et al., 1979; El Ghawabi et al.,
1975). Therefore, it is possible that the observed reproductive effects following exposure to
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cyanide may be mediated through decreases in thyroid hormones mediated through the cyanide
metabolite thiocyanate.
Thyroid hormones are important in the growth and development of a wide range of
tissues, including the male reproductive tract (Kobayashi et al., 2007; Wistuba et al., 2007).
Some reports investigating the developmental effects of hypothyroidism in animal models have
indicated male reproductive effects, including altered maturation of male reproductive organs
and impaired spermatogenesis (Wistuba et al., 2007; Del Rio et al., 2003; Maran and Aruldhas,
2002). Persistent neonatal hypothyroidism in animal models has been shown to result in reduced
reproductive organ weight and decreased sperm count and motility (Sahoo et al., 2008; Hamouli-
Said et al., 2007; Del Rio et al., 1998; Kumar et al., 1994). Conversely, transient neonatal
hypothyroidism has been shown to cause increased testis size and increased sperm production
(Sahoo et al., 2008; Joyce et al., 1993; Cooke, 1991). These experimental observations indicate
that reproductive tissues are sensitive to thyroid hormone levels during development.
In addition to thyroid effects on the growth and development of the reproductive tissues,
some research has suggested that the adult reproductive system is also modulated by thyroid
hormones. In adults, proper thyroid function has also been shown to be important for
maintenance of fertility in adult males and females (Trokoudes et al., 2006; Poppe and
Velkeniers, 2004). Hypothyroidism in adult males has been noted to cause alterations in sex
steroid hormone metabolism, spermatogenesis, and fertility (Krassas and Pontikides, 2004).
Mechanistic studies in adult animals have indicated that reproductive organs, including
the testis and epididymis, may be sensitive to alterations in thyroid hormone levels. A study by
De Paul et al. (2008) demonstrated staining for thyroid receptor protein and mRNA in the adult
rat epididymis, which was shown to be increased in hypothyroid rats showing responsiveness of
adult epididymis tissue in response to decreased thyroid hormone. Additional studies have found
cellular and ultrastructural changes in the adult rat epididymis following induced hypothyroidism
(following thyroidectomy) (Del Rio et al., 2003, 2001, 1979), some of which were reversible
following supplementation with thyroid hormone (Del Rio et al., 1979). Similarly, another study
found reduced epididymis weight in thyroidectomized rats, which was reversible following T4
supplementation (Kala et al., 2002).
In summary, research exists to suggest that reproductive tissues in developing and adult
animals are responsive to alterations in thyroid hormone levels. Additional evidence also exists
to suggest that specific structural changes and decreased epididymis weight can be mediated
through hypothyroidism in the adult animal. Though some information supports this
hypothetical mode of action that reproductive effects observed in the NTP (1993) study may be
due to alterations in thyroid function due to exposure to cyanide, specifically the cyanide
metabolite thiocyanate, uncertainty exists due to the lack of any measurement of indicators of
thyroid function, such as thyroid hormones (TSH, T3, T4) or thyroid weight (NTP, 1993).
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4.6. SYNTHESIS OF MAJOR NONCANCER EFFECTS AND MODE OF ACTION
Tables 4-6 and 4-7 present summaries of noncancer effects from repeat oral and
inhalation exposure to cyanide. Chronic and subchronic cyanide oral exposure studies in
experimental animals indicate that the thyroid, CNS, and male reproductive organs are sensitive
targets of toxicity (Manzano et al., 2007; Soto-Blanco et al., 2002a, b; NTP, 1993; Jackson,
1988). Information from human occupational studies suggests that subchronic and chronic
inhalation exposure to cyanide may be associated with CNS symptoms (including headache,
weakness, and changes in taste and smell) and thyroid alterations (enlargement, altered iodine
uptake, increased TSH, and decreased T3 and T4) (Banerjee et al., 1997; Blanc et al., 1985;
El Ghawabi et al., 1975). Another study also suggests that chronic exposure to HCN in a metal-
tempering plant may reduce pulmonary function in chronically exposed workers
(Chatgtopadhyay et al., 2000).
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Table 4-6. Summary of subchronic and chronic oral toxicity studies for cyanide in animals
Species,
sex
Reference
Dose
(mg/kg-day CN )
Route
Duration
Response
at LOAEL
NOAEL
LOAEL
Comments
Subchronic and chronic
Pig
(6 or
10/group;
sex not
specified)
Manzano et
al. (2007)
0, 1.4,2.8,4.3
Diet;
KCN
10 weeks
Increased thyroid weight
2.8
4.3
All doses showed altered
histology in thyroid, liver,
kidney, and CNS (no
incidences given for
histologic lesions).
Rat, Wistar
(6-7 males/
group)
Soto-Blanco
et al.
(2002a)
0,0.06,0.12, 0.24
Gavage
12 weeks
Histopathologic changes in
CNS
NDa
ND
No incidences given; no
changes in T3, T4.
Rat, F344
(10/sex/
group)
NTP (1993)
Males:
0,0.16,0.48, 1.4,
4.5, 12.5
Females:
0,0.16, 0.53, 1.7,
4.9, 12.5
Drinking
water;
NaCN
13 weeks
Decrease in cauda epididymis
weight and sperm motility
ND
1.4
Thyroid hormones and
thyroid weight not measured.
Mouse,
B6C3F1
(10/sex/
group)
NTP (1993)
Males:
0, 0.26, 0.96, 2.7,
8.6, 24.4
Females:
0,0.32, 1.1,3.3,
10.1,28.8
Drinking
water;
NaCN
13 weeks
Decrease in cauda epididymis
and epididymis weight
8.6
24.3
Dog,
mongrel
(6 males/
group)
Kamalu
(1993);
Kamalu and
Agharanya
(1991)
0, 1.04
Diet;
NaCN
14 weeks
Casts in renal tubules, adrenal
gland hypertrophy and
hyperplasia, and decreased
spermatids in stage VIII of
the spermatogenic cycle;
decreased T3; increased
thyroid weight
None
1.04
Dogs suffered from parasitic
infections.
Goat (6-8
males/
group)
Soto-Blanco
et al.
(2002b)
0,0.12,0.24,
0.58, 1.2
Milk,
drinking
water
5 months
Histopathologic changes in
CNS
ND
ND
Inadequate dose-response
characterization to determine
NOAEL/LOAEL.
Pig
(3/group,
mixed sexes)
Jackson
(1988)
0, 0.4, 0.7, 1.2
Gavage in
water;
KCN
6 months
Decreased T3 and T4;
behavioral changes
0.7
1.2
Single daily bolus dose; no
other endpoints evaluated.
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Table 4-6. Summary of subchronic and chronic oral toxicity studies for cyanide in animals
Species,
sex
Reference
Dose
(mg/kg-day CN )
Route
Duration
Response
at LOAEL
NOAEL
LOAEL
Comments
Rabbit
(6 males/
group)
Okolie and
Osagie
(2000, 1999)
0.2, 20
Diet;
KCN
10 months
Decreased body weight, focal
liver necrosis, tubular and
glomerular necrosis of
kidneys, pulmonary edema
and necrosis
0.2
20
Cyanide in control group
from determination of basal
amount in feed.
Rat, strain
not specified
(10 males/
group)
Philbrick et
al. (1979)
0, 44
Diet
11.5
months
Vacuolation of spinal cord
white matter; decreased T4 at
4 months and increased
thyroid weight at 11.5 months
None
44
Rat (10/sex/
group)
Howard and
Hanzal
(1955)
0, 4.3, 10.8
Diet
2 years
None
10.8
None
Reproductive and developmental
Rat, Wistar
(20 dams,
5 pups/
group)
Imosemi et
al. (2005)
0, 20
Diet
Gestation,
up to PND
50
Decreased body weight,
brain, and cerebellar weight;
altered cerebellar dimensions
None
20
Same study as Malomo et al.
(2004); aggressive and
restless behavior noted in
treated dams.
Rat, Wistar
(20 dams,
5 pups/
group)
Malomo et
al. (2004)
0, 20
Diet
Gestation,
up to PND
50
Reduced thickness of ML and
increased thickness of EGL
of cerebellum
None
20
Indicative of delayed
maturation and migration of
cerebellar cells.
Goat
(5-8/group)
Soto-Blanco
and Gorniak
(2004)
0, 0.4, 0.8, 1.2
Gavage in
water;
KCN
Days 24-
150 (birth)
Increased T3 in dams and
offspring at birth
0.8
1.2
Some of the dams in the
highest dose group
experienced tremors and
ataxia.
Goat
(7/group)
Soto-Blanco
and Gorniak
(2003)
0, 0.4, 0.8, 1.2
Gavage in
water
Lactation
days 0-90
Vacuolation of thyroid,
kidney epithelial cells, and
hepatocytes in offspring and
dams
ND
ND
Histological lesions reported
in treated dams and offspring
but incidence not given.
Rat, strain
not specified
(20 females/
group)
Tewe and
Maner
(1981)
1.2,21.6 (adults)
1.9, 34.3
(weanlings)
Diet;
cassava
and KCN
Through-
out mating,
gestation,
and
lactation
Decreased food consumption,
growth rate, liver weight in
weanlings
1.9
34.3
Control animals fed basal
diet containing low-HCN
cassava; treated animals fed
basal diet supplemented with
KCN.
aND = not determined.
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Table 4-7. Summary of subchronic and chronic inhalation toxicity studies for cyanide in humans
Study population
Reference
Exposure
Duration
Response
NOAEL
mg HCN/m3
LOAEL
mg HCN/m3
Comments
Males
n = 36 exposed workers
n = 20 unexposed controls
Electroplating workers across
3 factories
El Ghawabi et al.
(1975)
7.07, 8.9, 11.5 mg/m3
HCN(6.4, 8.1, 10.4 ppm)
5-15 years
Thyroid
enlargement,
altered iodide
uptake, and CNS
symptoms
None
7.07
Urinary
thiocyanate
correlated with
HCN air
concentration
Males, n = 36 exposed
workers, divided into
low n= 13
medium n = 14 or
high n = 9 exposure
Silver reclaiming facility
Blanc et al.
(1985)
16.6 mg/m3 HCN
(15 ppm)
24-hour TWA taken 1 day
after plant closed
0.5-21
months;
median:
8.5 months
Increased TSH,
increased T3
uptake, CNS
symptoms
None
16.6
No thyroid
enlargement
seen; study
conducted
11 months
postexposure
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The CNS is a target of both acute and chronic cyanide exposure. Symptoms of severe
CNS toxicity following acute cyanide exposure include respiratory depression, convulsions,
coma, and death. Chronic and subchronic inhalation exposure in workers has been reported to
result in symptoms, including headaches, weakness, nausea, and changes in taste and smell at
doses ranging from 7 to 17 mg/m3 (Blanc et al., 1985; El Ghawabi et al., 1975). Behavioral
changes and CNS lesions have been reported in animals exposed orally to cyanide.
Histopathologic effects on various CNS structures also have been observed following subchronic
exposure in rats (Soto-Blanco et al., 2002a; Philbrick et al., 1979) and goats (Soto-Blanco et al.,
2002b). Philbrick et al. (1979) reported that vacuolation was observed in the spinal cord white
matter of rats treated with cyanide or thiocyanate for a year with 44 mg/kg-day. Studies by Soto-
Blanco et al. (2002a, b) in rats and goats at doses from 0.24-1.2 mg/kg-day reported effects,
including neuron loss in the hippocampus, spheroids on white and gray matter of the spinal cord,
damaged Purkinje cells, and loss of white matter in the cerebellum; however, no quantitative data
or statistical analyses were presented for these effects. Behavioral changes in pigs and rats have
also been observed with oral exposure to cyanide. Jackson (1988) observed that pigs
administered 0.4-1.2 mg/kg-day cyanide in the drinking water for 6 months exhibited increased
flight response, a decrease in fighting, and a decrease in exploratory behavior. Additionally,
pregnant rats treated with 20 mg/kg-day cyanide by gavage demonstrated aggressive and restless
behavior (Imosemi et al., 2005). Though the mode of action of acute cyanide toxicity is well
understood (Klaassen, 2001; Hall and Rumack, 1990), the mode of action of CNS changes
observed with chronic cyanide exposure is unclear. It is plausible, due to the mode of action of
cyanide of inhibition of ATP synthesis, that CNS changes upon chronic CN exposure may also
be due to energy deprivation in areas of high metabolic activity in the brain. Conversely, a
chronic study in rats found similarly increased vacuolation of spinal cord white matter,
astrogliosis, and fluid accumulation compared to controls regardless of whether animals were
treated with KCN or KSCN, indicating that these histologic lesions may be due to SCN
(Philbrick et al., 1979).
The thyroid is also an organ sensitive to cyanide, particularly following long-term
exposure. Cyanide's effects on the thyroid are mediated by its metabolite, thiocyanate. Thyroid
enlargement and altered iodine uptake have been seen in workers exposed for 5-15 years to
HCN at concentrations ranging from 7-12 mg/m3 (El Ghawabi et al., 1975). A retrospective
study of a group of 36 male former workers who had been exposed to HCN fumes in a silver-
reclaiming facility for an average duration of 10.5 months found that TSH levels, though still
within normal levels, were elevated compared with those of controls (Blanc et al., 1985).
Thyroid effects, including enlargement, decreased hormone levels, and altered histology have
also been seen in experimental animals orally treated with cyanide (Manzano et al., 2007;
Soto-Blanco and Gorniak, 2003; Jackson, 1988; Philbrick et al., 1979). Rats treated for 4 months
at 44 mg/kg-day had significantly decreased plasma T4 levels (53%) and decreased T4 secretion
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rates (68%) compared with those of controls; however, after 11 months of cyanide treatment, T4
levels no longer differed from controls, though T4 secretion rates were depressed 27%. At the
termination of the study, thyroid weights were significantly increased by 43% in treated animals
(Philbrick et al., 1979). A study in pigs found a dose-related decrease in T3 and T4 (23 and 13%,
respectively) in animals administered 1.2 mg/kg-day CN (Jackson, 1988). Another study in pigs
found increased thyroid weight (24%) in animals administered 4.3 mg/kg-day CIST for 10 weeks,
though significant alterations in thyroid hormones were not observed (Manzano et al., 2007).
Histologic changes in the thyroid, characterized by an increased number of reabsorption vacuoles
on the colloid of the thyroidal follicles, were observed in dams and offspring treated for 90 days
with 0.4-1.2 mg/kg-day CIST (Soto-Blanco and Gorniak, 2003). However, this study found
significantly increased T4 levels in dams treated with 1.2 mg/kg-day, which adds uncertainty to
the interpretation of effects observed in this study. It is well established that the anti-thyroid
effects of cyanide are due to its primary metabolite, SCNThiocyanate competitively inhibits
the active transport of iodide into the thyroid gland, and thus, if homeostatic mechanisms of the
thyroid are overwhelmed, the available concentration of the iodine-based thyroid hormones T3
and T4 can be decreased (Crump and Gibbs, 2005; NRC, 2005; Guyton and Hall, 2000).
Alterations in the male reproductive system have also been observed in some studies.
