Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD Final draft XIII 1 Chapter VIII. Quantification of Toxicological Effects The quantification of toxicological effects of a chemical consists of separate assessments of noncarcinogenic and carcinogenic health effects. Unless otherwise specified, chemicals which do not produce carcinogenic effects are believed to have a threshold dose below which no adverse, noncarcinogenic health effects occur, while carcinogens are assumed to act without a threshold.

A. Introduction to Methods A. 1 Quantification of Noncarcinogenic Effects In quantification of noncarcinogenic effects, a Reference Dose ( RfD) ( formerly called the Acceptable Daily Intake ( ADI)) is calculated. The RfD is " an estimate ( with uncertainty spanning approximately an order of magnitude) of a daily exposure to the human population ( including sensitive subgroups) that is likely to be without appreciable risk of deleterious effects over a lifetime" ( U. S. EPA, 1993). The RfD is derived from a no observed adverse effect level ( NOAEL), lowest observed adverse effect level ( LOAEL), or a NOAEL surrogate such as a benchmark dose identified from a subchronic or chronic study, and divided by a composite uncertainty factor( s). The RfD is calculated as follows:

RfD = NOAEL ( LOAEL) UF × MF Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD Final draft XIII 2 where:

NOAEL = No observed adverse effect level from a high quality toxicological study of an appropriate duration LOAEL = Lowest observed adverse effect level from a high quality toxicological study of an appropriate duration. In situations where there is no NOAEL for a contaminant but there is a LOAEL, the LOAEL can be used for the RfD calculation with the inclusion of an additional uncertainty factor.

UF = Uncertainty factor chosen according to EPA/ NAS guidelines MF = Modifying factor Selection of the uncertainty factor to be employed in calculation of the RfD is based on professional judgment, while considering the entire database of toxicological effects for the chemical. To ensure that uncertainty factors are selected and applied in a consistent manner, the Office of Water ( OW) employs a modification to the guidelines proposed by the National Academy of Sciences ( NAS, 1977, 1980). According to the EPA approach ( U. S. EPA, 1993),

uncertainty is broken down into its components, and each dimension of uncertainty is given a quantitative rating. The total uncertainty factor is the product of the component uncertainties.

The individual components of the uncertainty are as follows:

UF H A factor of 1, 3, or 10 is used when extrapolating from valid data in studies using long term exposure to average healthy humans. This factor is intended to account for the variation in sensitivity ( intraspecies variation) among the members of the human population.

UF A An additional factor of 1, 3, or 10 is used when extrapolating from valid results of long term studies on experimental animals when results of studies of human exposure are not available or are inadequate. This factor is intended to account for the uncertainty involved in extrapolating from animal data to humans ( interspecies variation). Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD Final draft XIII 3 UF S An additional factor of 1, 3, or 10 is used when extrapolating from lessthan chronic results on experimental animals when there are no useful longterm human data. This factor is intended to account for the uncertainty involved in extrapolating from less than chronic NOAELs to chronic NOAELs.

UF L An additional factor of 1, 3, or 10 is used when deriving an RfD from a LOAEL, instead of a NOAEL. This factor is intended to account for the uncertainty involved in extrapolating from LOAELs to NOAELs.

UF D An additional factor of 1, 3 or 10 is used when deriving an RfD from an " incomplete" database. This factor is meant to account for the inability of any single type of study to consider all toxic endpoints. The intermediate factor of 3 ( approximately ½ log 10 unit, i. e., the square root of 10) is often used when there is a single data gap exclusive of chronic data. It is often designated as UF D.

On occasion, EPA also uses a modifying factor in the determination of the RfD. A modifying factor is an additional uncertainty factor that is greater than zero and less than or equal to 10. The magnitude of the MF depends upon the professional assessment of scientific uncertainties of the study and database not explicitly treated above ( e. g., the number of species tested). The default value for the MF is 1.

In establishing the UF or MF, it is recognized that professional scientific judgment must be used. The total product of the uncertainty factors and modifying factor should not exceed 3000.

If the assignment of uncertainty results in a UF/ MF product that exceeds 3000, then the database does not support development of an RfD. The quantification of toxicological effects of a chemical consists of separate assessments of noncarcinogenic and carcinogenic health effects.

Unless otherwise specified, chemicals which do not produce carcinogenic effects are believed to have a threshold dose below which no adverse, noncarcinogenic health effects occur, while carcinogens are assumed to act without a threshold. Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD Final draft XIII 4 A. 1.1. Drinking Water Equivalent Level The drinking water equivalent ( DWEL) is calculated from the RfD. The DWEL represents a drinking water specific lifetime exposure at which adverse, noncarcinogenic health effects are not anticipated to occur. The DWEL assumes 100% exposure from drinking water.

The DWEL provides the noncarcinogenic health effects basis for establishing a drinking water standard. For ingestion data, the DWEL is derived as follows:

DWEL = ( RfD) × BW WI where:

BW = 70 kg adult body weight WI = Drinking water intake ( 2 L/ day)

A. 1.2. Health Advisory Values In addition to the RfD and the DWEL, EPA calculates Health Advisory ( HA) values for noncancer effects. HAs are determined for lifetime exposures as well as for exposures of shorter duration ( 1 day, 10 day, and longer term). The shorter duration HA values are used as informal guidance to municipalities and other organizations when emergency spills or contamination situations occur. The lifetime HA becomes the MCLG for a chemical that is not a carcinogen. Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD Final draft XIII 5 The shorter term HAs are calculated using an equation similar to the ones for RfD and DWEL; however, the NOAELs or LOAELs are derived from acute or subchronic studies and identify a sensitive noncarcinogenic endpoint of toxicity. The HAs are derived as follows:

HA = NOAEL or LOAEL × BW UF × WI where:

NOAEL or LOAEL = No or lowest observed adverse effect level in mg/ kg bw/ day BW = Assumed body weight of a child ( 10 kg) or an adult ( 70 kg)

UF = Uncertainty factor, in accordance with EPA or NAS/ OW guidelines WI = Assumed daily water consumption of a child ( 1 L/ day) or an adult ( 2 L/ day)

Using the above equation, the following drinking water HAs are developed for noncarcinogenic effects:

° 1 day HA for a 10 kg child ingesting 1 L water per day.

° 10 day HA for a 10 kg child ingesting 1 L water per day.

° Longer term HA for a 10 kg child ingesting 1 L water per day.

° Longer term HA for a 70 kg adult ingesting 2 L water per day.

Each of these shorter term HA values assumes that the total exposure to the contaminant comes from drinking water. Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD Final draft XIII 6 The lifetime HA is calculated from the DWEL and takes into account exposure from sources other than drinking water. It is calculated using the following equation:

Lifetime HA = DWEL × RSC where:

DWEL = Drinking water equivalent level RSC = Relative source contribution. The fraction of the total exposure allocated to drinking water.

A. 2 Quantification of Carcinogenic Effects Under the 1986 guidelines, the EPA categorizes the carcinogenic potential of a chemical based on the overall weight of evidence according to the following scheme:

° Group A: Human Carcinogen. Sufficient evidence exists from epidemiology studies to support a causal association between exposure to the chemical and human cancer.