Male reproductive effects were reported in rats and mice following exposure in drinking water
(NTP, 1993) and in dogs following exposure in feed (Kamalu, 1993). Decreased epididymis,
cauda epididymis, and testis weight, and decreased sperm counts were seen in rats treated with
>12.5 mg/kg-day for 13 weeks (NTP, 1993). Decreased relative epididymis and cauda
epididymis weights were also observed in mice treated with >24.3 mg/kg-day for 13 weeks (NTP,
1993). In Kamalu (1993), dogs treated with 1.04 mg/kg-day for 14 weeks exhibited histologic
changes in the testis, including a significantly decreased percentage of tubules in stage VIII of
the spermatogenic cycle. The NTP (1993) authors suggested that the male reproductive effects
may be related to perturbations in hormonal balance, though thyroid hormones or thyroid weight
were not evaluated as part of the study. However, some data from hypothyroid animals suggest
that thyroid disruption can result in reproductive changes, including sperm decrements and
reproductive organ weight decreases (see section 4.5.4). It is not known whether workers
occupationally exposed to cyanide would be expected to suffer reproductive effects, since it does
not appear that sperm or other reproductive parameters were investigated in the available
occupational studies (Blanc et al., 1985; El Ghawabi et al., 1975).
Cyanide affects the kidneys and liver at doses similar to those that result in male
reproductive toxicity, depending on the species tested. Kidney effects, characterized by casts in
renal tubules, were reported in dogs following subchronic oral dosing at 1.04 mg/kg-day for
14 weeks (Kamalu, 1993). Rabbits dosed at 20 mg/kg-day for 40 weeks had evidence of tubular
and glomerular necrosis (Okolie and Osagie, 1999). Additionally, goat dams and offspring
treated with 0.4-1.2 mg/kg-day throughout lactation were reported to have vacuolation of the
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kidney epithelial cells. Similarly, liver effects with cyanide exposure have been seen in rats,
mice, rabbits, and goats. Goats treated with 0.4-1.2 mg/kg-day for 90 days were reported to
have vacuolation of hepatocytes (Soto-Blanco and Gorniak, 2003). Focal necrosis was seen in
rabbits treated with 20 mg/kg-day in the diet for 10 months (Okolie and Osagie, 1999), and
increased liver weight was observed in female rats and mice administered 12.5 or 23.4 mg/kg-
day, respectively, in drinking water for 13 weeks (NTP, 1993). The mode of action for the
effects of cyanide on the liver and kidneys has not been identified. One hypothesis suggested by
Okolie and Osagie (1999) is that interference with thyroid function by thiocyanate can lead to
decreased tissue protein turnover, resulting in depressed growth, as well as liver and kidney
effects. Additionally, high-dose exposure of cyanide in heavily oxygen-dependant tissues, such
as the liver, may result in ATP deprivation of the tissues.
4.7. EVALUATION OF CARCINOGENICITY
Under the U.S. EPA (2005a) "Guidelines for Carcinogen Risk Assessment," data are
inadequate for an assessment of the human carcinogenic potential of cyanide. No
carcinogenicity studies of cyanide are available in animals or humans. The currently available
data indicate that cyanide is not genotoxic. Bacterial mutagenicity assays, with and without
activation, are predominantly negative (NTP, 1993; De Flora et al., 1984; De Flora, 1981). No
sufficient animal bioassays are available to assess the potential carcinogenicity of cyanide. The
only available chronic study of cyanide that analyzed a wide variety of tissues is an oral rat study
(Howard and Hanzal, 1955); no tumors or lesions were associated with either dose group
following dietary administration of cyanide at doses up to 10.8 mg/kg-day for 2 years. This
study is limited by small sample sizes (n = 10), histopathologic assessment of only a subset of
potential target organs of carcinogenicity, and uncertainty regarding dose due to HCN volatility
issues. Therefore, the available data for cyanide are inadequate to assess the carcinogenic
potential of cyanide.
4.8. SUSCEPTIBLE POPULATIONS
4.8.1. Possible Childhood Susceptibility
Due to the known mode of action of the primary cyanide metabolite, thiocyanate, fetuses
exposed in utero may be susceptible to the effects of cyanide exposure, leading to potentially
reduced thyroid hormone production during critical periods of brain development. The effects of
significantly reduced thyroid hormone levels that result from untreated subclinical or clinical
hypothyroidism are much more severe in the developing young than in adults (NRC, 2005;
Guyton and Hall, 2000). While hypothyroidism in adults typically results in goiter (an enlarged
thyroid), congenital hypothyroidism has been associated with stunted bodily growth and
impaired mental development in fetuses, infants, and young children. Hypothyroidism has been
associated with neurodevelopmental delay and functional and structural neurological deficits in
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young humans and/or rats (NRC, 2005). In human fetuses and neonates, the effects of severe
congenital hypothyroidism associated with severe maternal hypothyroidism during pregnancy
are irreversible and are characterized by long-term impairment of behavior, locomotor ability,
speech, hearing, and cognition (Chan and Kilby, 2000). In moderately hypothyroid neonates
born to mothers who were hypothyroid during pregnancy, prompt supplementation with thyroid
hormone may restore neurodevelopmental function (NRC, 2005). A retrospective study of over
25,000 pregnant women indicated that mild subclinical maternal hypothyroidism can adversely
affect neurological development. Children (aged 7-9 years) whose mothers had subclinical
hypothyroidism during pregnancy were found to have IQ scores 7 points lower than the matched
control children (Haddow et al., 1999). Another study observed that pregnant women with
subclinical hypothyroidism were three times more likely to have a placental abruption (relative
risk [RR] 3.0, 95% confidence interval [CI] 1.1-8.2) and twice as likely to have a preterm birth
(RR 1.8, 95% CI 1.1-2.9) than controls (Casey et al., 2005). These authors speculated that the
reduced IQ of children born to mothers with subclinical hypothyroidism may be related to the
effects of prematurity; however, the mean gestational week at delivery in hypothyroid women in
Haddow et al. (1999) was no different from controls. Additional studies support the indication
of lower neurodevelopmental scores in offspring of women with free T4 levels in the lowest
tenth percentile and normal TSH values during early gestation (Kooistra et al. 2006; Pop et al.,
2003).
In rats, hypothyroidism during development has been associated with anatomical
alterations in the cerebellum, including reduction of growth and branching of Purkinje cells,
delayed proliferation and migration of granule cells, delayed myelination, and changes in
synaptic connections among cerebellar neurons (Koibuchi and Chin, 2000). Although animal
models may provide information on the potential neuroanatomical and neurophysiological
effects of highly reduced maternal thyroid hormone levels during gestation, they are limited in
the ability to assess subtle changes in neurodevelopment, cognition, and behavior that may occur
in humans. Further, the homeostatic system in humans is regarded as more robust and elastic
than that in rodents (NRC, 2005). Thus, animal models provide qualitative, but not quantitative,
information on the effects of low human fetal availability of thyroid hormones during gestation
and early development (Jahnke et al., 2004). The magnitude of decrease in serum T4 levels that
might result in neurodevelopmental effects during gestation and early childhood is not well
characterized and would depend on many factors, including whether T4 levels are reduced during
critical windows of development, the extent of the ability of the fetus to produce its own thyroid
hormones to compensate for decreased maternal thyroid hormone availability, and the nature and
extent of other nutritional deficiencies associated with thyroid hormone production (e.g., iodine,
selenium).
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Although cyanide is a known neurotoxicant, a dose response characterization of
neurodevelopmental toxicity resulting from competitive inhibition of iodide uptake in the thyroid
gland by its thiocyanate metabolite has not been demonstrated.
4.8.2. Possible Gender Differences
Experimental animal studies have indicated that male reproductive toxicity is a target of
chronic cyanide toxicity in rats, mice, and dogs (Kamalu, 1993; NTP, 1993). NTP (1993) found
reduced testicular spermatid count in male rats and decreased reproductive organ weights,
including the epididymis and testes, in both rats and mice. However, effects on female
reproductive organs were not reported in rats or mice at any dose tested (NTP, 1993). Few
studies identifying effects from cyanide exposure studied both sexes; therefore, information
suitable to assess gender differences is lacking.
Population based studies of thyroid disorders indicate gender related trends, with women
being more likely to develop goiter and hypothyroidism (Knudsen et al., 2002; NLM 2008b).
Additionally, analysis of a large data set from the 2001-2002 National Health and Nutrition
Examination Survey (NHANES) indicated statistically significant associations between urinary
levels of perchlorate (which shares a common MOA of competitive iodine inhibition with SCN)
and changes in TSH and T4 levels in women but not in men, with the association strongest in
women with low iodine intake (Steinmaus et al., 2007; Blount et al., 2006). Therefore, women,
especially those with low iodine intake, may be more susceptible to thyroid disruption compared
to men.
4.8.3. Other Susceptible Populations
Due to the ability of thiocyanate to competitively inhibit iodine uptake, people with
preexisting hypothyroidism or low iodine intake, especially pregnant and lactating women, may
represent a susceptible population due to an increased need for iodine during these periods
(WHO 1994). Kreutler et al. (1978) observed that thyroid effects in rats induced by exposure to
KCN could be attenuated by administering iodine concurrently with cyanide. Additionally, an
epidemiologic study by Brauer et al. (2006) demonstrates that populations with low ratios of
urinary iodine to urinary thiocyanate are at increased risk of developing enlarged thyroid.
Populations with low iodine intake exposed to additional chemicals that operate by a
similar mode of action as thiocyanate also represent an additional sensitive population because of
expected additive effects. In addition to thiocyanate, other chemicals, such as perchlorate and
nitrate, can act as competitive inhibitors of NIS, the membrane protein that actively transports
iodine into the thyroid follicular cells (De Groef et al., 2006; Van Sande et al., 2003). A recent
study by Steinmaus et al. (2007) investigated the relationship between urinary levels of
thiocyanate and perchlorate and thyroid hormone levels in women with low iodine intake
(categorized as <100 |ig/L urinary iodine) by using cross-sectional human data gathered as part
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of the National Health and Nutrition Examination Survey. The authors found no association
between urinary thiocyanate levels and T4 or TSH levels in women with low iodine intake.
However, an association was seen between exposure to perchlorate and decreased serum T4
levels. The authors found that this observed association was strengthened when thiocyanate
exposure was considered together with perchlorate.
People with protein deficiency may also be more sensitive to the cyanide-induced thyroid
effects. Kreutler et al. (1978) observed that rats on a low-protein diet (2% casein) demonstrated
increased plasma TSH and thyroid weights following KCN administration, whereas rats on a
normal-protein diet (20% casein) exposed to the same concentrations of cyanide did not develop
these effects. Reduced availability of sulfur-containing amino acids in protein-deficient diets
may impact the concentration of sulfur donors available for the detoxification of cyanide (Frakes
et al., 1986). Studies in human populations ingesting large quantities of cyanide-containing
foods, such as those made from cassava flour, also suggest that increased susceptibility to
thyrotoxic effects are also associated with dietary deficiencies of protein, iodide, vitamin Bi2, or
other vitamins (Makene and Wilson, 1972; Osuntokun et al., 1969).
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5. DOSE RESPONSE ASSESSMENTS
5.1. ORAL REFERENCE DOSE (RfD)
5.1.1. Choice of Principal Study and Critical Effect
The data available on subchronic or chronic oral exposure to cyanide are limited to
experimental studies in animals. Though clinical data from several acute human exposures are
available, no chronic or subchronic studies of oral exposure to cyanide in humans exist. Two
chronic oral exposure studies exist for cyanide, Philbrick et al. (1979) and Howard and Hanzal
(1955), both in rats, though they are limited by the use of only one high dose (Philbrick et al.,
1979) or the failure to detect effects (Howard and Hanzal, 1955). Additionally, there are two
well-designed subchronic studies in rats and mice that tested multiple dose levels and examined
an array of endpoints and tissues (NTP, 1993), and other subchronic studies in a variety of
animal models assessing more limited endpoints and tissues (Manzano et al., 2007; Soto-Blanco
et al., 2002a, b; Okolie and Osagie, 2000, 1999; Kamalu, 1993; Jackson, 1988). Furthermore,
several developmental studies on oral cyanide exposure in rats and goats exist (Imosemi et al.,
2005; Malomo et al., 2004; Soto-Blanco and Gorniak, 2004; Tewe and Maner, 1981).
Manzano et al. (2007) treated 6 or 10 pigs per group (sex not specified and number of
animals unclear due to inconsistencies in reporting) with KCN administered in the diet at 1.4, 2.8,
or 4.3 mg/kg-day for 10 weeks. An increase of 24% in thyroid weight was seen at 4.3 mg/kg-
day. Histologic alterations were reported in the thyroid, liver, kidney, and CNS in all dosed
animals. However, no incidence data or statistical analysis was provided for these histologic
findings, precluding a characterization of the dose response for these effects. This study
identified a LOAEL of 4.3 mg/kg-day and a NOAEL of 2.8 mg/kg-day for a statistically
significant increase in thyroid weight. This study is limited by poor reporting of study design
and observed histologic effects. Due to these limitations and the availability of studies
demonstrating effects at lower levels, this study was not selected as the principal study.
Jackson (1988) evaluated the effects of gavage administration of KCN on behavior and
thyroid function in miniature pigs. Doses of 0, 0.4, 0.7, or 1.2 mg/kg-day KCN were
administered by gavage to three pigs per group (mixed sexes) for 24 weeks. Both T3 and T4
demonstrated a dose-related decrease (23 and 13%, respectively) that was statistically significant
by week 18 of the study. Changes in thyroid hormones were portrayed graphically as means,
without reporting variance or data for individual animals. The author concluded that the overall
pattern of behavioral changes, characterized as an increased ambivalence and slower response to
stimuli, was different in the highest dose group compared to control animals. Based on
behavioral changes and decreased thyroid hormones, the LOAEL and NOAEL are 1.2 and
0.7 mg/kg-day CN , respectively. The biological significance of the behavioral changes
observed in this study is unclear. In addition, the utility of this study is limited by the use of
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bolus dosing. In comparison to relatively steady intake throughout the day via dietary
administration, bolus dosing produces higher peak blood levels as the entire daily dose is rapidly
absorbed. This difference is especially important considering that the toxicity of cyanide is
highly rate dependent. Thus, the findings from bolus exposure to cyanide are considered less
relevant to subchronic or chronic exposure conditions in humans. Due to the use of a bolus
regimen, this study was not considered appropriate for selection as the principal study.
Soto-Blanco et al. (2002a) treated Wistar rats (six to seven per group) with 0.06, 0.12, or
0.24 mg/kg-day by gavage for 12 weeks and reported histopathologic changes in the CNS. The
same authors also conducted a 5-month drinking water study in female goats (six to eight per
group) with concentrations ranging from 0.12 to 1.2 mg/kg-day (Soto-Blanco et al., 2002b). In
these studies, the authors reported a variety of histopathologic changes in the CNS described as
spheroids on the spinal cord, neuronal loss in the hippocampus, damaged Purkinje cells, gliosis,
and loss of cerebellar white matter. However, the authors provided no quantitative data,
precluding a dose-response characterization of the reported effects. The lack of quantitation of
the observed histologic effects and the use of bolus dosing precluded further consideration as the
principal study.
Kamalu (1993) evaluated the toxicity of NaCN administered to 22-week-old mongrel
dogs (six males per group) for 14 weeks. The diet was supplemented with NaCN corresponding
to 1.04 mg/kg-day CN , Lesions in the kidneys and adrenal gland were reported at the only dose
tested; however, no quantitation of these lesions was provided. In the testes, a specialized
morphologic analysis indicated that the treated dogs had a significantly decreased percentage of
tubules in stage VIII of the spermatogenic cycle, as compared with controls (p < 0.01). An
evaluation of the thyroid of animals in this study was published by Kamalu and Agharanya
(1991). Serum T3 was significantly decreased (55%) and thyroid weight was significantly
increased (23%) in the cyanide-exposed group. Based on thyroid enlargement and
histopathologic changes in the kidneys, testes, and adrenal glands, the only dose tested,
1.04 mg/kg-day, was considered to be the LOAEL. The authors reported that the animals
suffered from recurring parasitic infestations that required treatment with pharmaceuticals
throughout the study. Therefore, the use of the data from the Kamalu (1993) and Kamalu and
Agharanya (1991) studies are limited by the use of dogs of compromised health status and were
not selected as the principal study for the derivation of the RfD.