° Group B: Probable Human Carcinogen. Sufficient evidence of carcinogenicity in animals with limited ( Group B1) or inadequate ( Group B2) evidence in humans.

° Group C: Possible Human Carcinogen. Limited evidence of carcinogenicity in animals in the absence of human data. Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD Final draft XIII 7 ° Group D: Not classified as to Human Carcinogenicity. Inadequate human and animal evidence of carcinogenicity or for which no data are available.

° Group E: Evidence of Noncarcinogenicity for Humans. No evidence of carcinogenicity in at least two adequate animal tests in different species or in both adequate epidemiologic and animal studies.

If toxicological evidence leads to the classification of the contaminant as a genotoxic,

probable or possible human carcinogen, mathematical models are used to calculate the estimated excess cancer risk associated with ingestion of the contaminant in drinking water. The data used in these estimates usually come from lifetime exposure studies in animals. In order to predict the risk for humans from animal data, animal doses must be converted to equivalent human doses.

This conversion includes correction for noncontinuous exposure, less than lifetime studies and differences in size. It is assumed that the average adult human body weight is 70 kg and that the average water consumption of an adult human is two liters of water per day.

For contaminants with a carcinogenic potential, chemical levels are correlated with a carcinogenic risk estimate by employing a cancer potency ( unit risk) value together with the assumption for lifetime exposure via ingestion of water. Under the 1986 Carcinogen Risk Assessment Guidelines, the cancer unit risk is usually derived from a linearized multistage model with a 95% upper confidence limit providing a low dose estimate; that is, the true risk to humans,

while not identifiable, is not likely to exceed the upper limit estimate and, in fact, may be lower.

Excess cancer risk estimates may also be calculated using other models such as the one hit, Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 8 Final draft Weibull, logit and probit models. There is little basis in the current understanding of the biological mechanisms involved in cancer to suggest that any one of these models is able to predict risk more accurately than any of the others. Because each model is based upon differing assumptions, the estimates that are derived for each model can differ by several orders of magnitude.

The scientific data base used to calculate and support the setting of cancer risk rates has an inherent uncertainty due to the systematic and random errors in scientific measurement. In most cases, only studies using experimental animals have been performed. Thus, there is uncertainty when the data are extrapolated to humans. When developing cancer risk rates, several other areas of uncertainty exist, such as the incomplete knowledge concerning the health effects of contaminants in drinking water, the impact of the experimental animal's age, sex and species, the nature of the target organ system( s) examined and the actual rate of exposure of the internal targets in experimental animals or humans. Dose response data usually are available only for high levels of exposure, not for the lower levels of exposure at which a standard may be set. When there is exposure to more than one contaminant, additional uncertainty results from a lack of information about possible synergistic or antagonistic effects.

The quantification of toxicological effects of a chemical consists of separate assessments of noncarcinogenic and carcinogenic health effects. Chemicals that do not produce carcinogenic effects are believed to have a threshold dose below which no adverse, noncarcinogenic health effects occur, while carcinogens are assumed to act without a threshold. Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 9 Final draft B. Noncarcinogenic Effects The useful toxicity data for the cyanogen chloride and its known metabolites ( cyanide,

thiocyanate) are provided in Tables VIII 1 through VIII 4. ( The useful toxicity data for the putative metabolites cyanate and cyanamide, and for HCl, which is not appropriate as a surrogate,

are summarized in Appendix Tables C 1, D 1, and E 1, respectively.) As noted in the Executive Summary, because of the lack of adequate toxicity for cyanogen chloride, use of data from known or potential cyanogen chloride metabolites was considered as the basis for the cyanogen chloride Health Advisories. Therefore, this section first provides the derivation of Health Advisories for cyanide and thiocyanate, the known metabolites of cyanogen chloride. The Health Advisories for cyanogen chloride are then derived, based on consideration of the Health Advisories derived for the other chemicals, as well as consideration of the limited toxicity and toxicokinetics data available for cyanogen chloride.

B. 1. Cyanide B. 1.1 One day Health Advisory for Cyanide No suitable studies were located. Only human poisoning studies, in which severe effects were observed ( e. g., Liebowitz and Schwarz, 1948; Saincher et al., 1994), and animal lethality studies ( e. g., Ferguson, 1962; Smyth et al., 1969), were identified. In the absence of adequate data, the ten day HA value is recommended as a conservative estimate of an appropriate one day HA value. Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 10 Final draft B. 1.2. Ten day Health Advisory for Cyanide Two suitable studies of appropriate duration were located. The first was that of Kreutler et al. ( 1978). In this study, rats were exposed in the diet for 2 weeks to 87 mg CN/ kg day, and were provided diets that either supplied adequate protein or that were deficient in protein and iodine. No effect was seen in the group receiving adequate protein and iodine, but decreased plasma TSH and increased thyroid weight were observed in the group provided the diet deficient in protein and iodine. Although changes in TSH and thyroid weight alone can be considered adaptive rather than adverse effects, this study did not include assessment of thyroid histopathology to determine if thyroid hyperplasia was present. Thus, one can consider this study as identifying a minimal LOAEL of 40 mg CN/ kg day in a sensitive population. Benchmark dose analysis of TSH levels resulted in a BMDL of 2.1 mg/ kg day. Confidence in this BMDL is limited, however, because only a single dose level and a control were available, meaning that only a linear model could be fit to the data. In addition to the absence of information on the shape of the dose response curve, the BMDL is markedly lower than the NOAEL, meaning that calculating the BMDL involved significant extrapolation below the data, a practice generally not needed for BMD modeling. Thus, although this BMDL can be considered a NOAEL surrogate because the modeled effect is an adaptive effect, overall confidence in the BMDL is limited.

In the second study ( Palmer and Olson, 1979), no effect was seen in rats administered 8 mg CN/ kg day in feed for 21 days. Analysis of the feed by the study authors led them to conclude that < 20% of the administered dose of cyanide was bioavailable; therefore, the NOAEL can be estimated as 1.6 mg CN/ kg day. In a parallel study by the same authors, a LOAEL of 14 mg Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 11 Final draft CN/ kg day was identified in rats administered cyanide in drinking water for the same duration,

based on increased liver weight; no drinking water NOAEL corresponding to this LOAEL was identified. BMDL modeling could not be conducted for this study, because no measure of variability was provided.

It is of interest that the doses in the two short term studies are higher than the doses reported to cause death in rats and mice when delivered in a single dose ( 4 8 mg/ kg, Ferguson,

1962; Smyth et al., 1969). However, this finding is consistent with the mechanism of action of cyanide. A bolus dose results in rapid absorption of cyanide, which can overwhelm the ability of the liver to metabolize cyanide to thiocyanate by first pass metabolism. By contrast,

administration of the same dose over the course of the day in feed or water allows the liver to fully metabolize the cyanide.

In light of the uncertainty regarding the actual dose administered in feed by Palmer and Olson ( 1979), and because drinking water is the exposure route of interest, the drinking water LOAEL of 14 mg CN/ kg day ( Palmer and Olson, 1979) is used as the basis for the Ten day Health Advisory. This LOAEL is supported by the BMDL of 2.1 mg/ kg day calculated based on increased plasma TSH in protein and iodine deficient rats ( a sensitive population) in the Kreutler et al. ( 1978) study. Note that the drinking water LOAEL in Palmer and Olson ( 1979) is approximately a factor of 9 times the NOAEL in feed in this study. This means that the Health Advisory calculated using the drinking water study ( and an uncertainty factor of 10 for extrapolating from a LOAEL) is similar to the value that would be calculated using the NOAEL in Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 12 Final draft feed, but the latter value would have greater uncertainty, due to the uncertainty in the ingested dose. An uncertainty factor of 10 is used for animal to human extrapolation, an uncertainty factor of 10 is used to protect sensitive human populations ( i. e., to account for human variability), and an uncertainty factor of 10 is used to account for extrapolation from a LOAEL.

A similar Health Advisory could be calculated from the BMDL in the Kreutler et al.

( 1978) study. In this case, no factor would be needed for extrapolation from a LOAEL, and the default factor of 10 would be used to extrapolate from animals to humans. Since an effect was seen only in the rats fed a protein deficient diet, the study was done in a sensitive population, and no extrapolation would be needed to consider toxicodynamic variability relating to a sensitive subpopulation. Therefore the uncertainty factor for human variability would be reduced to a factor of 3, for consideration of toxicokinetic variability. However, in light of the large degree of extrapolation involved and the uncertainty related to the BMDL being based on only two data points, confidence in the BMDL is limited. Therefore, it was not considered appropriate to consider the Kreutler et al. ( 1978) study as a co critical study. As shown in Table VIII 5:

( 14 mg/ kg day) ( 10 kg) Ten day HA ( for a child) = = 0.14 mg/ L, rounded to 0.1 mg/ L ( 1000) ( 1 L/ day) where:

14 mg/ kg day = LOAEL, based on increased liver weight in rats receiving 14 mg CN/ kgday for 21 days in drinking water ( Palmer and Olson, 1979)

10 kg = assumed body weight of a child.

1000 = composite uncertainty factor, chosen to account for extrapolation from animals, protection of sensitive subpopulations, and extrapolation from a LOAEL Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 13 Final draft 1 L/ day = assumed daily water consumption by a 10 kg child.

B. 1.3. Longer term Health Advisory for Cyanide The most appropriate study on which to base the longer term HA is the subchronic bioassay by NTP ( 1993), which evaluated the effects of several doses on all organ systems in both mice and rats. The study identified a NOAEL of 4.5 mg/ kg day and a LOAEL of 12.5 mg/ kg day based on male reproductive effects ( decreased epididymal weight, testis weight, and sperm count)

in rats. Although decreases in left cauda epididymis weight and sperm motility were observed at doses of 1.4 and 4.5 mg CN/ kg day, these changes were insufficient for these doses to be considered LOAELs. NTP ( 1993) did not consider the effects on sperm motility to be adverse because they were small and well within the range of historical controls. Similarly, the decrease in left cauda epididymis weight was small at 1.4 and 4.5 mg/ kg day, and there was little evidence of a dose response at these doses. A BMDL of 0.79 mg/ kg day was calculated for this study, based on decreased epididymal weight, and supported by a BMDL of 1.3 mg/ kg day for both decreased spermatid heads/ testis and decreased spermatid count. No studies on the effects of cyanide on reproduction function in males have been located, although NTP ( 1993) stated that comparison with other studies indicates that the observed decreases would not result in functional effects.

Two other subchronic studies identified lower NOAELs and LOAELs; however, these studies are not as appropriate. Jackson ( 1988) identified a NOAEL of 0.7 and a LOAEL of 1.2 mg/ kg day in miniature pigs based on altered thyroid hormones and behavioral changes. This NOAEL is not considered reliable, however, because the study was not controlled sufficiently well to ensure that the observed differences were not due to variability, particularly since unequal numbers of males and females were used in each dose group and the observed endpoints tended to differ between Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 14 Final draft the two sexes. Kamalu ( 1993) identified a LOAEL of 1.04 mg/ kg day in dogs based on male reproductive effects and histopathological changes in the kidney and adrenal gland. This appears to be a well conducted study, but dogs are not a suitable model for the toxicity of cyanide in humans, because their levels of rhodanese ( the enzyme which detoxifies cyanide) are much lower than the levels in humans ( ATSDR, 1997).

The database for cyanide includes a subchronic systemic toxicity study in rats and mice ( NTP, 1993). No standard multigeneration or developmental toxicity studies of cyanide are available. The single available developmental toxicity ( Tewe and Maner, 1981) found no effects on litter size, birth weight of pups, or pup mortality in the offspring of female rats receiving cyanide in the diet throughout mating, gestation, and lactation, although decreased pup body weight gain compared to the low cyanide control was seen in the high cyanide dose ( weanling dose) of 34.3 mg/ kg day; morphological assessment of developmental effects was not conducted.

The absence of a study evaluating reproductive function is of particular concern, since the critical effect is histopathological changes to the male reproductive organs ( NTP, 1993). In addition, the absence of a developmental toxicity study that evaluated thyroid hormone or neurodevelopmental endpoints is of concern, since prolonged exposure to cyanide is known to affect thyroid hormone levels ( with a LOAEL of 44 mg/ kg day and no NOAEL identified in Philbrick et al., 1979), and neurodevelopmental effects are known to result from decreased T4 levels ( Chan and Kilby,

2000). As discussed in Chapter 7, there is not a direct correlation between serum levels of thyroid hormones and neurodevelopmental effects. Therefore, the Philbrick et al. ( 1979) data are inadequate to determine what doses may cause developmental neurotoxicity. The absence of data on doses that affect male reproductive function, and particularly the absence of a two generation Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 15 Final draft study, warrants at least a factor of 3 to account for database deficiencies. The absence of any studies on neurobehavioral or neurodevelopmental toxicity, secondary to decreased T4 levels warrants an additional factor of 3 for database deficiencies. Therefore, a total database uncertainty factor of 10 is used. No quantitative data are available on toxicokinetic and toxicodynamic differences between animals and humans, or on human variability in these areas.

Therefore, the default uncertainty factors of 10 each are used for extrapolation from animals to humans, and for protection of sensitive individuals. A composite uncertainty factor of 1000 results, based on default factors of 10 each for interspecies extrapolation and human variability,

and a factor of 10 for database uncertainties. The same composite uncertainty factor of 1000 was used for the derivation of a Health Advisory based on the BMDL. No additional uncertainty factor is needed for extrapolation from the BMDL, in light of the minimal adversity of the observed effects, and since the BMDL is lower than the corresponding NOAEL.

Derivation of the Longer term HA based on the study NOAEL ( 4.5 mg/ kg day) ( 10 kg) Longer Term HA ( for a child) = = 0.045 mg/ L ( rounded to 0.05 mg/ L) ( 1000) ( 1 L/ day)

where:

4.5 mg/ kg day = NOAEL, based on absence of effects on male reproductive organ weight and spermatid counts in rats exposed to cyanide in drinking water for 13 weeks, with a corresponding LOAEL of 12.5 mg/ kg day ( NTP, 1993).

10 kg = assumed body weight of a child.