Studies observing low-level developmental effects were also considered in the selection
of the principal study and critical effect (Soto-Blanco and Gorniak, 2004, 2003). Soto-Blanco
and Gorniak (2004) administered gavage doses of CN~ equivalent to 0, 0.4, 0.8, or 1.2 mg/kg-
day throughout gestation (GD 24 to birth) to pregnant goats (six per group) and observed
elevated T3 (but not T4) levels in dams and offspring tested at birth in the highest dose group.
Another publication by the same authors (Soto-Blanco and Gorniak, 2003) treated goats with 0,
0.4, 0.8, or 1.2 mg/kg-day during lactation (PNDs 0-90) and identified vacuolation of kidney
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epithelial cells and hepatocytes in offspring and dams at unspecified doses. Incidence or severity
of the reported histologic lesions was not provided, precluding any characterization of dose
response. These studies are limited by the use of bolus doses and the lack of dose-response
characterization, which precludes their selection as principal studies.
A study by NTP (1993) examined the toxicity of cyanide over a wide range of doses.
NTP administered NaCN in drinking water to rats and mice (10/sex/group) at concentrations of
0, 0.16, 0.48, 1.4, 4.5, and 12.5 mg/kg-day CIST in male rats; 0, 0.16, 0.53, 1.7, 4.9, and
12.5 mg/kg-day in female rats; 0, 0.26, 0.96, 2.7, 8.6, and 24.4 mg/kg-day CIST in male mice; 0,
0.32, 1.1, 3.3, 10.1, and 28.8 mg/kg-day in female mice. Reproductive effects were observed in
male animals of both species, though rats appeared to be the more sensitive species. In rats, a
statistically significant decrease in cauda epididymis weight (7%) was seen at doses >1.4 mg/kg-
day. A 7% decrease in whole epididymis weight (as compared to cauda epididymis weight) was
seen at 12.5 mg/kg-day. At the highest dose tested, 12.5 mg/kg-day, epididymis and cauda
epididymis weights were decreased by 7 and 13%, respectively. Dose-related decreases in testis
weight (8%), number of spermatid heads (14%), and spermatid concentration (14%) were also
found to be significant at doses >12.5 mg/kg-day. No change in epididymal sperm count was
observed at any dose, however, a statistically significant decrease in epididymal sperm motility
was observed at doses >1.4 mg/kg-day CN , though it did not appear to increase in severity with
dose.
In consideration of the available studies reporting low-dose effects of chronic and
subchronic oral exposure to cyanide in animals, the NTP (1993) study was chosen as the
principal study. This study was well designed, with five dose groups of 10 animals per group per
sex and species. Numerous tissues and endpoints were assessed, and methods and observed
effects were thoroughly reported. This study identified statistically significant male reproductive
effects in rats and mice that increased in severity in a dose-dependent manner. The observed
effects included decreased cauda and whole epididymis weights, decreased testes weight, and
altered sperm parameters.
The reproductive effects observed by NTP (1993) are consistent with an effect on male
reproductive endpoints, including organ weights and sperm parameters, although the magnitude
of the effects alone may be insufficient to decrease fertility in rats. However, human males have
markedly lower rates of sperm production and sperm counts compared with rats, thus the
potential impact of decrements in sperm quality in humans is considered to be greater than that
of rats (U.S. EPA, 1996; Working, 1988). Furthermore, the cyanide database contains limited
additional support for the specific endpoint of reproductive toxicity (Kamalu, 1993). Therefore,
for the above reasons, NTP (1993) was chosen as the principal study, and all statistically
significantly altered reproductive endpoints in rats and mice were benchmark dose (BMD)
modeled and are presented in section 5.1.2 and Appendix B.
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EPA has selected decreased cauda epididymis weight as the critical effect because it was
determined that this effect represents the most sensitive endpoint indicative of male reproductive
toxicity. The cauda epididymis is one of the three primary subsections of the epididymis (along
with the caput and corpus) and functions as the site of sperm storage and maturation. Because
the cauda is part of the epididymis, these weights are not independent endpoints. BMD analysis
of the observed reproductive data from rats and mice indicated decreased (absolute) whole
epididymis and cauda epididymis weights to be the most sensitive reproductive effects observed.
Points of departure (PODs) for these endpoints identified through BMD modeling were virtually
identical. However, examination of the organ weight data from the principal study (NTP, 1993)
indicated that data for decreased cauda epididymis weight was statistically significantly
decreased (7%) at the lowest dose tested, whereas the decrease in whole epididymis weight (2%)
was not. It is possible that use of whole epididymis weight may mask the effect first observed in
the cauda region of the epididymis. Thus, decreased cauda epididymis weight was considered a
more sensitive effect.
Altered sperm parameters support the observed decreases in reproductive organ weight
seen in the NTP (1993) study. At the lowest dose examined, 1.4 mg/kg-day CN, a modest, but
statistically significant decrease in epididymal sperm motility (4%) was observed, though its
severity did not increase with dose. Additionally, at the highest dose tested, testicular spermatid
count was statistically significantly decreased (14%). Epididymal sperm count was not affected
at any dose tested.
Human male fertility is established to be lower than that of rodent test species, thus
human fertility may be more susceptible to damage from toxic agents (Working, 1998; US EPA,
1996). Therefore, according to the US EPA Guidelines for Reproductive Toxicity Risk
Assessment (US EPA, 1996), statistically significant changes to measures in sperm parameters
including sperm count, morphology, or motility are considered adverse. However, no decrease
in epididymal sperm count and only a modest decrease in sperm motility were observed at doses
which the cauda epididymis weight was statistically significantly decreased. The data from NTP
(1993) suggests that cauda epididymis weight is an effect which precedes more severe
decrements in sperm parameters, such as decreased testicular spermatid count, seen at the highest
dose. Therefore, decreased cauda epididymis weight, the most sensitive effect observed in this
study, was chosen as the critical effect.
Several studies in a variety of experimental models described above have reported
LOAELs in the same range or lower than the reproductive effects identified by NTP (1993).
However, interpretations of these studies are complicated by various issues, including limited
reporting of methods, incidences, severity, and statistical significance of observed effects in
addition to the use of bolus dosing regimens and animals of compromised health status
(Manzano et al., 2007; Soto-Blanco and Gorniak, 2004, 2003; Soto-Blanco et al., 2002a, b;
Kamalu, 1993; Jackson, 1988). Nevertheless, possible reference values (RfVs) based on the
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observed effects from Jackson (1988), Manzano et al. (2007), and Kamalu (1993) are presented
for comparison in section 5.1.4.
5.1.2. Method of Analysis
Statistically significantly altered reproductive endpoints in rats and mice observed in the
NTP (1993) study were BMD modeled, including decreased cauda and whole epididymis
weights, decreased testes weight, and altered sperm parameters (Table 5-1). Epididymal sperm
motility, though statistically significantly decreased in all treated groups, did not exibit a dose-
response relationship, and thus was not amenable to BMD modeling. For reproductive organ
weight changes, absolute reproductive organ weights, as opposed to relative organ weights, were
modeled. The absolute reproductive organ weight data presented by NTP (1993) showed dose-
related decreases in rats and mice. Relative organ weights did not show a stonger dose-response
than absolute organ weights. The study found body weight decreases in the highest dose group of
male rats (6%, p < 0.05) and in the highest dose of male mice (4%, not statistically significant).
Given the lack of substantive body weight changes in rats and mice, especially at lower doses
which showed organ weight changes, relative organ weights were not analyzed further.
Table 5-1. Reproductive endpoints in male rats and mice observed following
administration of NaCN in drinking water for 13 weeks
Concentration (ppm)
0
30
100
300
Rats
Dose (mg/kg-day)
0
1.4
4.5
12.5
Weights (g)a
Cauda epididymis, absolute
0.162 ±0.009
0.150 ±0.013b
0.148 ±0.013b
0.141 ±0.009c
Epididymis, absolute
0.448 ±0.019
0.437 ±0.016
0.425 ± 0.022b
0.417 ±0.016c
Testis, absolute
1.58 ±0.094
1.56 ±0.063
1.52 ±0.063
1.46 ± 0.063°
Spermatid measurements3
Spermatid count (/10 4 mL)
89.28 ± 9.64
84.68 ± 12.74
82.90 ± 9.99
77.10 ± 6.96b
Mice
Dose (mg/kg-day)
0
2.7
8.6
24.3
Weights (g)a
Epididymis, absolute
0.049 ± 0.003
0.047 ± 0.006
0.047 ± 0.003
0.044 ± 0.003b
Cauda epididymis, absolute
0.017 ±0.003
0.016 ±0.000
0.015 ±0.003b
0.014 ±0.003b
aValues are mean ± SD.
bSignificantly different from control atp< 0.05 using Shirley's test.
Significantly different from control atp< 0.01 using Shirley's test.
Source: NTP (1993).
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Continuous models (i.e., linear, polynomial, and power) with constant variance were fit
to the data by using U.S. EPA BMD software (BMDS) (version 1.4.1). The other continuous
model available in BMDS, the Hill model, was not fit to these data because fitting of the Hill
model requires the estimation of four parameters (i.e., intercept, v, n, and k), which necessitates
having a minimum of five dose groups in order to have adequate degrees of freedom for testing
model fit. The NTP (1993) study employed only four dose groups, and thus the Hill model could
not be fit to these data. A benchmark response (BMR) level was selected corresponding to a
change in the mean response equal to one SD from the control mean for cauda epididymis
weight. In this case, a one SD change in cauda epididymis weight was selected under an
assumption that it represents a minimal biologically significant change. By using the best fitting
model for this data set, a one SD change was equivalent to a 7% decrease in cauda epididymis
weight. The BMD modeling reports generated from modeling the reproductive endpoints from
the NTP (1993) study are summarized below in Table 5-2.
Table 5-2. BMD modeling results for observed reproductive endpoints
Endpoint
Fitted model
Goodness-of-fit
p value
AICa
BMD
BMDLa
Rats
Cauda epididymis weight
(absolute)
Linear
0.08
-312.75
8.4
5.6
Polynomial
0.11
-313.10
3.5
1.9
Power
0.08
-312.75
8.4
5.6
Epididymis weight (absolute)
Linear
0.22
-274.73
8.2
5.6
Polynomial
0.73
-275.64
3.2
1.8
Power
0.22
-274.73
8.2
5.6
Testis weight (absolute)
Linear
0.82
-167.94
7.4
5.1
Polynomial
0.98
-166.32
5.3
2.4
Power
0.82
-167.94
7.4
5.1
Spermatid concentration
Linear
0.70
227.04
11.2
6.9
Polynomial
0.53
228.73
8.5
2.9
Power
0.70
227.04
11.2
6.9
Mice
Epididymis weight (absolute)
Linear
0.67
-378.71
21.5
13.0
Polynomial
0.38
-376.73
20.5
6.6
Power
0.67
-378.71
21.5
13.0
Cauda epididymis weight
(absolute)
Linear
0.87
-402.99
25.9
14.6
Polynomial
0.82
-399.67
16.3
5.2
Power
0.69
-400.99
16.3
14.6
aAIC = Akaike's Information Criterion. BMDL = 95% lower confidence limit on the BMD.
Data source: NTP (1993).
All three models provided an adequate fit to this data set based on the goodness-of-fit
statistic (p value > 0.1). Of these three models, the polynomial model provided the best fit to the
data based on this model's exhibiting the lowest Akaike's Information Criterion (AIC) and visual
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inspection of the plot of observed versus expected values across the three models. The detailed
BMD modeling output for the selected polynomial model is presented in Appendix B. The BMD
associated with a one SD decrease in cauda epididymis weight in rats is 3.5 mg/kg-day, and its
95% lower confidence limit (BMDL) is 1.9 mg/kg-day.
5.1.3. RfD Derivation-Including Application of Uncertainty Factors (UFs)
The BMDL of 1.9 mg/kg-day based on decreased cauda epididymis weight in rats was
used as the POD for the derivation of the RfD. A total uncertainty factor (UF) of 3000 was
applied to the POD of 1.9 mg/kg-day: 10 for interspecies extrapolation from animals to humans
(UFa), 10 for human intraspecies variability (UFH), 10 to account for the use of a subchronic
study (UFS), and 3 to account for database deficiencies (UFD).
A 10-fold UF was used to account for uncertainties in extrapolating from laboratory
animals to humans. Humans and laboratory animals have qualitatively similar absorption,
distribution, metabolism, and excretion of cyanide. However, quantitative comparisons of
toxicokinetic parameters are lacking. Additionally, a wide range of sensitivity to effects of
cyanide has been observed between different species of experimental animals. The available
data do not provide quantitative information on the difference in susceptibility to cyanide
between rats and humans.
A default 10-fold UF was used to account for variation in susceptibility among members
of the human population (i.e., interindividual variability). Insufficient information is available to
quantitatively estimate variability in human susceptibility to cyanide.
A 10-fold UF was applied for the extrapolation of subchronic-to-chronic exposure
duration since a 13-week subchronic study (NTP, 1993) was used as the basis for the RfD. It is
not known whether progression of the reproductive effects observed in the principal study would
be observed with increased duration of exposure or whether reproductive or other effects would
be observed following chronic exposure at a lower dose.
An UF of 3 was applied to account for deficiencies in the cyanide toxicity database,
including the lack of a multigenerational reproductive study and a sensitive developmental
neurotoxicity study. The database includes limited human data from epidemiological studies of
workers exposed by inhalation to HCN (Chatgtopadhyay et al., 2000; Banerjee et al., 1997;
Blanc et al., 1985; El Ghawabi et al., 1975). The database also includes studies in laboratory
animals, including chronic and subchronic dietary exposure studies and developmental studies.
The database includes oral toxicity studies in various animal species, including rats, mice, rabbits,
dogs, pigs, and goats. A developmental study with skeletal and visceral examination has not
been conducted for cyanide; however, developmental studies in rats and goats evaluating the
thyroid, kidney, liver, pancreas, brain, and CNS system of gestationally and/or lactationally
exposed offspring exist (Imosemi et al., 2005; Malomo et al., 2004; Soto-Blanco and Gorniak,
2004, 2003; Tewe and Maner, 1981). External or overt developmental effects with cyanide
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exposure have not been noted at doses up to 1.2 mg/kg-day in goats (Soto-Blanco and Gorniak,
2004. 2003) and 21.6 mg/kg-day in rats (Imosemi et al., 2005; Malomo et al., 2004; Tewe and
Maner, 1981). However, due to the mode of action of thiocyanate involving competitive iodine
uptake inhibition and implications for neurotoxicity in the developing animal, the lack of
developmental studies that assess learning and development is an additional weakness in this
database. The cyanide database does not have an appropriately designed multigenerational
reproductive study, but assessment of reproductive organs was included as a component of the
13-week NTP (1993) studies in rats and mice, and testicular histology was also assessed in dogs
(Kamalu, 1993). No available studies for cyanide have assessed reproductive success in
cyanide-treated animals.
An UF for LOAEL-to-NOAEL extrapolation was not used because the current approach
is to address this factor as one of the considerations in selecting a BMR for BMD modeling. In
this case, a BMR of a one SD change from the control mean in epididymis weight was selected
under an assumption that it represents a minimal biologically significant change.