1000 = composite uncertainty factor, chosen to account for extrapolation from a NOAEL in animals, protection of sensitive individuals, and insufficiencies in the database. The database insufficiencies include the absence of a 2 generation reproduction study, and the absence of adequate developmental toxicity studies, particularly a neurodevelopmental toxicity study, in light of the potential for low level cyanide exposure to decrease T4 levels. Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 16 Final draft 1 L/ day = assumed daily water consumption by a 10 kg child.

The Longer term HA for a 70 kg adult consuming 2 L/ day of water is calculated as follows:

( 4.5 mg/ kg day) ( 70 kg) Longer Term HA ( for adults) = = 0.158 mg/ L ( rounded to 0.2 mg/ L) ( 1000)( 2 L/ day)

where:

4.5 mg/ kg day = NOAEL, based on absence of effects on male reproductive organ weight and spermatid counts in rats exposed to cyanide in drinking water for 13 weeks, with a corresponding LOAEL of 12.5 mg/ kg day ( NTP, 1993).

70 kg = assumed body weight of an adult.

1000 = composite uncertainty factor, chosen to account for extrapolation from a NOAEL in animals, inter individual variability in humans, and insufficiencies in the database. The database insufficiencies include the absence of a 2 generation reproduction study, and the absence of adequate developmental toxicity studies, particularly a neurodevelopmental toxicity study, in light of the potential for low level cyanide exposure to decrease T4 levels.

2 L/ day = assumed daily water consumption by a 70 kg adult.

Derivation of the Longer term HA based on the study BMDL ( 0.79 mg/ kg day) ( 10 kg) Longer Term HA ( child) = = 0.0079 mg/ L ( rounded to 0.008 mg/ L) ( 1000) ( 1 L/ day)

where:

0.79 mg/ kg day = BMDL, based on decreased left epididymis weight in male rats exposed to cyanide in drinking water for 13 weeks ( NTP, 1993).

10 kg = assumed body weight of a child.

1000 = composite uncertainty factor, chosen to account for extrapolation from a BMDL in animals, protection of sensitive individuals, and insufficiencies in the database. The database insufficiencies include the absence of a 2 generation reproduction study, and the absence of adequate developmental toxicity studies, particularly a neurodevelopmental toxicity study, in light of the potential for low level cyanide exposure to decrease T4 levels. Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites 1An uncertainty factor of 10 each was used for extrapolation from animals to humans, and for protection of sensitive human populations. A modifying factor of 5 was used to account for the apparent tolerance to cyanide when it is ingested with food rather than when it is administered by gavage or by EPA/ OW/ OST/ HECD VIII 17 Final draft 1 L/ day = assumed daily water consumption by a 10 kg child.

The Longer term HA for a 70 kg adult consuming 2 L/ day of water is calculated as follows:

( 0.79 mg/ kg day) ( 70 kg) Longer Term HA ( for adults) = = 0.028 mg/ L ( rounded to 0.03 mg/ L) ( 1000)( 2 L/ day)

where:

0.79 mg/ kg day = BMDL, based on decreased left epididymis weight in male rats exposed to cyanide in drinking water for 13 weeks ( NTP, 1993).

70 kg = assumed body weight of an adult.

1000 = composite uncertainty factor, chosen to account for extrapolation from a NOAEL in animals, inter individual variability in humans, and insufficiencies in the database. The database insufficiencies include the absence of a 2 generation reproduction study, and the absence of adequate developmental toxicity studies, particularly a neurodevelopmental toxicity study, in light of the potential for low level cyanide exposure to decrease T4 levels.

2 L/ day = assumed daily water consumption by a 70 kg adult.

B. 1.4 Reference Dose, Drinking Water Equivalent Level, and Lifetime Health Advisory for Cyanide The only chronic oral study of cyanide ( Howard and Hanzal, 1955) identified a freestanding NOAEL of 10.8 mg/ kg day. This study, together with the dietary study of Philbrick et al. ( 1979), which identified a LOAEL of 44 mg/ kg day based on myelin degeneration and increased thyroid weight, were used by EPA in 1985 to develop an RfD for free cyanide of 0.02 mg/ kg day, and an RfD for cyanogen chloride of 0.05 mg/ kg day ( as reported in U. S. EPA,

2001a). 1 However, the more recent subchronic bioassay by NTP ( 1993) is a better quality study Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites drinking water. These RfDs were developed prior to EPA's adoption of the database uncertainty factor. In addition, a default food factor of 0.05 was used to calculate the dose in the Philbrick et al. ( 1979) study, leading to a LOAEL of 30 mg/ kg day.

EPA/ OW/ OST/ HECD VIII 18 Final draft with less uncertainty, because it identifies both a NOAEL and a LOAEL, and, therefore, is more appropriate as the basis of the RfD. ATSDR ( 1997) used the NTP ( 1993) study as the basis of an intermediate oral minimal risk level ( MRL) for cyanide. ATSDR determined that the high dose in male rats, 12.5 mg CN/ kg day, was the LOAEL based on the spectrum of reproductive effects observed at this dose. The next lowest dose, 4.5 mg CN/ kg day, was designated as the NOAEL.

Although decreases in left cauda epididymis weight and sperm motility were observed at doses of 1.4 and 4.5 mg CN/ kg day, NTP ( 1993) ( and ATSDR) determined that these effects by themselves were not adverse. A total uncertainty factor of 100 was applied, a factor of 10 to extrapolate from animals to humans, and another factor of 10 to account for human variability.

The resulting value for the intermediate MRL is 0.05 mg/ kg day.

As discussed in the context of the Longer term Health Advisory, NTP ( 1993) identified a NOAEL of 4.5 mg/ kg day and a LOAEL of 12.5 mg/ kg day based on male reproductive effects ( decreased epididymal weight, testis weight, and sperm count). The effects on sperm motility at lower doses were not considered adverse because the changes were small and well within the range of historical controls. Similarly, the decrease in left cauda epididymis weight was small at 1.4 and 4.5 mg/ kg day, and there was little evidence of a dose response at these doses. A BMDL of 0.79 mg/ kg day was calculated for this study, based on decreased left epididymis weight, and supported by a BMDL of 1.3 mg/ kg day for both decreased spermatid heads/ testis and decreased spermatid count. No studies on the effects of cyanide on reproductive function in males have Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 19 Final draft been located, although NTP ( 1993) stated that comparison with other studies indicates that the observed decreases would not result in functional effects. For development of the Lifetime Health Advisory, full uncertainty factors of 10 were used to account for extrapolation from animals and to account for human variability. The database for cyanide includes systemic toxicity studies in rats and mice, as well as lower quality studies in pigs and dogs. However, no standard multigeneration or developmental toxicity studies of cyanide are available. In the only available study on developmental effects of cyanide ( Tewe and Maner, 1981) the only observed effect was that pup body weight gain was decreased at the high dose of 34.3 mg/ kg day ( weanling dose)

compared to the low " control" cyanide dose; morphological assessment of developmental effects was not conducted. As noted above, the absence of a study evaluating reproductive function and of a developmental toxicity study ( particularly the absence of a neurodevelopmental study) are of particular concern, since the current critical effect is histopathological changes to the male reproductive organs ( NTP, 1993). In addition, cyanide is known to affect thyroid hormone levels,

and agents that affect thyroid hormone levels may also affect neurological development. Based on the lack of a multigeneration reproductive toxicity study and concerns about potential neurodevelopmental toxicity, a factor of 10 was used for database uncertainties for the Longerterm cyanide Health Advisory.