The oral RfD for CN was calculated as follows:
RfD = BMDLio - UF
= 1.9 mg/kg-day ^ 3000
= 6.3 x 10 4 mg/kg-day (rounded to 6 x 10 4 mg/kg-day)
The RfDs for simple cyanide salts like NaCN and KCN, which freely dissociate into
cyanide, are calculated from the RfD for CIST by adjusting for molecular weight (i.e., the RfD is
multiplied by the ratio of the total molecular weight of the compound to the molecular weight of
the CN~):
RfD for aqueous HCN [HCN(aq)] = 6.3 x 10 4 x 27/26 = 7 x 10 4 mg/kg-day
RfD for NaCN = 6.3 xlO^x 49/26 =lx] o 3 mg/kg-day
RfD for KCN = 6.3 x 10~4 x 65/26 = 2 x 10~3 mg/kg-day
RfD for calcium cyanide4 [Ca(CN)2] = 6.3 x 10 4 x 92/(2 x 26) = 1 x 10 3 mg/kg-day
RfD for potassium silver cyanide5 [KAg(CN)2] = 6.3 x 10 4 x 199/26 = 5 x 10 3 mg/kg-day
RfD for cyanogen5 (CN)2 = 6.3 x 10 4 x 52/26 = 1 x 10 3 mg/kg-day
Use of the RfD for free cyanide to calculate RfDs of other cyanide compounds may be
merited, but the ability of the individual cyanogenic species to dissociate and release free
cyanide in aqueous solution (and at physiological pHs) should be taken into consideration. If
4 Two molar equivalents of free CN released in water.
5 One molar equivalent of free CIST released in water.
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dissociation of the compound is expected, liberated cations should be considered for potential
toxicity independent of CN~. Also, some metallocyanides, such as copper cyanide, have
chemical-specific data and are not included in this analysis.
5.1.4. RfD Comparison Information
Reduced reproductive organ weight, altered sperm parameters, increased thyroid weight,
altered thyroid hormones, and altered testicular, kidney, and adrenal histopathology are observed
low-level effects following subchronic oral exposure to cyanide (Manzano et al, 2007; Kamalu,
1993; NTP, 1993; Jackson, 1988). Table 5-3 provides a summary of alternate PODs and
resulting potential reference values (RfVs) derived from these endpoints. Additionally, Figure 5-
1 provides a graphical representation of this information. This figure should be interpreted with
caution since the PODs across studies are not necessarily comparable, nor is the confidence the
same in the data sets from which the PODs were derived. The PODs presented in this figure are
based on either a BMDLisd, NOAEL, or LOAEL (if no NOAEL was available).
A composite UF of 30,000 was applied to the LOAEL for the endpoints identified in the
Kamalu (1993) study, which is generally considered too large to support derivation of a
reference value. In the report, A Review of the Reference Dose and Reference Concentration
Processes (U.S. EPA, 2002), the RfD/RfC Technical Panel concluded that, in cases where the
total UF is more than 3,000, it is unlikely that the study or database is sufficient to derive a
reference value. Thus, the magnitude of the uncertainty associated with this specific study
indicates it is insufficient to support derivation of a reference value.
Some indication of the confidence associated with the resulting potential RfVs is
reflected in the magnitude of the total UF applied to the POD (i.e., the size of the bar); however,
the text of sections 5.1.1 and 5.1.2 should be consulted for a more complete understanding of the
issues associated with each data set, the rationale for the selection of the principal study, and the
critical effect used to derive the RfD. As discussed in section 5.1.1., among the studies
considered, the subchronic study by NTP (1993) provided the data set for the derivation of the
RfD.
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Table 5-3. Alternate PODs with applied UFs and resulting potential RfVs
Effect
POD
Species
UFe
Potential RfV
Total
A
H
L
s
D
Decreased testis weight
5.r
Rat
3,000
10
10
NA
10
3
2 x 10~3
Decreased epididymis
weight
1.8a
Rat
3,000
10
10
NA
10
3
6 x 10~4
Decreased cauda
epididymis weight
1.9a
Rat
3,000
10
10
NA
10
3
6 x 10~4
Decreased testicular
spermatid concentration
6.9a
Rat
3,000
10
10
NA
10
3
2 x 10~3
Decreased epididymis
weight
13a
Mouse
3,000
10
10
NA
10
3
4 x 10~3
Decreased cauda
epididymis weight
14.6a
Mouse
3,000
10
10
NA
10
3
5 x 10~3
Altered thyroid hormones,
behavioral changes
0.7b
Pig
3,000
10
10
1
10
3
2 x 10~4
Increased thyroid weight
2.8°
Pig
3,000
10
10
1
10
3
9 x 10~4
Kidney, testes, and
adrenal gland effects
1.04d
Dog
30,000
10
10
10
10
3
4 x 10~5
aBMDL based on BMD modeling of a one SD change. Source: NTP (1993); NA, not applicable.
bPOD based on NOAEL. Sources: Jackson (1988);
°POD based on NOAEL. Source: Manzano et al. (2007).
dPOD based on LOAEL. Source: Kamalu (1993).
eUFs: A = animal to human (interspecies); H = interindividual (intraspecies); L = LOAEL-to-NOAEL; S =
subchronic to chronic duration; D = database deficiency.
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100
10
0.1
0.01
0.001
0.0001
0.00001
Decreased
cauda
epididymis
weight, mice
(NTP, 1993)
Decreased
epididymis
weight
mice
(NTP, 1993)
Increased
thyroid
weight, pigs
(Manzano
2007)
Decreased
testis
weight, rats
(NTP, 1993)
Decreased
epididymis
weight, rats
(NTP, 1993)
*Decreased
cauda
epididymis
weight, rats
(NTP, 1993)
Decreased
testicular
spermatid
concentration.
rats
(NTP, 1993)
_ hyroid
hormone and
behavioral
changes, pigs
(Jackson, 1988;
Changes in
kidney, testes
and adrenal
gland, dogs
(Kamalu, 1993)
Figure 5-1. Potential RfV comparison array for alternate PODs.
Point of departure
• Potential RfV
* Critical effect and RfD
UF, animal to human
UF, human variability
UF, subchronic to chronic
UF, LOAEL to NOAEL
UF, database
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5.1.5. Previous RfD Assessment
An RfD for cyanide was posted on the IRIS database in 1985 and was based on
coprincipal studies, previously described in section 4.2.1. Howard and Hanzal (1955) fed food
fumigated with HCN to rats for 2 years and identified the high dose of 10.8 mg/kg-day as the
NOAEL. Philbrick et al. (1979) evaluated the thyroid and nervous system in rats fed KCN for
11.5 months. The single dose tested, 44 mg/kg-day, was identified as a LOAEL, based on
myelin degeneration in the CNS and increased thyroid gland weight. These two studies were
considered together to identify the critical effect and the POD. The previous RfD was based on a
NOAEL of 10.8 mg/kg-day. The NOAEL was divided by a UF of 500, including a factor of 10
each for extrapolation from animals to humans and intraspecies variability. A modifying factor
of 5 was used to account for the apparent tolerance to cyanide when it is ingested with food
compared with administration by gavage or by drinking water (U.S. EPA, 1992). However, as
discussed in section 4.4.1 (Palmer and Olson, 1979), the apparent difference may have been due
to instability of the cyanide concentrations in the feed rather than differences in bioavailability.
In addition, newer methodology has been developed, including the application of uncertainty
factors for database deficiencies such as a lack of developmental and reproductive studies. The
resulting RfD was 10.8/500 = 0.02 mg/kg-day.
5.2. INHALATION REFERENCE CONCENTRATION (RfC)
5.2.1. Choice of Principal Study and Critical Effect
Limited data are available on the effects of long-term inhalation exposure to cyanide.
Several occupational studies investigated the effects of inhalation exposure to HCN, and three of
these studies provide evidence of effects on the thyroid. One of these studies included
environmental exposure data based on breathing zone samples for the individual study
participants. There are no subchronic or chronic inhalation exposure studies of HCN in animals.
El Ghawabi et al. (1975) reported statistically significantly altered rates of iodide uptake
by the thyroid, thyroid enlargement, and CNS symptoms (e.g., self-reported increased incidence
of headache, weakness, and sensory changes for taste and smell) in workers (n = 36) exposed to
HCN for 5-15 years in three electroplating factories. Individual breathing zone measurements of
HCN were collected from each worker. The mean concentrations across factories ranged from
7.1 to 11.5 mg/m3 HCNand the values for individual workers ranged from 4.6 to 13.7 mg/m3
HCN. Urine SCN levels, an indicator of internal dose, collected from workers were highly
correlated with individual HCN exposure concentrations (see Figure 4-1). Twenty of the
exposed workers (56%) were identified with mild to moderate thyroid enlargement. Radioactive
iodine uptake measured following a 2-day break in HCN exposure indicated statistically
significantly elevated iodide uptake after 4 hours (38.7% compared to 22.4%) and 24 hours
(49.3%) compared to 39.9%>) as compared to controls.
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Increased 24 hour uptake of radioactive iodide by the thyroid has been reported to occur
in hyperthyroidism, iodine deficiency and goiter (Ravel 2005; NLM 2008a). The study authors
concluded that the increased iodine uptake observed in the workers following the 2-day cessation
in exposure was a postexposure response to depletion of iodine in the thyroid. A similar increase
in iodide uptake has been seen with perchlorate (CIO4 ), another competitive inhibitor of iodide
uptake, following cessation of exposure. Lawrence et al. (2000) measured iodine uptake in
volunteers administered doses of C104 at baseline, at 2 weeks of dosing, and then 2 weeks
postexposure cessation. The authors reported that iodide uptake decreased 10-38% in the low-
and high-dose groups (compared to baseline) at 2 weeks of dosing. Two weeks after exposure
was discontinued, iodide uptake was statistically significantly increased 22 and 25%, indicating a
rebound effect in iodide accumulation postexposure.
The lowest mean concentration of HCN recorded in the three factories, 7.1 mg/m3, is
designated as a LOAEL for thyroid enlargement and altered iodide uptake. The study authors
also indicated some coexposure of the workers to gasoline, alkali, and acid during the
electroplating process, although the magnitudes of these exposures were not quantified and it is
unclear if these exposures would impact the observed thyroid effects. Blanc et al. (1985)
conducted a study of silver-reclamation workers (n = 36) examined an average of 11 months
following exposure. The median length of employment was 8.5 months. Workers were
categorized into low-, moderate-, or high-exposure groups based on their primary job activities.
Information on exposure was limited, as the plant had been shut down following the death of one
worker from cyanide overexposure. Environmental monitoring conducted the day after the plant
was shut down found that the 24-hour TWA HCN exposure was 16.6 mg/m3. Serum TSH levels
in workers were significantly elevated relative to laboratory controls. The authors noted a
significant positive trend with increasing exposure level for self-reported weight loss and several
symptoms, including dizziness, syncope, and nausea and vomiting. Serum TSH levels in
workers were reported as being significantly elevated in workers relative to laboratory controls.
T3 uptake in the highest exposed workers (n = 9) was statistically significantly elevated
compared to laboratory controls. The authors reported that this elevation may reflect a
post-inhibitory response. Because there were multiple possible routes of cyanide exposure,
including dermal exposure and contamination of food, and because earlier air levels were likely
higher than the measured TWA concentration, the environmental monitoring data do not allow
for the selection of a reliable LOAEL. Additionally, this study examined workers an average of
11 months post occupational HCN exposure and so may have missed effects that have the
potential to regress following cessation of exposure. The observation of significant effects on
the thyroid almost 1 year after cessation of exposure indicates that these observed thyroid effects
are not transient. Due to limitations of this study based on its retrospective design and because
of the observation of significant effects at lower levels in the El Ghawabi et al. (1975) study, this
study was not selected for the derivation of the RfC.
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In another occupational study of electroplating workers exposed to HCN, workers
(n = 35) exposed for 5 years had significantly decreased T3 (48%) and T4 (37%) and significantly
increased TSH (142%) as compared to controls (Baneijee et al., 1997). Serum SCN was
elevated in workers compared to controls. A significant negative correlation between serum T4
and SCIST concentrations and a significant positive correlation between TSH and SCIST
concentrations were observed. However, no information was provided on exposure levels,
therefore no NOAEL or LOAEL could be determined from this study.
Chandra et al. (1980) reported on a group of 23 electroplating workers chronically
exposed to average breathing zone concentrations of 0.15 mg/m3 HCN. The authors noted that
the workers complained of symptoms typical of cyanide poisoning but provided no additional
information on specific symptoms or further analysis. In the absence of measured adverse
endpoints, no NOAEL or LOAEL could be determined from this study, precluding its use in a
quantitative risk assessment.
Chatgtopadhyay et al. (2000) found some indication of decreased pulmonary function in
workers at a metal-tempering plant. Specifically, the authors observed decreased pulmonary
function in 24 workers exposed for a mean duration of 24 years. This study provided no
information regarding the environmental exposure levels of the workers, and thus no NOAEL or
LOAEL could be determined, limiting this study's utility for risk assessment.
Considering the availability of studies in the HCN database, El Ghawabi et al. (1975) was
chosen as the principal study. The results of this study indicate that chronic, low-level exposure
to cyanide was associated with thyroid enlargement and altered iodine uptake in humans. This
study examined workers exposed to HCN for extended durations (5-15 years). Although this
study is limited by small sample size, it used matched controls and is not confounded by
smoking since all workers and controls were nonsmokers. The authors collected individual
breathing zone measurements of HCN exposure, which were strongly correlated with urinary
SCN , a measure of internal exposure. Mean individual HCN concentrations reported from all
three plants were close (6.4-10.4 ppm) and the range among the 36 individuals was also
relatively small (4.2-12.4 ppm), indicating similar magnitude of exposure for exposed workers.
Thyroid enlargement was strongly associated with HCN exposure with 56% of the exposed
workers diagnosed with mild to moderate thyroid enlargement. This observation is supported by
an increased radioactive iodide uptake in workers (p < 0.001). Increased uptake of radioactive
iodide has been reported to occur in hyperthyroidism, iodine deficiency, recovery from thyroid
suppression and goiter (Ravel 2005; NLM 2008a; Spencer 2008). The increase in iodide uptake
may have resulted from temporary weekend cessation of exposure. A similar phenomenon of
post-inhibitory response was also seen in the occupational study by Blanc et al. (1985), which
noted significantly increased T3 uptake observed in workers several months following HCN
exposure. The thyroid alterations reported in El Ghawabi et al. (1975) are believed to be
biologically significant effects. These effects, particularly thyroid enlargement, are consistent
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with those observed in oral exposure animal studies (Manzano et al., 2007; Jackson, 1988;
Philbrick et al., 1979). Additionally, other human inhalation studies have indicated thyroid
effects in exposed workers (Blanc et al., 1985; Baneijee et al., 1997). The thyroid effects
observed in El Ghawabi et al. (1975) are also supported by mode-of-action data for cyanide,
indicating competitive iodide uptake (see section 4.6). The thyroid enlargement observed in the
HCN-exposed workers likely indicates antagonism of iodine uptake by the cyanide metabolite
SCNThis biological response indicates a stress on the homeostatic mechanisms of the thyroid,
which is of special concern to populations that include individuals with iodine deficiency,
individuals with clinical or subclinical hypothyroidism, and the developing fetus.
The HCN inhalation database is limited by a lack of reliable exposure-response data.
Information on acute human occupational inhalation exposure to HCN does exist but is limited to
case reports of accidental overexposures with unclear exposure concentrations and/or durations.