The data indicate that a reduced uncertainty factor for extrapolation from subchronic to chronic duration is appropriate. There was no progression in liver effects ( actually, a regression)

with increased exposure. Palmer and Olson ( 1979) found a LOAEL of 12 mg CN/ kg day for increased liver weight in male rats exposed via a drinking water study for 21 days, but there was no effect on liver weight in male rats exposed via drinking water for 13 weeks ( NTP, 1993). Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 20 Final draft There was an increase in liver weight in female rats in the NTP ( 1993) study, but this increase was comparable to the increase seen in males in the short term study; Palmer and Olson ( 1979) did not evaluate female rats. As discussed in the context of the Longer term Health Advisory for thiocyanate, progression of effect from subchronic to chronic exposure would not be expected for effects of cyanide that are secondary to thyroid hormone changes, since the thyroid responds rapidly to iodine like ions, and thiocyanate does not accumulate in the body. However, the mode of action of the critical effect on the male reproductive tract has not been identified ( e. g., whether it is due to cyanide or thiocyanate, and whether it is secondary to effects on thyroid hormones). It is also not known whether the observed effects would progress with continued exposure.

Therefore, the default factor of 10 is used for subchronic to chronic extrapolation for the male reproductive effects.

An endpoint by endpoint analysis is needed, however, for determination of the appropriate composite uncertainty factor for subchronic to chronic extrapolation and for database uncertainties. If the " true" critical effect is decreased weight of male reproductive organs ( as in the NTP, 1993 study), a factor of 10 would be needed for subchronic to chronic extrapolation,

but no database uncertainty factor would be needed ( since the critical effect has been identified).

If neurodevelopmental effects due to alterations in thyroid hormones are the " true" critical effect,

a full database uncertainty factor of 10 would be needed, but a factor of 1 would be adequate for subchronic to chronic extrapolation, as described in the context of the Longer term health advisory for thiocyanate. Similarly, if the " true" critical effect were a functional effect identified in a 2 generation reproduction study, a database factor of 3 would be adequate, based on the usual approach in the absence of a multigeneration study. No subchronic to chronic factor would Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 21 Final draft be needed, because a subchronic to chronic uncertainty factor is not usually used with a multigeneration study. Therefore, regardless of the " true" critical effect, a composite uncertainty factor of 10 is adequate for database insufficiencies and for subchronic to chronic extrapolation.

Thus, full factors of 10 were used for extrapolation from animals to humans and for human variability, and composite factor of 10 was used for database uncertainties and for subchronic to chronic extrapolation. A composite uncertainty factor of 1000 results. Because the principal study ( NTP, 1993) used drinking water administration, no additional modifying factor is needed. The same composite uncertainty factor of 1000 would be used with the BMDL of 0.79 mg/ kg day, in light of the sensitivity of the critical effect ( decreased left epididymis weight), and because the BMDL is markedly lower than the NOAEL.

Derivation of the Lifetime HA based on the study NOAEL Step 1: Determination of a RfD for Cyanide RfD = 4.5 mg/ kg day = 0.0045 mg/ kg day ( rounded to 0.005 mg/ kg day) 1000 where:

4.5 mg/ kg day = NOAEL, based on absence of effects on male reproductive organ weight ( epididymis, cauda epididymis, and testes weight) and spermatid counts in rats exposed to cyanide in drinking water for 13 weeks, with a corresponding LOAEL of 12.5 mg/ kg day ( NTP, 1993).

1000 = composite uncertainty factor, chosen to account for extrapolation from a NOAEL in animals, protection of sensitive subpopulations, and insufficiencies in the database. The database insufficiencies include the absence of a 2 generation reproduction study, and the absence of adequate developmental toxicity studies, particularly a neurodevelopmental toxicity study, in light of the potential for low level cyanide exposure to decrease T4 levels. Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 22 Final draft Step 2: Determination of a Drinking Water Equivalent Level ( DWEL) for Cyanide DWEL = ( 0.0045 mg/ kg day) ( 70kg) = 0.16 mg/ L ( rounded to 0.2 mg/ L) ( 2 L/ day)

where:

0.0045 mg/ kg day = RfD ( before rounding) 70 kg = assumed body weight of an adult 2 L/ day = assumed drinking water consumption of a 70 kg adult Step 3: Determination of Lifetime HA for Cyanide Lifetime HA = ( 0.16 mg/ L) ( 20%) = 0.032 mg/ L ( rounded to 30

g/ L)

where: 0.16 mg/ L = DWEL 20% = assumed relative source contribution from water Derivation of the Lifetime HA based on the study BMDL Step 1: Determination of a RfD for Cyanide RfD = 0.79 mg/ kg day = 0.00079 mg/ kg day ( rounded to 0.0008 mg/ kg day) 1000 where:

0.79 mg/ kg day = BMDL, based on decreased left epididymis weight in male rats exposed to cyanide in drinking water for 13 weeks ( NTP, 1993).

1000 = composite uncertainty factor, chosen to account for extrapolation from a BMDL in animals, protection of sensitive subpopulations, and insufficiencies in the database. The database insufficiencies include the absence of a 2 generation reproduction study, and the absence of adequate developmental toxicity studies, particularly a neurodevelopmental toxicity study, in light of the potential for low level cyanide exposure to decrease T4 levels.

Step 2: Determination of a Drinking Water Equivalent Level ( DWEL) for Cyanide DWEL = ( 0.00079 mg/ kg day) ( 70kg) = 0.028 mg/ L ( rounded to 0.03 mg/ L) ( 2 L/ day) Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 23 Final draft where:

0.00079 mg/ kg day = RfD ( before rounding) 70 kg = assumed body weight of an adult 2 L/ day = assumed drinking water consumption of a 70 kg adult Step 3: Determination of Lifetime HA for Cyanide Lifetime HA = ( 0.028 mg/ L) ( 20%) = 0.0056 mg/ L ( rounded to 6

g/ L)

where: 0.028 mg/ L = DWEL 20% = assumed relative source contribution from water B. 2 Thiocyanate B. 2.1 One day Health Advisory for Thiocyanate No suitable studies were located. Decreased hemoglobin was observed in guinea pigs following 2 doses of 120 mg SCN/ kg/ dose, the only dose tested ( Taubman and Heilborn, 1930),

but this study is not of adequate quality to use as the basis for the development of a Health Advisory. In the absence of adequate data, the ten day HA value is recommended as a conservative estimate of an appropriate one day HA value.

B. 2.2 Ten day Health Advisory for Thiocyanate No NOAEL was identified in the available short term animal toxicity studies ( Wolff et al.,

1946; Rawson et al., 1944; De Groot et al., 1991). The lowest reliable LOAEL was 123 mg SCN/ kg day, based on decreased T4, increased thyroid weight, and activation of thyroid follicles ( De Groot et al., 1991); no NOAEL was identified in this study. Smith and Rudolf ( 1928)

demonstrated a 15 30 mm Hg drop in systolic blood pressure in normotensive human subjects treated with three daily doses of 3.3 mg SCN/ kg day ( for a cumulative dose of 10 mg/ kg day) for Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 24 Final draft 1 week, with no other effects reported. However, the likely target organ ( the thyroid) was not evaluated in this study. A human LOAEL of 1.5 1.7 mg SCN/ kg day was identified by Palmer et al. ( 1929) and Palmer and Sprague ( 1929), based on weakness, angina, and precordial distress in hypertensive subjects administered the TWA dose over 3 4 weeks ( decreasing daily doses). It was unclear, however, whether the observed effects were secondary to a rapid decrease in blood pressure. Thyroid effects were not evaluated by Palmer and colleagues. However, the LOAEL in the hypertensive subjects is consistent with the NOAEL of 0.11 mg SCN/ kg day identified by Dahlberg et al. ( 1984), based on the absence of changes in T3, T4, or thyrotropic hormone in normotensive subjects treated for 12 weeks. Higher doses were not tested by Dahlberg et al.

( 1984). The results of Dahlberg et al. ( 1984) indicate that the NOAEL for a ten day exposure would be 0.11 mg SCN/ kg day or higher, and that the thyroid is adequately protected by deriving the Ten day Health advisory from a LOAEL of 1.5 mg SCN/ kg day, and using an uncertainty factor of 10 to extrapolate to a NOAEL. Thus, the Ten day Health Advisory is based on the LOAEL of 1.5 mg SCN/ kg day identified by Palmer et al. ( 1929) and Palmer and Sprague ( 1929)

in hypertensive subjects. An uncertainty factor of 10 is used to protect sensitive subpopulations,

and an uncertainty factor of 10 is used to account for the use of a LOAEL, resulting in a composite uncertainty factor of 100. ( See Table VIII 6.)

( 1.5 mg/ kg day) ( 10 kg) Ten day HA ( for a child) = = 0.15 mg/ L, rounded to 0.2 mg/ L ( 100)( 1 L/ day)

where:

1.5 mg/ kg day = LOAEL, at which weakness, angina, and precordial distress were seen in hypertensive subjects administered the TWA dose for 3 4 weeks ( Palmer et al., 1929; Palmer and Sprague, 1929) Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 25 Final draft 10 kg = assumed body weight of a child.

100 = composite uncertainty factor, chosen to account for extrapolation from a LOAEL in humans and protection of sensitive subpopulations 1 L/ day = assumed daily water consumption by a 10 kg child.

B. 2.3 Longer term Health Advisory for Thiocyanate Long term toxicity data are available from mice ( Nagasawa et al., 1980) and rats ( Pyska,

1977; Philbrick et al., 1979; Kanno et al., 1990) exposed to thiocyanate via the diet and drinking water. These data support the identification of thyroid as the target organ. None of the studies that evaluated thyroid endpoints identified a NOAEL. The lowest LOAEL was identified by Philbrick et al. ( 1979), who reported a statistically significant decrease in plasma T4 and increased thyroid weight ( but no thyroid histopathology) in rats administered 67 mg SCN/ kg day in diet for 11.5 months; plasma T4 was increased at this dose after 4 months of exposure. Pyska et al.

( 1977) also reported decreased plasma protein bound iodine ( a measure of plasma T3 and T4) in rats administered 90 mg SCN/ kg day for approximately 10 weeks in drinking water. No standard multigeneration reproduction or developmental toxicity study of thiocyanate by any route was located. However, a multigeneration study that evaluated thyroid hormone levels in rat pups is available ( Raghunath and Bala, 1998), as well as two developmental toxicity studies that evaluated thyroid effects in rat pups ( Bala et al., 1996; Kreutler et al., 1978). These studies have found effects on maternal thyroid hormones at the lowest dose evaluated in dams ( 52 mg SCN/ kg day; Pyska, 1977) and in pups ( 108 mg SCN/ kg day, Raghunath and Bala, 1998). A mouse reproductive toxicity study found no effect on pup weight up to 341 mg SCN/ kg day, but did not evaluate pup morphology or thyroid endpoints . Kreutler et al. ( 1978) found increased relative thyroid weight in pups born to the lowest dose group, 5.6 mg SCN/ kg day in drinking Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 26 Final draft water. Unfortunately, thyroid hormone levels and histopathology were not evaluated, and increases in thyroid weight may be adaptive or adverse, depending on the alterations in thyroid hormones and histopathology.

Despite the availability of LOAELs ( without corresponding NOAELs) for thyroid hormone effects, one can not confidently conclude that the NOAEL for neurodevelopmental effects is known within a factor of ten ( the uncertainty factor for extrapolating from a LOAEL to a NOAEL) of the LOAEL for thyroid hormone changes in pups. There are several reasons for this uncertainty. First, the degree of decrease in serum T4 that results in neurodevelopmental effects has not been determined. In addition, thyroid hormones are controlled by a negative feedback loop, and rats rapidly up regulate serum thyroid hormone levels. Because of this rapid feedback, the peak degree of change in hormone levels in the developmental toxicity studies could have been much larger at earlier time points. Since developmental toxicity depends on exposure during a critical developmental window, a large, but brief, change in thyroid hormone levels might result in marked neurodevelopmental effects. This means that the ratio between the free standing LOAELs identified for changes in thyroid hormone levels in rat pups and the short term NOAEL ( in the same species) for this effect may be more than an a factor of ten. Because a short term decrease in T4 levels ( of sufficient magnitude) can affect neurodevelopment if the change occurs during the window of sensitivity at the appropriate stage in development, even short term deviations in T4 are of concern. Although the increased relative thyroid weight observed by Kreutler et al. ( 1978) at 5.6 mg/ kg day may have been adaptive, that observation ( at a dose 20

fold below the LOAEL for thyroid hormone changes observed by Bala et al. [ 1996]) suggests that there may have been a short term change in thyroid hormone levels at lower doses, to which the Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 27 Final draft rats later adapted. Based on these considerations, a neurodevelopmental toxicity study would be important in determining whether the brain is affected at doses below those observed to affect serum hormone levels in rat pups.

Human studies provide a reliable apparent NOAEL/ LOAEL pair, based on thyroid effects in normotensive populations. Dahlberg et al. ( 1984) found no effect on serum T3, T4, or thyrotropic hormone, or the T3: T4 ratio in 37 volunteers administered 8 mg/ day thiocyanate in milk ( 0.11 mg SCN/ kg day) for 12 weeks. Banerjee et al. ( 1997) found decreased serum T4 and increased TSH compared to matched controls in 35 women in India who ingested thiocyanate and hydrogen peroxide as a bacteriocide in milk for at least 5 years. The exposed women ingested approximately 0.19 mg SCN/ kg day. This LOAEL is approximately a factor of 2 higher than the NOAEL, suggesting that the NOAEL/ LOAEL boundary is well defined. The LOAEL identified by Banerjee et al. ( 1997) is also supported by the results of Beamish et al. ( 1954), who reported decreased protein bound plasma iodine and decreased thyroid uptake of iodine in a clinical study of 5 hypertensive adults in Canada. Although doses were not reported in the Beamish et al.