Additionally, no chronic or subchronic HCN inhalation studies exist in animals. The only
available studies that report exposure data and potential health effects of inhalation of HCN are
occupational exposure studies. In consideration of this limited inhalation database, the El
Ghawabi et al. (1975) study was chosen as the principal study. This study of electroplating
workers at three factories in Egypt included individual breathing zone measurements from the
study participants and reported a strong correlation between these measurements and urinary
levels of SGST. However, the exposure assessment was based on a single 15-minute breathing
zone sample for each worker and the potential for dermal exposure was not explicitly discussed.
Despite the significant weakness of limited exposure assessment, a relatively narrow range in
exposure levels within the study, and lack of quantitative measurement of other exposures in this
work setting, EPA has selected the El Ghawabi et al. (1975) study as the principal study, and
thyroid enlargement and altered iodide uptake were designated as the critical effects. The lowest
mean concentration of HCN reported, 7.1 mg/m3 HCN, is designated as the LOAEL. The choice
of thyroid dysfunction as a critical effect is supported by other epidemiologic studies in a silver-
reclaiming factory (Blanc et al., 1985) and in electroplating workers (Baneijee et al., 1997).
5.2.2. Method of Analysis
A LOAEL, but no NOAEL, was available from the principal study. Because quantitative
concentration-response data were available only for one concentration, benchmark concentration
modeling could not be conducted.
The lowest mean concentration recorded among the three factories evaluated by
El Ghawabi et al. (1975) was 6.4 ppm. Assuming a temperature and pressure of 25°C and
760 mm Hg, this LOAEL in ppm was multiplied by the molecular weight of HCN and divided
by 24.45 to determine the LOAEL in mg/m3 HCN.
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LOAEL (ppm) x 27/24.45 = 6.4 ppm x 27/24.45 = 7.08 mg/m3 HCN
The standard method for adjustment of LOAEL or NOAEL values from occupational
studies was employed as described in U.S. EPA (2002). Because El Ghawabi et al. (1975) did
not report daily exposure durations for exposed workers, an 8 hour/day, 5 day/week exposure
scenario was assumed. A default occupational ventilation rate of 10 m3/8-hour day and a default
ventilation rate for continuous ambient exposure of 20 m3/24-hour day were used. The exposure
was also adjusted to account for the difference between occupational exposure for 5 days/week
versus continuous ambient exposure for 7 days/week. Thus,
LOAEL(ADj) = 7.08 mg/m3 HCN x 10/20 x 5 days/7 days = 2.5 mg/m3 HCN
5.2.3. RfC Derivation—Including Application of Uncertainty Factors (UFs)
The RfC is based on thyroid enlargement and altered iodide uptake reported in an
occupational study of electroplating workers and is supported by other occupational studies and
oral animal studies reporting similar thyroid effects. The LOAEL from this study (adjusted for
continuous exposure) of 2.5 mg/m3 HCN was used as the POD.
A total UF of 3,000 was applied to the POD: 10 for the extrapolation of a LOAEL to a
NOAEL (UFl), 3 for the extrapolation from a subchronic to chronic exposure duration (UFS), 10
for human intraspecies variability (UFH), and 10 to account for database deficiencies (UFD).
An UF of 10 was used for extrapolating from a LOAEL to a NOAEL (UFl) since no
NOAEL was available in the principal study.
An UF of 3 was used to account for extrapolation from a subchronic to chronic exposure
duration. The workers in the principal study were exposed to cyanide for 5-15 years.
Approximately 60% of the workers in this study were exposed for more than 5 years, and 25%
were exposed for more than 10 years. Twenty of the 36 exposed workers had thyroid
enlargement; however, the authors found no correlation between duration of exposure and either
incidence or magnitude of thyroid enlargement in the workers.
An UF of 10 was used to account for variation in susceptibility to cyanide among
members of the human population. Although some information is available on potential
sensitive populations, as described in section 4.7, there are insufficient quantitative data to
inform the UF for human variability with chemical-specific data.
An UF of 10 was applied to account for deficiencies in the cyanide inhalation database.
The database includes limited human data from epidemiologic studies of inhalationally exposed
workers (Blanc et al., 1985; El Ghawabi et al., 1975). The database lacks developmental and
multigenerational reproductive toxicity studies. Inhalation studies on ACH evaluated limited
male and female reproductive endpoints and were negative for impacts on fertility (Monsanto
Co., 1985a, b). Oral studies of cyanide exposure in rodents have suggested that the male
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reproductive tract is a sensitive target of cyanide toxicity following subchronic exposure (NTP,
1993). Due to the proposed cyanide mode of action of thyroid disruption (through the metabolite
thiocyanate), developmental neurotoxicity studies or developmental studies assessing maternal
and fetal thyroid function are also considered potential data insufficiencies.
The RfC for HCN was calculated as follows:
RfC = LOAEL(Adj) - UF
= 2.5 mg/m3 - 3000
= 0.00083 mg/m3 (rounded to 8 x 10^ mg/m3)
It is not recommended that the RfC for HCN be extrapolated to cyanide salts due to
inhalation considerations. Specifically, exposure to HCN occurs as a gas, whereas the extremely
high boiling points and vapor pressures of cyanide salts predict that inhalation exposure would
occur as aerosols. Different dosimetric approaches would apply to the aerosol (or particle)
exposures that would result from exposure to cyanide salts, compared with exposure to HCN gas.
5.2.4. Previous RfC Assessment
An RfC for HCN of 3 x 10 3 mg/m3 was posted on the IRIS database in 1994. This RfC
was based on findings of thyroid effects and neurological symptoms in workers from the study
by El Ghawabi et al. (1975). The POD for this RfC was based on an adjusted LOAEL of 7.07
mg/m3. The LOAEL was divided by a total UF of 1000 comprised of UFs of 10 each to account
for the lack of a NOAEL and intrahuman variability. UFs of 3 each were applied to account for
the use of a study of less than chronic duration and deficiencies in the database (lack of chronic
and multigenerational reproduction studies). Since the posting of the previous IRIS RfC for
cyanide, no new chronic or subchronic inhalation studies for HCN with quantitative information
are available in the literature. However new data are available which evaluate potential health
effects at low levels of exposure i.e., perturbations of thyroid function in pregnant women and
their offspring (see Section 4.8.1). Specifically, these studies show increased pregnancy
complications and decrements in learning and memory in offspring of women with subclinical
hypothyroidism (Kooistra et al., 2006; Casey et al., 2005; Pop et al., 2003; Haddow et al., 1999).
The current RfC for cyanide was derived from the same study as the past assessment based on
the observation of altered thyroid function (as indicated by iodine uptake) in occupationally
exposed male workers. The change in the selection of the database UF was based on this new
data and consideration of the MOA of SCN.
5.3. UNCERTAINTIES IN THE ORAL REFERENCE DOSE AND INHALATION
REFERENCE CONCENTRATION
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The following discussion identifies uncertainties associated with the quantification of the
RfD and RfC for cyanide. Following EPA practices and guidance (U.S. EPA, 1994b, 1993), the
UF approach was applied to the chosen PODs to derive an RfD and RfC (see sections 5.1.3 and
5.2.3). Factors accounting for uncertainties associated with a number of steps in the analyses
were adopted to account for extrapolating from an animal study to human exposure, a diverse
human population of varying susceptibilities, and database deficiencies.
The database for cyanide includes limited human data from studies of occupationally
exposed workers. Endpoints observed in inhalationally exposed workers include altered thyroid
function and CNS symptoms (including headache, weakness, nausea, and vomiting). The
database also includes oral exposure studies in laboratory animals, including limited chronic
studies, subchronic dietary exposure studies, and several developmental studies, including one
specifically assessing gross and microscopic brain morphology in rats (as discussed in
section 4.3). Effects seen with low-dose oral exposure to cyanide include decreased reproductive
organ weight, decreased spermatid concentration, increased thyroid weight, and histologic
alterations in the CNS, kidney, testis, and adrenal glands. In addition to oral and inhalation data,
the database for cyanide includes studies on absorption, distribution, metabolism, and excretion.
Uncertainty exists in the selection of the most relevant animal species for human health
assessment. Studies in several species, including rodents, pigs, goats, and dogs, were considered
in the development of the RfD; however, limited data exist on differential species' sensitivity to
cyanide, especially in the context of long-term exposure.
The RfD was derived from a BMDL of 1.8 mg/kg-day, which was based on the
observation of decreased epididymis weight in male F344 rats exposed to NaCN in drinking
water for 13 weeks (NTP, 1993). This study treated male and female rats with doses of CN
ranging from 0.16 to 12.5 mg/kg-day. Other reproductive effects observed at higher doses
included decreased caudal epididymis and testis weights and decreased spermatid count. After
consideration of all potential PODs, the RfD of 6 x 10 4 mg/kg-day was based on the observation
of decreased cauda epididymis weight in male F344 rats following subchronic dietary
administration of cyanide (NTP, 1993).
The mode of action of the decrease in cauda epididymis weight in rats is uncertain,
though limited information from other model systems of hypothyroid animals suggests that it
may be related to thyroid disruption from the primary metabolite thiocyanate (see section 4.5.4).
However, the critical study used as the basis for the RfD (NTP, 1993) did not assess thyroid-
related parameters, such as T4, T3, TSH, or thyroid weight. Therefore, it is not known whether
direct indicators of disruption of thyroid homeostasis accompanied the observed reproductive
effects.
Additional studies exist that determined different effects at lower doses, as discussed in
section 5.1.1, including behavioral changes and decreased serum T4 in pigs (Jackson, 1988) and
kidney, adrenal, and testicular effects in dogs (Kamalu, 1993). Selection of either of these
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studies would result in a lower RfD as portrayed in Table 5-3 and Figure 5-1. Ultimately, these
studies were deemed of lower confidence, due to issues concerning study design and reporting
(see section 5.1.1), than the 3-month dietary study in rats and mice conducted by NTP (1993)
and thus were not chosen as the principal study. To derive the RfD, UFs were applied to the
POD determined through BMD modeling of the critical effect of reduced cauda epididymis
weight in male rats. This study was well designed and conducted with several dose levels,
sufficient numbers of animals, and a wide range of tissues and endpoints assessed; however,
significant areas of uncertainty exist in the animal data relied upon for the RfD. UFs associated
with the extrapolation from the POD derived from an animal study to a diverse human
population of varying susceptibilities were applied.
Uncertainty exists in the selection of the BMR level utilized in the BMD modeling of the
critical effect (decreased cauda epididymis weight) to determine the POD. At this response level
in the cauda epididymis, no alteration in epididymal sperm count was detected, and thus fertility
in these animals would not likely be affected. However, human males have markedly lower
fertility than do rats (U.S. EPA, 1996; Working et al., 1988) and thus changes in male
reproductive endpoints may be expected to have greater impact in humans as compared to
rodents. In the absence of clear information to determine the level of change in cauda
epididymis weight related to a biologically significant change, a decrease of one SD in organ
weight was selected to represent a minimally adverse change.
The choice of BMD model is not expected to introduce a considerable amount of
uncertainty since one SD in the reduction of cauda epididymis weight is within the observable
range of the data. Other available continuous, constant variance models available in the BMD
software (i.e., the power and linear models) also showed acceptable fits (p value > 0.1) and had
AIC values within 1 unit of the selected polynomial model. However, the polynomial model had
superior visual fit to the data, especially in the low-dose range. The BMDL estimates for various
models are not within a factor of 3, indicating some model dependence; therefore, in accordance
with the current Benchmark Dose Technical Guidance Document (External Review Draft, U.S.
EPA, 2000b), the model with the lowest BMDL estimate was selected.
Additional BMD modeling for other data sets, including additional reproductive
endpoints from the NTP (1993) study, was also conducted to provide other PODs for comparison
purposes (see Appendix B). A graphical representation of these and other potential PODs and
resulting RfVs is shown in Figure 5-1 (see section 5.1.4).
The default UF of 10 for the extrapolation from animals to humans accounts for
toxicokinetic differences and toxicodynamic differences. Physiologically based toxicokinetic
models can be useful for the evaluation of interspecies toxicokinetics; however, the cyanide
database lacks an adequate model that would inform potential differences. Data from workers
occupationally exposed to cyanide provide some information on the absorption, metabolism, and
elimination of cyanide in humans and indicate qualitatively that the toxicokinetics of cyanide are
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similar between humans and animals. Additionally, some biological effects, including thyroid
enlargement and neurological symptoms observed in animals and humans (such as ataxia,
weakness, and behavioral changes) are similar in nature, indicating similar toxicodynamics.
However, the magnitude of the similarities or differences in toxicokinetic and toxicodynamic
parameters cannot be calculated due to uncertainties regarding routes of exposure and doses for
the occupationally exposed workers. Therefore, a UF of 10 to account for interspecies
differences was used.
An UF of 3 was applied to account for less than chronic exposure duration in the
occupationally exposed workers. Uncertainty exists as to whether additional or more severe
effects would be expected in these workers over longer durations.
Limited data exist on effects of cyanide in populations of occupationally exposed workers.
However, since potential variability in responses to cyanide in the greater human population is
unknown, the default UF of 10 for intrahuman variability was used. Human variation may be
larger or smaller; however, chronic cyanide-specific data to examine the potential magnitude of
human variability of response were not found.
Uncertainties associated with data gaps in the cyanide database have been identified.
Effects on reproductive organ weight and sperm parameters have been identified in rats and mice
subchronically exposed to cyanide in the diet. However, data more fully characterizing potential
multigenerational reproductive effects are lacking. Gross developmental effects have not been
observed in the few, limited developmental studies available (Imosemi et al., 2005; Malomo et
al., 2004; Tewe and Maner, 1981). Due to thiocyanate's proposed mode of action of competitive
inhibition of iodide uptake in the thyroid, the lack of studies evaluating subtle
neurodevelopmental and behavioral outcomes adds uncertainty to this assessment. It is unclear
from the available database whether perturbation of thyroid function sufficient for the induction
of subclinical or clinical hypothyroidism would be expected to occur below the POD for reduced
epididymis weight in rats identified in NTP (1993), since thyroid hormone levels were not
measured in this study (though examination of the thyroid showed no increase in weight or
histologic lesions). Therefore, a UF of 3 for database deficiencies was applied to the POD to
account for uncertainty regarding potential neurological effects during development and for the
lack of data on multigenerational reproductive toxicity.
The lack of specific immune-related data on cyanide represents a data gap. Subchronic
and chronic studies on cyanide and cyanide-related compounds have evaluated limited immune
endpoints, such as organ weights, histopathology (NTP, 1993; Monsanto Co., 1985a, b; Lewis et
al., 1984; Howard and Hanzal, 1955), and hematological parameters (NTP, 1993; Blanc et al.,
1985; Monsanto Co., 1985a, b; Howard and Hanzal, 1955), and are generally negative (see
section 4.4.4). One study found an elevation in percent lymphocytes in exposed workers as
compared to controls (El Ghawabi et al., 1975). This finding is of unclear significance,
considering the nonspecific nature of this hematological parameter and the lack of support for
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this finding in the database. Overall, though the examined immune endpoints in the cyanide
database appear normal, the lack of functional immune assays precludes a confident conclusion
regarding potential immune toxicity of cyanide.
The HCN inhalation database is limited by a lack of reliable exposure-response data.
Additionally, no chronic or subchronic HCN inhalation studies exist in animals. The only
available studies reporting exposure data and potential health effects of inhalation of HCN are
occupational exposure studies. In consideration of this limited inhalation database, El Ghawabi
et al. (1975) was selected as the most appropriate study for the derivation of the RfC.