( 1954) study, the study subjects had plasma thiocyanate levels of 1.3 5 mg/ 100 mL,

concentrations very similar to the average blood thiocyanate level of 1.3 mg/ 100 mL reported by Banerjee et al. ( 1997). This study was not considered co critical, because actual intake was not reported. Note that the Indian population might constitute a sensitive population if the women had borderline protein intake. There is, however, no support for this hypothesis, other than the high percentage of the Indian population that is malnourished. Since the study subjects were from Calcutta, a coastal city, it appears that an iodine deficiency was unlikely. In the absence of definitive information on a protein or iodine deficiency, the study population is conservatively Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 28 Final draft assumed not to constitute a sensitive population. Although the NOAEL is from a study of shorter duration, the rapid nature of the thyroid response to iodine like ions, and the lack of accumulation of thiocyanate in the body, indicate that the 12 week NOAEL is adequately protective for longerterm exposure. Therefore, progression of the effect with increased exposure duration from 12 weeks to 5 years would not be expected.

Using the NOAEL of 0.11 mg SCN/ kg day from Dahlberg et al. ( 1984), an uncertainty factor of 10 is applied to protect sensitive subpopulations. A factor of 10 is applied for database deficiencies. A factor of 3 would normally be applied to account for the absence of a standard multi generation study. This default was increased to 10 because a NOAEL for effects on the thyroid in developing fetuses has not been identified; effects on the fetal thyroid are of particular concern, because neurodevelopmental effects are known to result from decreased T4 levels ( Chan and Kilby, 2000), and no neurological effects were evaluated in pups exposed to thiocyanate during gestation. The composite uncertainty factor is 100.

( 0.11 mg/ kg day) ( 10 kg) Longer Term HA ( for a child) = = 0.011 mg/ L ( rounded to 0.01 mg/ L) ( 100)( 1 L/ day) where:

0.11 mg/ kg day = NOAEL, based on absence of effects on hormone levels in human subjects exposed to thiocyanate for 12 weeks ( Dahlberg et al., 1984)

10 kg = assumed body weight of a child.

100 = composite uncertainty factor, chosen to account for inter individual variability in humans, and insufficiencies in the database, including the absence of a 2 generation reproductive study, and the lack of adequate information on effects on the fetal thyroid effects and associated neurodevelopmental effects Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 29 Final draft 1 L/ day = assumed daily water consumption by a 10 kg child.

The Longer term HA for a 70 kg adult consuming 2 L/ day of water is calculated as follows:

( 0.11 mg/ kg day) ( 70 kg) Longer Term HA ( for adults) = = 0.039 mg/ L ( rounded to 0.04 mg/ L) ( 100)( 2 L/ day) where:

0.11 mg/ kg day = NOAEL, based on absence of effects on thyroid hormone levels in human subjects exposed to thiocyanate for 12 weeks ( Dahlberg et al., 1984).

70 kg = assumed body weight of an adult.

100 = composite uncertainty factor, chosen to account for inter individual variability in humans and insufficiencies in the database, including the absence of a 2 generation reproductive study, and the lack of adequate information on effects on the fetal thyroid effects and associated neurodevelopmental effects.

2 L/ day = assumed daily water consumption by a 70 kg adult.

B. 2.4 Reference Dose, Drinking Water Equivalent Level, and Lifetime Health Advisory for Thiocyanate The only chronic animal studies of thiocyanate did not evaluate sensitive endpoints, such as effects on thyroid hormones. As discussed in the context of the Longer term Health Advisory,

the thyroid responds rapidly to effects of ions such as thiocyanate. Therefore, the human NOAEL of 0.11 mg SCN/ kg day, based on a 12 week study ( Dahlberg et al., 1984), coupled with a human LOAEL of 0.19 mg SCN/ kg day, based on a 5 year study ( Banerjee et al., 1997), is appropriate as the basis for the Lifetime health advisory, with no further adjustment. The LOAEL identified by Beamish et al. ( 1954) also supports the identified critical effect level. However, Beamish et al. Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 30 Final draft ( 1954) is not appropriate as a co critical study, because only serum thiocyanate levels were reported, without any information on intake.

As noted for the Longer term Health Advisory, two uncertainty factors are used with this NOAEL. A factor of 10 is applied to protect sensitive subpopulations, and a factor of 10 is applied for database deficiencies, including the absence of a 2 generation reproductive study and the absence of a neurodevelopmental toxicity study, in light of the neurological effects of decreased T4 levels in the developing animal. No uncertainty factor is necessary for subchronicto chronic extrapolation, in light of the rapid response of the thyroid.

Step 1: Determination of RfD for Thiocyanate ( 0.11 mg/ kg day) RfD = = 0.0011 mg/ kg day, rounded to 0.001 mg/ kg day ( 100)

where:

0.11 mg/ kg day = NOAEL, based on absence of effects on thyroid in human subjects exposed to thiocyanate for 12 weeks ( Dahlberg et al., 1984).

100 = composite uncertainty factor chosen to protect sensitive subpopulations and to account for insufficiencies in the database, including the absence of a 2 generation reproductive study, and the lack of adequate information on effects on the fetal thyroid effects and associated neurodevelopmental effects.

Step 2: Determination of the Drinking Water Equivalent Level ( DWEL) for Thiocyanate ( 0.0011 mg/ kg day) ( 70 kg) DWEL = = 0.039 mg/ L ( rounded to 0.04 mg/ L) ( 2 L/ day)

where:

0.0011 mg/ kg day = RfD ( before rounding)

70 kg = assumed body weight of an adult. Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 31 Final draft 2 L/ day = assumed water consumption of a 70 kg adult.

Step 3: Determination of Lifetime HA for Thiocyanate Lifetime HA = ( 0.039 mg/ L) ( 20%) = 0.0078 mg/ L ( rounded to 8 µ g/ L)

where:

0.039 mg/ L = DWEL ( before rounding)

20% = assumed relative source contribution from water B. 3 Cyanogen chloride.

No suitable studies are available for developing quantitative assessments of cyanogen chloride. No oral studies of cyanogen chloride toxicity are available. The available studies ( Reed,

1920; Haymaker, 1952; Aldridge and Evans, 1946; Flury and Zernik, 1931) were done prior to the development of modern toxicology methods, used inadequate numbers of animals, did not evaluate standard endpoints, and used either inhalation or injection exposure. In the absence of adequate toxicity data on cyanogen chloride itself, an alternative approach to dose response assessment for cyanogen chloride is to use the quantitation for cyanogen chloride metabolites as a surrogate for cyanogen chloride. Early studies demonstrate that cyanide and thiocyanate are metabolites of cyanogen chloride. Cyanide accounts for approximately 30 40% of the cyanogen chloride dose at high doses, and 60 80% of the cyanogen chloride dose at lower doses ( Aldridge and Evans, 1946; Midwest Research Institute, 1997). Because thiocyanate was not measured and a mass balance was not calculated, actual conversion to cyanide may have been higher, since some of the cyanide would have been converted to thiocyanate. Conversion to cyanide would be Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 32 Final draft expected to be on the higher end ( i. e., closer to the 80% estimate) at environmentally relevant doses. As discussed in Chapter 3, there are uncertainties in this conclusion, in light of the absence of quantitative mass balance data on cyanogen chloride metabolism. For example, it is not known whether cyanide and thiocyanate together account for 100% of the cyanogen chloride dose, or if there are additional metabolites of cyanogen chloride that have yet to be identified. As discussed in Chapter 3, the aqueous chemistry of cyanogen chloride indicates that other potential metabolites are cyanate, cyanamide, and HCl ( Price et al., 1947; Migridichian, 1946); these chemicals are discussed in Appendices C, D, and E, respectively. HCl production has not been quantified in vivo or in vitro, but mass balance considerations mean that metabolism of cyanogen chloride would produce a mole of HCl for every mole of cyanogen chloride.