The RfC was derived from a LOAEL(adj) of 2.5 mg/m3 HCN, which was based on
thyroid enlargement and altered iodide uptake in a cohort of workers in three electroplating
facilities who had been exposed to HCN for 5-15 years (El Ghawabi et al., 1975). This study is
the only extended duration epidemiologic study in which concurrent exposure concentrations
were measured. The mean cyanide air concentrations in the breathing zone of workers at the
three plants were 7.1 to 11.5 mg/m3 HCN. The lowest mean concentration recorded in the three
factories, 7.1 mg/m3 HCN, is designated as a LOAEL. Twenty male volunteers of the same age
group and socioeconomic status who had no occupational exposure to cyanide were chosen as
controls. Effects observed in exposed workers included thyroid enlargement and increased
iodide uptake in the thyroid. The effects observed are consistent with known effects of cyanide
and reported effects in other studies of cyanide-exposed workers and are also supported by
similar effects observed in animals orally exposed to cyanide. However, significant areas of
uncertainty exist in the human data relied upon for the RfC.
Some uncertainties exist in the exposure doses measured. The individual breathing zone
concentrations of HCN were measured in the three different factories over a period of 2 months.
No information was given regarding how the current cyanide environmental monitoring data
may compare to conditions over the last several years. Furthermore, the authors acknowledged
that workers were co-exposed to other chemicals during the electroplating process, including
gasoline, alkali, and acid, but did not quantitate the magnitude of these exposures. It is unclear
how these exposures would be expected to impact the effects observed in El Ghawabi et al.
(1975). However, the thyroid effects, specifically the thyroid enlargement, are supported by
human and animal data and are consistent with the mode of action of the cyanide metabolite
SCN (see section 4.6). Additionally, as with most occupational exposure scenarios, the
possibility exists for exposure through the dermal route. However, Table 4-1 provides evidence
indicating that the individual breathing zone measurements of HCN closely correlated with an
internal measure of exposure (urinary SCN~).
No NOAEL was identified in the El Ghawabi et al. (1975) study. Effects in this study
were found at the lowest exposure concentrations measured; therefore, the LOAEL determined
for this study does not indicate where a threshold of effects would lie and the data provided in
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the study are not sufficient for a dose-response analysis. To account for the uncertainty in the
use of a LOAEL for the POD, a factor of 10 was applied in the derivation of the RfC.
Several assumptions were made in the conversion of the LOAEL observed in El Ghawabi
et al. (1975) to an adjusted dose. The average temperatures in the factories were not included in
the study; therefore, conversion from ppm to mg/m3 exposure concentrations were based on
standard temperature and pressure. Additionally, the daily and weekly cyanide exposure
durations were not explicitly stated in the study; therefore, an 8 hour/day, 5 day/week exposure
was assumed. Other uncertainties in the exposure assessment of the workers include potential
variability in exposures among workers based on specific duties or locations in the factories;
however, this uncertainty is limited since breathing zone samples from individual workers were
averaged to determine mean factory exposure.
Although significant areas of uncertainty remain in the human epidemiologic data relied
upon for the RfC, the use of human exposure data eliminates the substantial uncertainty inherent
in the extrapolation of an animal study to humans. The study by El Ghawabi et al. (1975) was
conducted by using a small population of male workers (n = 36) and cannot be expected to
capture the human variability of response to cyanide exposure. Additionally, the workers
included in the study may represent a low-sensitivity group with other more affected workers not
continuing employment (i.e., the healthy worker effect). Therefore, in the absence of cyanide-
specific data to account for the heterogeneity of human sensitivity, a factor of 10 was used to
account for uncertainty associated with human variation in the derivation of the RfC. Human
variation in response to cyanide exposure may be larger or smaller; however, chemical-specific
data to assess the potential magnitude of variability are unavailable. Of the 36 workers, average
exposure duration was about 7 years, with the minimum exposure duration being 5 years and the
maximum being 15 years.
Uncertainty exists regarding whether progression of effects would be expected with
longer exposure time. Twenty of the 36 exposed workers had thyroid enlargement rated as being
mild to moderate; however, the authors found no correlation between duration of employment at
the factory and either incidence or magnitude of enlargement. Additionally, it is not known
whether CNS symptom incidence or severity would be expected to increase with increasing
exposure duration or whether CNS effects would be detected in chronically exposed workers at
lower concentrations. Additionally, it is possible that different sensitive endpoints may be
detected in studies of longer duration. For instance, chronic inhalation exposure in workers at a
metal tempering plant indicated some deterioration in pulmonary function (Chatgtopadhyay et al.,
2000). However, no exposure monitoring from this study was available for dose-response
comparison to the El Ghawabi et al. (1975) study, identified as the principal study. Therefore, to
account for uncertainties regarding the exposure duration of the POD, a UF of 3 was applied.
Uncertainties associated with data gaps in the cyanide database have been identified. The
database includes limited human data from epidemiologic studies of occupationally exposed
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workers. No animal studies exist that employed extended duration inhalation exposure to HCN,
although inhalation studies on the related compounds ACH and (CN)2 exist. Studies to assess
developmental or multigenerational reproductive toxicity of HCN for the inhalational route of
exposure are not available. Thus, a UF of 10 was applied to account for limitations in the
inhalation database.
5.4. CANCER ASSESSMENT
The only available chronic study of cyanide that analyzed a wide variety of tissues is an
oral rat study (Howard and Hanzal, 1955); no tumors or lesions were associated with either dose
group following dietary administration of cyanide at doses up to 10.8 mg/kg-day for 2 years.
This study is limited by small sample sizes, histopathologic assessment of only a subset of
potential target organs of carcinogenicity, and uncertainty regarding dose due to volatility.
Overall, the data are inadequate for an assessment of the human carcinogenic potential of
cyanide, based on EPA's "Guidelines for Carcinogenic Risk Assessment' (U.S. EPA, 2005a).
Therefore, no quantitative cancer assessment was conducted.
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6. MAJOR CONCLUSIONS IN CHARACTERIZATION OF
HAZARD AND DOSE RESPONSE
6.1. HUMAN HAZARD POTENTIAL
Cyanide compounds are used in a number of industrial processes, including mining,
metallurgy, manufacturing, and photography, due to their ability to form stable complexes with a
range of metals. Cyanide has been employed extensively in electroplating, in which a solid
metal object is immersed in a plating bath containing a solution of another metal with which it is
to be coated in order to improve the durability, electrical resistance, and/or conductivity of the
object. HCN has also been used in gas chamber executions and in chemical warfare. The
cyanide salts NaCN and KCN have also been used as rodenticides. Use in industrial processes is
the main origin of cyanide in the environment, but cyanide is also released from biomass burning,
volcanoes, and natural biogenic processes from higher plants, bacteria, and fungi (ATSDR,
2006). Additionally, cyanogenic compounds, which are converted to cyanide in the body,
naturally occur in many plant foods, including cassava root, almonds, millet sprouts, lima beans,
soy, spinach, bamboo shoots, and sorghum. Exposure to cyanide also occurs from smoking.
Thiocyanate, the primary metabolite of cyanide, is found in plasma or blood at approximately
0.5-4 |ig/L in nonsmokers and approximately 6-22 |ig/L in smokers (Chandra et al., 1980).
The available data show that cyanide is rapidly and extensively absorbed via the oral,
inhalation, and dermal routes, although quantitative data on the percent or extent of absorption
are limited. At physiological pH, cyanide is distributed in the body as HCN, and thus the
toxicokinetics for freely dissociating cyanide compounds are the same. Cyanide distributes
rapidly and fairly uniformly throughout the body following absorption. Inhaled or dermally
absorbed HCN enters the systemic circulation immediately. In contrast, ingested cyanide is
primarily converted to thiocyanate via first-pass metabolism in the liver. Immediately following
oral exposure in humans, tissues containing cyanide included the liver, brain, spleen, blood,
kidneys, and lungs (Ansell and Lewis, 1970; Gettler and Baine, 1938). Following acute
inhalation exposure in humans and animals, cyanide is found in the lung, heart, blood, kidneys,
and brain (Ballantyne, 1983; Gettler and Baine, 1938). The major metabolic pathway for
cyanide is conversion to thiocyanate, primarily by rhodanese. Detoxification of cyanide by
rhodanese is rapid with the concentration of sulfur-containing donor molecules as the rate-
limiting factor. Rhodanese is widely distributed throughout the body but is located at the highest
concentration in the liver. Toxicokinetic studies in animals indicate rapid decreases in cyanide
blood concentration within 3 hours following dosing, with the half-life of elimination for
thiocyanate for all species about 10 times longer (Sousa et al., 2003). Cyanide is primarily
excreted in the urine as thiocyanate following both inhalation and oral exposures and is not
thought to accumulate in the blood and tissues.
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Several reports on occupationally exposed workers indicate that chronic inhalational
exposure to low concentrations of cyanide can cause thyroid effects and CNS symptoms
(Banerjee et al., 1997; Blanc et al., 1985; El Ghawabi et al., 1975). The results of these
occupational studies suggest that chronic exposure to cyanide may be associated with alterations
in thyroid gland function, including enlargement, altered iodine uptake, and decreased thyroid
hormones, and subjective CNS symptoms. Another study also suggests that chronic exposure to
cyanide fumes in a metal-tempering plant may reduce pulmonary function in chronically exposed
workers (Chatgtopadhyay et al., 2000). Chronic or subchronic inhalation studies of HCN in
experimental animals were not found.
Epidemiologic studies of populations in developing countries consuming cyanogenic
compounds in food have been conducted (Madhusudanan et al., 2008; Oluwole et al., 2003;
Osuntokun, 1973; Makene and Wilson, 1972). These studies are confounded by the presence of
other potentially toxic dietary components associated with cyanogenic foods, such as the
cyanogenic glycoside linamarin, and the high prevalence of iodine, protein, and vitamin
deficiencies in the studied populations. Because of the aforementioned challenges, use of
epidemiologic studies of human dietary cyanogenic exposure is limited for the purposes of this
hazard assessment for cyanide.
No epidemiologic studies exist of long-term human exposure to cyanide by the oral route.
Information on human oral exposure to cyanide is limited to acute effects following suicide
attempts or accidental poisoning. Acute oral exposure to cyanide has been observed to result in
typical signs of cyanide poisoning, including CNS depression, convulsions, coma, and death.
Chronic and subchronic oral studies in experimental animals indicate that the thyroid, CNS, and
male reproductive organs are sensitive targets of cyanide toxicity (Manzano et al., 2007; Soto-
Blanco et al., 2002a, b; NTP, 1993; Jackson, 1988).
Histologic changes in the CNS have been observed following longer-term exposure to
cyanide in some animal models. In rats exposed to cyanide in the diet for 1 year, increased
vacuolation in the spinal cord white matter and exacerbation of methionine deficiency-induced
spinal cord demyelination were observed (Philbrick et al., 1979). In addition, histopathologic
effects, including neuronal loss, spheroids, damaged Purkinje cells, and loss of white matter in
various CNS structures, were observed in rats following a 12-week oral exposure period (Soto-
Blanco et al., 2002a) and in goats following 5 months of oral exposure (Soto-Blanco et al.,
2002b). However, histopathology of the nervous system has not been identified in other chronic
or subchronic studies (NTP, 1993; Howard and Hanzal, 1955) conducted at doses lower than that
administered by Philbrick et al. (1979). In addition to histologic changes observed in some
studies, subtle behavioral changes were noted in pigs orally exposed to 1.2 mg/kg-day cyanide in
drinking water for 6 months.
Chronic and subchronic exposure to cyanide is known to induce thyroid effects due to the
cyanide metabolite, thiocyanate. Thiocyanate adversely affects the thyroid gland via competitive
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inhibition of iodide uptake and perturbation of the homeostatic feedback mechanisms that
regulate the synthesis and secretion of essential thyroid hormones. Philbrick et al. (1979)
reported decreased serum T4 levels and increased thyroid weights in rats but no histopathologic
changes in the thyroid gland. Subchronic studies in rats and mice (NTP, 1993) conducted with a
range of doses lower than the single dose tested by Philbrick et al. (1979) did not observe
adverse histopathology or increased weight of the thyroid gland, though thyroid hormone levels
were not evaluated. Studies in pigs have noted increased thyroid weights, altered thyroid
histology, and decreased thyroid hormones (Manzano et al., 2007; Jackson, 1988) at doses in the
range of the NTP (1993) study and several times lower than the Philbrick et al. (1979) study. It
is apparent through comparisons of thyroid effects in animal models that sensitivity of the
thyroid to the effects of cyanide appears to vary widely among species.
Adverse reproductive effects, including decreased epididymis, cauda epididymis, and
testis weights and decreased sperm parameters (epididymal sperm motility and testicular
spermatid counts), have been observed in rats in a subchronic dietary study by NTP (1993).
Decreases in the cauda epididymis and epididymis weights were also seen in mice (NTP, 1993).
Histologic examination of reproductive organs did not reveal any lesions. Additionally,
reproductive effects, specifically, alterations in testicular histology, have also been observed in a
14-week study in dogs (Kamalu, 1993). The mode of action of the reproductive effects
following subchronic cyanide exposure in rodents is unclear, though some data in hypothyroid
animal models suggest that these effects may be secondary to thyroid perturbation.
Cyanide exerts its acute effects, including CNS depression, convulsions, coma, and death,
by binding with cytochrome c oxidase, a key enzyme in the production of ATP by way of
oxidative phosphorylation. The steep dose response occurring with acute high-dose exposures is
thought to be due to cyanide overload, resulting in saturation of detoxification pathways that
metabolize cyanide to less acutely toxic intermediate compounds. At lower dose rates, an
efficient detoxification system (primarily via rhodanese with sulfur donors as the rate-limiting
factor) catalyzes the transformation of cyanide to thiocyanate, its primarily metabolite.
Thiocyanate is not known to be acutely toxic, although long-term exposures can adversely affect
the thyroid gland via iodide uptake inhibition and decreased thyroid hormone synthesis. The
chronic effects of cyanide and thiocyanate on other organ groups are not clear.
6.2. DOSE RESPONSE
6.2.1. Noncancer—Oral
The male reproductive effects observed by NTP (1993) have been identified as the
critical effects for the development of the cyanide RfD. The NTP (1993) study was well
designed with five treatment groups with doses spanning two orders of magnitude. Numerous
tissues and endpoints were assessed in both rats and mice. This study identified a suite of
statistically significant reproductive effects in both species, including decreased epididymis
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weights (cauda and whole), decreased testes weight, and altered sperm parameters. BMD
analysis of the observed reproductive data from rats and mice indicated decreased cauda
epididymis weight to be the most sensitive reproductive effect observed. Thus, decreased cauda
epididymis weight was chosen as the critical effect. This effect is believed to be one which
would likely precede substantial decrements in sperm parameters and fertility in this test species.
The cyanide database contains additional, limited support for the specific endpoint of
reproductive toxicity. Altered testicular histopathology, including a significantly decreased
percentage of tubules in stage VIII of the spermatogenic cycle, was observed in dogs ingesting 1
mg/kg-day cyanide for 14 weeks (Kamalu, 1993).
In addition to the reproductive effects observed in NTP (1993), other sensitive effects
observed in animals, including increased thyroid weight, altered thyroid hormones, and altered
testicular, kidney, and adrenal histopathology, were also considered as potential critical effects
(see discussion in section 5.1.1). Though these effects were not ultimately selected for the
derivation of the RfD, RfDs for these endpoints were quantified for comparison purposes.
BMD modeling was conducted to calculate potential PODs for deriving the RfD by
estimating the effective dose at a specified level of response (BMDX) and its 95% lower bound
confidence limit (BMDLX). A BMR level was selected corresponding to a change in the mean
response equal to one control SD from the control mean for cauda epididymis weight. In this
case, a one SD change in cauda epididymis weight was selected under an assumption that it
represents a minimal biologically significant change. Using the best fitting model for this data
set, a one SD change was equivalent to a 7% decrease in cauda epididymis weight. Additional
BMD modeling for other amenable data sets was also conducted to provide other PODs for
comparison purposes (see Appendix B). PODs for these endpoints and other PODs determined
through a NOAEL/LOAEL approach were considered for the derivation of the RfD. Tabular and
graphical representations of these potential PODs and resulting RfVs are shown in Table 5-2 and
Figure 5-1, respectively.
The RfD of 6 x 10^ mg/kg-day was calculated from a BMDLisd of 1.9 mg/kg-day based
on decreased cauda epididymis weight in rats in the subchronic oral study conducted by NTP
(1993). A total UF of 3,000 was applied to the POD: 10 for the extrapolation of a LOAEL to a
NOAEL (UFl), 10 for the extrapolation from a subchronic to chronic exposure duration (UFS);
10 for human intraspecies variability (UFH); and 3 to account for database deficiencies (UFD).
Information was unavailable to quantitatively assess toxicokinetic or toxicodynamic differences
between animals and humans and the potential variability in human susceptibility; thus, the
interspecies and intraspecies UFs of 10 were applied. To address the subchronic, 13-week study
duration of the principal study, a UF of 10 (UFS) was also applied. Additionally, a threefold
database UF was considered necessary due to the lack of information regarding potential
multigenerational reproductive effects and the lack of a sensitive neurodevelopmental study.
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The overall confidence in the RfD is medium. Confidence in the principal study (NTP,
1993) is medium. This study was well conducted, involved a sufficient number of animals per
group (including both sexes of two species), used several dose levels, and assessed a wide range
of tissues and endpoints. Confidence in the database is medium. The cyanide database includes
occupational inhalation exposure studies in humans, chronic and subchronic dietary exposure
studies in laboratory animals, and several developmental studies in laboratory animals. However,
the database is lacking a multigenerational reproductive toxicity study and a sensitive
neurodevelopmental study. Therefore, reflecting medium confidence in the database and
medium confidence in the principal study, the overall confidence in the RfD is medium.
6.2.2. Noncancer—Inhalation
No new chronic or subchronic inhalation studies for HCN (with quantitative information)
have been published in the literature since the development of the previous HCN RfC. Therefore,
the inhalation database was reevaluated, but the principal study and critical effect selected for the
RfC were unchanged. The current RfC is based on an occupational study reporting thyroid
enlargement and altered iodide uptake. The principal study (El Ghawabi et al., 1975) identified
effects on the thyroid that were consistent with the proposed mode of action of cyanide. Only a
LOAEL could be determined from this study. In addition, there was potential coexposure of the
workers to other substances, including gasoline, alkali, and hydrochloric acid, through the
electroplating process. Nonetheless, the reported thyroid alterations are consistent with the
reported effects of cyanide exposure in other occupational studies and animal studies.
The RfC of 8 x 10"4 mg/m3 was derived from a LOAEL(Adj) of 2.5 mg/m3 HCN, which
was based on thyroid enlargement and altered iodide uptake in a cohort of workers in three
electroplating facilities who had been exposed to HCN for 5-15 years (El Ghawabi et al., 1975).
The authors recorded multiple individual breathing zone samples and reported the mean HCN
exposure level for each factory. The lowest mean HCN concentration was designated as the
LOAEL. The study authors did not report the data in a manner that allowed evaluation of an
exposure duration response or a concentration response. Other studies of occupationally exposed
workers either did not provide exposure data or included higher exposure levels and did not
control for confounding variables.
A total UF of 3,000 was applied to the POD: 10 for the extrapolation of a LOAEL to a
NOAEL (UFl), 3 for the extrapolation from a subchronic study (UFS), 10 for human intraspecies
variability (UFH), and 10 to account for database deficiencies (UFD). The occupational study by
El Ghawabi et al. (1975) identified a LOAEL but no NOAEL. Therefore, a UF of 10 was
applied (UFl). Additionally, a UF was deemed necessary to account for the exposure duration of
the studied workers that ranged from subchronic to chronic. Because greater than half of the
workers were exposed for over 5 years, and 25% of workers were exposed for at least 10 years, a
partial UF of 3 was applied (UFS) to account for the less than chronic exposure duration of a
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subset of workers. Information was unavailable to quantitatively assess the variability in
susceptibility to cyanide in the human population; therefore, a UF of 10 for intraspecies
variability was applied (UFH). A UF of 10 was applied for uncertainties in the database,
specifically the lack of a multigenerational toxicity study and a sensitive neurodevelopmental
study. Several limited occupational inhalation studies are available in the cyanide database. No
chronic or subchronic inhalation studies evaluating a wide array of endpoints are available in
experimental animals, though inhalation studies on the related compounds (CN)2 and ACH exist.
However, due to the lack of data evaluating potential multigenerational reproductive toxicity and
neurodevelopmental toxicity, a database UF of 10 (UFD) was applied.
Reflecting low confidence in the principal study (El Ghawabi et al., 1975) and low to
medium confidence in the inhalation database, the overall confidence in the cyanide RfC is low.
6.2.3. Cancer
Cyanide has not been subjected to a complete standard battery of genotoxicity assays,
though, overall, the available data indicate that cyanide is not genotoxic. No adequate
carcinogenicity studies of cyanide are available in animals or humans. In a 2-year study in rats,
no evidence of tumorigenicity was observed (Howard and Hanzal, 1955). However, the number
of animals per dose group limited the power of the study and only a limited set of target tissues
was evaluated histopathologically. Based on these considerations and in accordance with the
U.S. EPA (2005a) "Guidelines for Carcinogen Risk Assessment," the data are inadequate for an
assessment of the human carcinogenic potential of cyanide.
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Walton, DC; Witherspoon, MG. (1926) Skin absorption of certain gases. J Pharmacol Exp Ther 26:315-324.
Way, JL. (1984) Cyanide intoxication and its mechanism of antagonism. Annu Rev Pharmacol Toxicol 24:451-481.
Westley, JL. (1981) Thiosulfate-cyanide sulfur-transferase (rhodanese). Methods Enzymol 77:285-291.
WHO (World Helath Organization). (1994) Indicators for assessing iodine deficiency disorders and their control
through salt iodization. WHO/NUT/94.6. Geneva: International Council for the Control of Iodine Deficiency
Disorders.
Wing, DA; Baskin, SI. (1992) Modifiers of mercaptopyruvate sulfurtransferase catalyzed conversion of cyanide to
thiocyanate in vitro. J Biochem Toxicol 7(2):65-72.
Wistuba, J; Mittag, J; Luetjens, CM; et al. (2007) Male congenital hypothyroid Pax8-/- mice are infertile despite
adequate treatment with thyroid hormone. J Endocrinol 192(1):99-109.
Wolff, J. (1998) Perchlorate and the thyroid gland. Pharmacol Rev 50(1):89-105.
Wolfsie, JH; Shaffer, CB. (1959) Hydrogen cyanide. Hazards, toxicology, prevention and management of poisoning.
JOccupMed l(5):281-288.
Wood, JL. (1975) Biochemistry. In: Newman, AA; ed. Chemistry and biochemistry of thiocyanic acid and its
derivatives. London; New York, NY: Academic Press; pp. 156-221.
Wood, JL; Cooley, SL. (1956) Detoxication of cyanide by cystine. J Biol Chem 218(l):449-457.
Working, PK. (1988) Male reproductive toxicology: comparison of the human to animal models. Environ Health
Perspect 77:37-44.
Yamamoto, K; Yamamoto, Y; Hattori, H; et al. (1982) Effects of routes of administration on the cyanide
concentration distribution in the various organs of cyanide-intoxicated rats. Tohoku J Exp Med 137(l):73-78.
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APPENDIX A. SUMMARY OF EXTERNAL PEER REVIEW AND PUBLIC
COMMENTS AND DISPOSITION
[place holder]
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APPENDIX B. BENCHMARK DOSE MODELING RESULTS
Decreased Absolute Cauda Epididymis Weight in Rats Exposed to NaCN in Drinking Water for
13 Weeks (NTP, 1993)
Table B-l. Decreased cauda epididymis weight in F344 rats following
administration of NaCN in drinking water for 13 weeks
Male rats
0 ppm
30 ppm
100 ppm
300 ppm
Dose (mg/kg-day)
0
1.4
4.5
12.5
Weight (g) ± SI)
Cauda epididymis, absolute
0.162 ±0.009
0.150 ±0.013
0.148 ±0.013
0.141 ±0.009
All models for continuous variables available in the EPA BMDS version 1.4.1c, except
the Hill model, were fit to the data in the Table B-l. The Hill model was not fit to these data
because fitting of the Hill model requires the estimation of four parameters (i.e., intercept, v, n,
and k), which necessitates having a minimum of five dose groups in order to have adequate
degrees of freedom for testing model fit. The NTP (1993) study has only four dose groups, and
thus the Hill model could not be fit to these data. All models fit were constant variance models.
All models tested provided adequate fit to the data, based on the summary results reported by the
BMDS output and visual examination of the graphs. A summary of the goodness-of-fit statistics
for the tested models and resulting BMD and BMDL is presented in Table B-2.
Table B-2. BMD modeling results for decreased cauda epididymis weight in rats
Study
Endpoint
Model
AIC
Goodness-of-
fit p value
BMD
BMDL
NTP (1993);
male rats
Cauda epididymis
weight (absolute)
Linear (1° polynomial)
-312.75
0.08
8.4
5.6
Polynomial (2°)
-313.10
0.11
3.5
1.9
Power
-312.75
0.08
8.4
5.6
All models adequately described the data as demonstrated by goodness-of-fitp values >
0.1. Of the available models, the polynomial (2°) was selected, based on having the lowest AIC
value. Visual inspection reveals that the model describes the data well.
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Polynomial Model with 0.95 Confidence Level
Polynomial
0.17
0.165
0.16
O)
w
c
0
Q.
W
O)
01
0.155
0.15
c
CO
O)
0.145
0.14
0.135
BMD
iBMD
0
2
4
6
8
10
12
dose
16:00 02/13 2008
Figure B-l. Observed and predicted decrease in cauda epididymis weight in
F344 rats following administration of NaCN in drinking water for 13 weeks.
The computer output for the polynomial model of decreased (absolute) cauda epididymis weight
follows:
Polynomial Model. (Version: 2.12; Date: 02/20/2007)
Input Data File: C:\BMDSl-4-lC\UNSAVEDl.(d)
Gnuplot Plotting File: C:\BMDSl-4-lC\UNSAVEDl.plt
Wed Feb 13 16:00:34 2008
BMDS MODEL RUN
The form of the response function is:
Y[dose] = beta 0 + beta l*dose + beta 2*doseA2 + . . .
Dependent variable = MEAN
Independent variable = COLUMN1
rho is set to 0
Signs of the polynomial coefficients are not restricted
A constant variance model is fit
Total number of dose groups = 4
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Default Initial Parameter Values
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alpha =
rho =
beta 0 =
beta 1 =
beta 2 =
0. 000125
0
0. 159311
-0. 00377087
0.0001857
Specified
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -rho have been estimated at a boundary point,
or have been specified by the user, do not appear in the correlation matrix )
alpha
beta 0
beta 1
beta 2
alpha
1
-8 . 4e-010
-3 . 3e-011
4 . 3e-010
beta 0
. 4e-010
1
-0.72
0. 61
beta 1
-3.3e-011
-0.72
1
-0.91
beta 2
4 . 3e-010
0. 61
-0.91
1
Parameter Estimates
Interval
Variable
Limit
alpha
0.00017266
beta_0
0.165233
beta_l
0.000635079
beta_2
0. 000419578
Estimate
0.000120048
0.159311
-0.00377087
0.0001857
Std. Err.
2.68434e-005
0.00302151
0.00159992
0.000119328
95.0% Wald Confidence
Lower Conf. Limit Upper Conf.
6.74354e-005
0.153389
-0.00690665
-4 . 81776e-005
Table of Data and Estimated Values of Interest
Dose
Obs Mean
Est Mean Obs Std Dev Est Std Dev Scaled Res.
0
1.4
4.5
12.5
10
10
10
10
0.162
0.15
0.148
0.141
0.159
0.154
0.146
0.141
0.009
0.013
0.013
0.009
0. 011
0. 011
0. 011
0. 011
0.776
-1.27
0.548
-0.0551
Model Descriptions for likelihoods calculated
Model A1: Yij = Mu(i) + e(ij)
Var{e(ij)} = SigmaA2
Model A2: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma(i)A2
Model A3: Yij = Mu(i) + e(ij)
Var{e(ij)} = SigmaA2
Model A3 uses any fixed variance parameters that
were specified by the user
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Model R: Yi = Mu + e(i)
Var{e(i)} = SigmaA2
Likelihoods of Interest
Model
A1
A2
A3
fitted
R
Log(likelihood)
161.851147
163.173943
161.851147
160.552452
153.631041
# Param's
5
8
5
4
2
AIC
-313.702293
-310.347886
-313.702293
-313.104903
-303.262082
Explanation of Tests
Test 1: Do responses and/or variances differ among Dose levels?
(A2 vs. R)
Test 2: Are Variances Homogeneous? (A1 vs A2)
Test 3: Are variances adequately modeled? (A2 vs. A3)
Test 4: Does the Model for the Mean Fit? (A3 vs. fitted)
(Note: When rho=0 the results of Test 3 and Test 2 will be the same.)
Test
Test 1
Test 2
Test 3
Test 4
Tests of Interest
-2*log(Likelihood Ratio) Test df
19.0858
2.64559
2.64559
2.59739
p-value
0. 004021
0.4496
0.4496
0. 107
The p-value for Test 1 is less than .05. There appears to be a
difference between response and/or variances among the dose levels
It seems appropriate to model the data
The p-value for Test 2 is greater than .1. A homogeneous variance
model appears to be appropriate here
The p-value for Test 3 is greater than .1.
to be appropriate here
The p-value for Test 4 is greater than .1.
to adequately describe the data
The modeled variance appears
The model chosen seems
Benchmark Dose Computation
Specified effect = 1
Risk Type = Estimated standard deviations from the control mean
Confidence level = 0.95
BMD = 3.51354
BMDL = 1.89704
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Decreased Absolute Epididymis Weight in Rats Exposed to NaCN in Drinking Water for 13
Weeks (NTP, J993)
Table B-3. Decreased epididymis weight in F344 rats following administration of
NaCN in drinking water for 13 weeks
Male rats
0 ppm
30 ppm
100 ppm
300 ppm
Dose (mg/kg-day)
0
1.4
4.5
12.5
Weight (g) ± SI)
Epididymis, absolute
0.448 ±0.019
0.437 ±0.016
0.425 ± 022a
0.417 ±0.016b
aNot reported as significant in NTP (1993) but significant by Dunnett's test in independent
analyses conducted for this assessment, p < 0.05.
Significant by Shirley's test, p £ 0.01,
Source: NTP (1993).
All models for continuous variables available in the EPA BMDS version 1.4.1c, except
the Hill model, were fit to the data in the Table B-3. The Hill model was not fit to these data
because fitting of the Hill model requires the estimation of four parameters (i.e., intercept, v, n,
and k), which necessitates having a minimum of five dose groups in order to have adequate
degrees of freedom for testing model fit. The NTP (1993) study has only four dose groups, and
thus the Hill model could not be fit to these data. All models fit were constant variance models.
All models tested provided adequate fit to the data, based on the summary results reported by the
BMDS output and visual examination of the graphs. A summary of the goodness-of-fit statistics
for the tested models and resulting BMD and BMDL is presented in Table B-4.
Table B-4. BMD modeling results for decreased epididymis weight in rats
Study
Endpoint
Model
AIC
Goodness-of-
fit p value
BMD
BMDL
NTP (1993);
male rats
Epididymis weight
(absolute)
Linear (1° polynomial)
-274.73
0.22
8.2
5.6
Polynomial (2°)
-275.64
0.73
3.2
1.8
Power
-274.73
0.22
8.2
5.6
All models adequately described the data as demonstrated by goodness-of-fitp values >
0.1. Of the available models, the polynomial (2°) was selected, based on having the lowest AIC
value. Visual inspection reveals that the model describes the data well.
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Polynomial Model with 0.95 Confidence Level
Polynomial
0.46
0.45
O)
w
§ 0.44
Q_
CO
CD
C 0.43
TO
0)
0.42
0.41
BMD
,BMD,
0.4
0
2
4
6
8
10
12
dose
15:40 02/13 2008
Figure B-2. Observed and predicted decrease in epididymis weight in F344
rats following administration of NaCN in drinking water for 13 weeks.
The computer output for the polynomial model of decreased (absolute) epididymis weight
follows:
Polynomial Model. (Version: 2.12; Date: 02/20/2007)
Input Data File: C:\BMDSl-4-lC\EPIDIDYMIS_WEIGHT_ABSOLUTE.(d)
Gnuplot Plotting File: C:\BMDSl-4-lC\EPIDIDYMIS_WEIGHT_ABSOLUTE.plt
Wed Feb 13 15:40:43 2008
BMDS MODEL RUN
The form of the response function is:
Y[dose] = beta 0 + beta l*dose + beta 2*doseA2 + ...
Dependent variable = MEAN
Independent variable = COLUMN2
rho is set to 0
Signs of the polynomial coefficients are not restricted
A constant variance model is fit
Total number of dose groups = 4
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Default Initial Parameter Values
alpha = 0.00033925
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rho =
beta 0 =
beta 1 =
beta 2 =
0 Specified
0.447054
-0. 00654
0. 000331286
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -rho have been estimated at a boundary point,
or have been specified by the user, and do not appear in the correlation
matrix )
alpha
beta 0
beta 1
beta 2
alpha
1
le-010
-3.9e-011
-1.8e-011
beta 0
le-010
1
-0.72
0. 61
beta 1
-3.9e-011
-0.72
1
-0.91
beta 2
-1.8e-011
0. 61
-0.91
1
Interval
Variable
Limit
alpha
0. 000440482
beta_0
0.456513
beta_l
0. 00153141
beta_2
0. 000704843
Estimate
0.00030626
0.447054
-0.00654
0.000331286
Parameter Estimates
Std. Err.
6. 84818e-005
0.00482605
0.00255545
0. 000190594
95.0% Wald Confidence
Lower Conf. Limit Upper Conf.
0.000172038
0.437595
-0.0115486
-4 .22712e-005
Table of Data and Estimated Values of Interest
Dose
Obs Mean
Est Mean Obs Std Dev Est Std Dev Scaled Res.
0
1.4
4.5
12.5
10
10
10
10
0.448
0.437
0.425
0.417
0.447
0. 439
0. 424
0. 417
0.019
0.016
0.022
0.016
0.0175
0.0175
0.0175
0.0175
0.171
-0.28
0.121
-0.0121
Model Descriptions for likelihoods calculated
Model A1: Yij
Var{e(ij)}
Model A2: Yij
Var{e (ij) }
Model A3: Yij
Var{e (ij) }
Mu(i) + e(i j)
S i gma A 2
Mu(i) + e(i j)
S i gma(i)A 2
Mu(i) + e(i j)
S i gma A 2
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Model A3 uses any fixed variance parameters that were specified by the
user
Model R: Yi = Mu + e(i)
Var{e(i)} = SigmaA2
Likelihoods of Interest
Model
A1
A2
A3
fitted
R
Log(likelihood)
141. 882675
142.610833
141.882675
141.821537
134.393157
# Param's
5
8
5
4
2
AIC
-273.765351
-269.221665
-273.765351
-275.643073
-264.786314
Test 1
Test 2
Test 3
Test 4
Explanation of Tests
Do responses and/or variances differ among Dose levels? (A2 vs. R)
Are Variances Homogeneous? (A1 vs A2)
Are variances adequately modeled? (A2 vs. A3)
Does the Model for the Mean Fit? (A3 vs. fitted)
(Note: When rho=0 the results of Test 3 and Test 2 will be the same.)
Tests of Interest
Test
Test 1
Test 2
Test 3
Test 4
-2*log(Likelihood Ratio) Test df
16.4354
1.45631
1.45631
0. 122278
p-value
0.0116
0.6924
0.6924
0.7266
The p-value for Test 1 is less than .05. There appears to be a
difference between response and/or variances among the dose levels
It seems appropriate to model the data
The p-value for Test 2 is greater than .1. A homogeneous variance
model appears to be appropriate here
The p-value for Test 3 is greater than .1. The modeled variance appears
to be appropriate here
The p-value for Test 4 is greater than .1. The model chosen seems
to adequately describe the data
Benchmark Dose Computation
Specified effect =
Risk Type =
Confidence level =
BMD =
BMDL =
1
Estimated standard deviations from the control mean
0. 95
3. 19201
1.79176
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Decreased Absolute Testis Weight in Rats Exposed toNaCN in Drinking Water for 13 Weeks
(NTP, 1993)
Table B-5. Decreased testis weight in F344 rats, following administration of NaCN
in drinking water for 13 weeks
Male rats
0 ppm
30 ppm
100 ppm
300 ppm
Dose (mg/kg-day)
0
1.4
4.5
12.5
Weight (g) ± SI)
Testis, absolute
1.58 ±0.094
1.56 ±0.063
1.52 ±0.063
1.46 ±0.063
Table B-6. BMD modeling results for decreased testis weight in rats
Study
Endpoint
Model
AIC
Goodness-of-
fit p value
BMD
BMDL
NTP (1993);
male rats
Testis weight
(absolute)
Linear (1° polynomial)
-167.94
0.82
7.4
5.1
Polynomial (2°)
-166.32
0.98
5.3
2.4
Power
-167.94
0.82
7.4
5.1
Linear Model with 0.95 Confidence Level
Linear
1.65
1.55
1.45
BMDI
BMD
8
0
2
4
6
10
12
dose
15:08 02/13 2008
Figure B-3. Observed and predicted decrease in epididymis weight in F344
rats following administration of NaCN in drinking water for 13 weeks.
AUGUST 2009
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The computer output from the linear model of decreased (absolute) testis weight follows:
Polynomial Model. (Version: 2.12; Date: 02/20/2007)
Input Data File: C:\BMDSl-4-lC\TESTIS_WEIGHT_ABSOLUTE.(d)
Gnuplot Plotting File: C:\BMDSl-4-lC\TESTIS_WEIGHT_ABSOLUTE.plt
Wed Feb 13 15:08:00 2008
BMDS MODEL RUN
The form of the response function is:
Y[dose] = beta 0 + beta l*dose + beta 2*doseA2 +
Dependent variable = MEAN
Independent variable = COLUMN2
rho is set to 0
Signs of the polynomial coefficients are not restricted
A constant variance model is fit
Total number of dose groups = 4
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Default Initial Parameter Values
alpha = 0.005233
rho = 0 Specified
beta 0 = 1.57305
beta~l = -0.00935835
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -rho have been estimated at a boundary point,
or have been specified by the user,and do not appear in the correlation
matrix )
alpha beta 0 beta 1
alpha 1 -4.le-010 3.3e-011
beta_0 -4.le-010 1 -0.69
beta 1 3. 3e-011 -0.69 1
Parameter Estimates
95.0% Wald Confidence
Interval
AUGUST 2009 B-10 DRAFT - DO NOT CITE OR QUOTE
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Variable
Limit
alpha
0.00683971
beta_0
1.60252
beta_l
0.00494569
Estimate
0. 00475554
1.57305
-0.00935835
Std. Err. Lower Conf. Limit Upper Conf.
0.00106337 0.00267137
0.0150381 1.54357
0.0022514 -0.013771
Table of Data and Estimated Values of Interest
Dose
Obs Mean
Est Mean Obs Std Dev Est Std Dev Scaled Res.
0
1.4
4.5
12.5
10
10
10
10
1.58
1.56
1.52
1.46
1.57
1.56
1.53
1.46
0.095
0.063
0.063
0.063
0. 069
0. 069
0. 069
0. 069
0.319
0.00244
-0.501
0.18
Model Descriptions for likelihoods calculated
Model A1: Yij = Mu(i) + e(ij)
Var{e(ij)} = SigmaA2
Model A2: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma(i)A2
Model A3: Yij = Mu(i) + e(ij)
Var{e(ij)} = SigmaA2
Model A3 uses any fixed variance parameters that
were specified by the user
Model R: Yi = Mu + e(i)
Var{e(i)} = SigmaA2
Likelihoods of Interest
Model Log(likelihood) # Param's AIC
A1 87.162621 5 -164.325243
A2 88.584611 8 -161.169222
A3 87.162621 5 -164.325243
fitted 86.968887 3 -167.937775
R 79.788144 2 -155.576289
Explanation of Tests
Test 1: Do responses and/or variances differ among Dose levels?
(A2 vs. R)
Test 2: Are Variances Homogeneous? (A1 vs A2)
Test 3: Are variances adequately modeled? (A2 vs. A3)
Test 4: Does the Model for the Mean Fit? (A3 vs. fitted)
(Note: When rho=0 the results of Test 3 and Test 2 will be the same.)
Tests of Interest
AUGUST 2009
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Test -2*log(Likelihood Ratio) Test df
p-value
Test 1
Test 2
Test 3
Test 4
0.387468
2 . 84398
2 . 84398
17 . 5929
6
3
3
2
0. 007334
0.4163
0.4163
0. 8239
The p-value for Test 1 is less than .05. There appears to be a
difference between response and/or variances among the dose levels
It seems appropriate to model the data
The p-value for Test 2 is greater than .1. A homogeneous variance
model appears to be appropriate here
The p-value for Test 3 is greater than .1. The modeled variance appears
to be appropriate here
The p-value for Test 4 is greater than .1. The model chosen seems
to adequately describe the data
Benchmark Dose Computation
Specified effect
1
Risk Type
Estimated standard deviations from the control mean
Confidence level
0. 95
BMD
7 . 36887
BMDL
5. 12669
AUGUST 2009
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Decreased Testicular Spermatid Concentration in Rats Exposed to NaCN in Drinking Water for
13 Weeks (NTP, J993)
Table B-7. Decreased testicular spermatid concentration in F344 rats following
administration of NaCN in drinking water for 13 weeks
Male rats
0 ppm
30 ppm
100 ppm
300 ppm
Dose (mg/kg-day)
0
1.4
4.5
12.5
Mean/1 Or4 mL suspension ± SI)
Spermatid count
89.28 ±9.64
84.68 ± 12.74
82.90 ±9.99
77.10 ±6.96
Table B-8. BMD modeling results for decreased testicular spermatid concentration
in rats
Study
Endpoint
Model
AIC
Goodness-of-
fit p value
BMD
BMDL
NTP (1993);
male rats
Spermatid
concentration
(testis)
Linear (1° polynomial)
227.04
0.70
11.2
6.9
Polynomial (2°)
228.73
0.53
8.5
2.9
Power
227.04
0.70
11.2
6.9
O)
w
c
o
Q.
0 2 4 6 8 10 12
dose
16:40 02/21 2008
Figure B-4. Observed and predicted decrease in testicular spermatid
concentration in F344 rats following administration of NaCN in drinking
water for 13 weeks.
Linear Model with 0.95 Confidence Level
Linear
95
90
85
80
75
BMDL
BMD
70
AUGUST 2009
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The computer output from the polynomial model of testicular spermatid concentration follows:
Polynomial Model. (Version: 2.12; Date: 02/20/2007)
Input Data File: C:\BMDSl-4-lC\UNSAVEDl.(d)
Gnuplot Plotting File: C:\BMDSl-4-lC\UNSAVEDl.plt
Thu Feb 21 16:40:27 2008
BMDS MODEL RUN
The form of the response function is:
Y[dose] = beta 0 + beta l*dose + beta 2*doseA2 + . . .
Dependent variable = MEAN
Independent variable = COLUMN1
rho is set to 0
Signs of the polynomial coefficients are not restricted
A constant variance model is fit
Total number of dose groups = 4
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Default Initial Parameter Values
alpha = 100.87
rho = 0 Specified
beta_0 = 87.4548
beta 1 = -0.861906
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -rho
have been estimated at a boundary point, or have been
specified by the user,
and do not appear in the correlation matrix )
alpha beta 0 beta 1
alpha 1 l.le-010 -2.5e-013
beta_0 l.le-010 1 -0.69
beta 1 -2.5e-013 -0.69 1
Parameter Estimates
95.0% Wald Confidence
Interval
AUGUST 2009
B-14
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Variable
Limit
132.879
91.563
alpha
beta 0
beta 1
Estimate
92.3886
87. 4548
-0.861906
Std. Err. Lower Conf. Limit Upper Conf.
20.6587 51.8982
2.09605 83.3466
0.313806 -1.47695
0.246857
Table of Data and Estimated Values of Interest
Dose
Obs Mean
Est Mean Obs Std Dev Est Std Dev Scaled Res.
0
1.4
4.5
12.5
10
10
10
10
89.3
84.7
82.9
77.1
87.5
86.2
83.6
76.7
9. 64
12 .7
9. 99
6. 96
9. 61
9. 61
9. 61
9. 61
0.6
-0.516
-0.222
0.138
Model Descriptions for likelihoods calculated
Model A1: Yij = Mu(i) + e(ij)
Var{e(ij)} = SigmaA2
Model A2: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma(i)A2
Model A3: Yij = Mu(i) + e(ij)
Var{e(ij)} = SigmaA2
Model A3 uses any fixed variance parameters that
were specified by the user
Model R: Yi = Mu + e(i)
Var{e(i)} = SigmaA2
Likelihoods of Interest
Model Log(likelihood) # Param's AIC
A1 -110.169386 5 230.338773
A2 -108.417108 8 232.834216
A3 -110.169386 5 230.338773
fitted -110.520064 3 227.040129
R -113.975552 2 231.951103
Explanation of Tests
Test 1: Do responses and/or variances differ among Dose levels?
(A2 vs. R)
Test 2: Are Variances Homogeneous? (A1 vs A2)
Test 3: Are variances adequately modeled? (A2 vs. A3)
Test 4: Does the Model for the Mean Fit? (A3 vs. fitted)
(Note: When rho=0 the results of Test 3 and Test 2 will be the same.)
Tests of Interest
AUGUST 2009
B-15
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Test -2*log(Likelihood Ratio) Test df
p-value
Test 1
Test 2
Test 3
Test 4
11.1169
3.50456
3.50456
0.701356
6
3
3
2
0.08483
0.3202
0.3202
0.7042
The p-value for Test 1 is greater than .05. There may not be a
diffence between responses and/or variances among the dose levels
Modelling the data with a dose/response curve may not be appropriate
The p-value for Test 2 is greater than .1. A homogeneous variance
model appears to be appropriate here
The p-value for Test 3 is greater than .1. The modeled variance appears
to be appropriate here
The p-value for Test 4 is greater than .1. The model chosen seems
to adequately describe the data
Benchmark Dose Computation
Specified effect
1
Risk Type
Estimated standard deviations from the control mean
Confidence level
0. 95
BMD
11.1519
BMDL
6.85
AUGUST 2009
B-16
DRAFT - DO NOT CITE OR QUOTE
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