As discussed in Chapters 3 and 7, ingested cyanogen chloride may react with nucleophiles such as proteins in the gastrointestinal tract contents; cyanate and cyanamide might also be formed slowly in the gastrointestinal tract, but hydrolysis would be expected to be very slow, due to the low pH in the stomach. Therefore it is plausible that absorption of cyanogen chloride is primarily as the parent form. Absorbed cyanogen chloride ( or its metabolites) would enter the blood stream either directly or via the portal vein. Cyanogen chloride that is absorbed via the portal vein would be expected to undergo rapid glutathione mediated reduction to cyanide and HCl in the blood and in the liver during first pass metabolism, in light of the rapid reactivity of cyanogen chloride in blood, and the high glutathione concentration in the liver. Cyanide produced in the portal vein or the liver would be expected to be metabolized to thiocyanate via the rhodanese enzyme in the liver. Cyanogen chloride absorbed directly into the blood stream would also be expected to react rapidly in the blood to form cyanide and HCl. Cyanide formed in the Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 33 Final draft blood following intestinal absorption of cyanogen chloride would be metabolized by rhodanese found elsewhere in the body, or by mercaptopyruvate sulfur transferase, which is concentrated in blood cells.

Based on this hypothesized toxicokinetic pathway, the primary metabolites of cyanogen chloride would be cyanide, HCl, and thiocyanate, and consideration of the toxic effects of cyanogen chloride should focus on these compounds. In addition to effects attributable to cyanide, HCl, and thiocyanate, primary irritation to the gastrointestinal tract might also occur,

based on the reports of irritation following exposure to cyanogen chloride vapor ( Flury and Zernick, 1931; Prentiss, 1937; Michigan Department of Public Health, 1977; Reed, 1920;

Aldridge and Evans, 1946; Haymaker et al., 1952). Insufficient data are available to quantify the oral doses that would cause this effect, but it is reasonable to expect that this irritation would occur at drinking water concentrations above those that could result in systemic effects. The hypothesized metabolic pathway would suggest that HCl production might result in transient pH decreases, particularly in the portal vein and liver ( with an associated potential for liver damage),

depending on the size of the bolus dose and the buffering capacity of these tissues. However, no effect on liver weight, liver histopathology, or serum enzymes indicative of liver damage was observed in rats and rabbits at drinking water doses of 36 and 25 mg HCl/ kg day, respectively ( Tober Meyer et al., 1981); there was also no acidosis at this dose. If HCl and cyanogen chloride are absorbed to a similar degree and at similar rates, these results would suggest that cyanogen chloride at similar molar doses would also not result in liver effects from HCl production.

However, uncertainties remain as to how differences between HCl and cyanogen chloride in the rate or degree of absorption from the stomach and/ or intestine would affect the peak hydrogen ion Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 34 Final draft concentration in the liver. Liver effects were not investigated in the feed studies that produced acidosis at higher doses of HCl ( Throssell et al., 1995; 1996), but liver damage does not appear to be one of the effects associated with acidosis ( Bookallil, 2001), suggesting that the liver is resistant to moderate changes in blood pH. Liver effects were not reported in the few studies on cyanogen chloride inhalation or injection ( Flury and Zernick, 1931; Prentiss, 1937; Michigan Department of Public Health, 1977; Reed, 1920; Aldridge and Evans, 1946; Haymaker et al.,

1952), but the liver was not evaluated, and the potential liver effects from cyanogen chloride ingestion would be related to first pass metabolism, which would not apply to these inhalation and injection studies. Overall, the data suggest that the HCl produced from metabolism of cyanogen chloride would not be sufficient to damage the liver, but there are several uncertainties in the data.

A second possible effect of HCl production from cyanogen chloride might be systemic metabolic acidosis. Due to the complexities of determining how the body adjusts to increased acid in the system, the potential for acidosis resulting from cyanogen chloride ingestion is discussed below in the context of specific potential exposure levels.

Because the only identified metabolites of cyanogen chloride are cyanide and thiocyanate,

consideration of appropriate surrogates was limited to these two chemicals. The calculated RfD for cyanide is 0.005 mg/ kg day based on the NOAEL, or 0.0008 mg/ kg day based on the BMDL.

For thiocyanate, the calculated RfD is 0.001 mg/ kg day. The corresponding values for the cyanogen chloride RfD that would be calculated from the cyanide NOAEL based RfD, the cyanide BMDL based RfD, and the thiocyanate RfD, after adjusting for the molecular weights of cyanogen chloride and the metabolites, are 0.01, 0.002, and 0.001 mg/ kg day, respectively. This Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 35 Final draft adjustment is based on a determination of the amount of the metabolite that could be produced from a given dose of cyanogen chloride, assuming that all of the cyanogen chloride is converted to that metabolite.

Cyanide has been selected as the surrogate for cyanogen chloride based on the fact that it is a known metabolite that represents a major portion of the cyanogen chloride dose. Although the lowest cyanogen chloride RfD would have resulted if thiocyanate were used as a surrogate,

the toxicokinetic data on thiocyanate and cyanogen chloride are insufficient to use thiocyanate as the surrogate. There are no quantitative data on thiocyanate production from cyanogen chloride.

If thiocyanate were to be used as a surrogate, quantitative metabolic data ( including mass balance data) on the degree of cyanogen chloride conversion to thiocyanate at environmentally relevant doses would be needed. Therefore, the uncertainty is greater when thiocyanate ( a more downstream metabolite) is used as a surrogate instead of cyanide. Given that a significant portion of an ingested cyanide dose would react or be excreted prior to metabolism to thiocyanate, the derivation of a thiocyanate RfD that is lower than the cyanide RfD is not inconsistent with the fact that cyanide is metabolized to thiocyanate. Similarly, a portion of an ingested cyanogen chloride dose may be converted to products other than cyanide or thiocyanate.

If the cyanogen chloride RfD were calculated using thiocyanate as the surrogate, the RfD would be approximately 1/ 10 the RfD calculated using cyanide as a surrogate ( when the RfD is based on the NOAEL) or ½ the RfD calculated using cyanide as a surrogate ( when the RfD is calculated based on the BMDL). Therefore, using the NOAEL based cyanide RfD, cyanide would be sufficiently health protective as a surrogate unless direct conversion to thiocyanate were Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 36 Final draft more than 1/ 10 of the cyanogen chloride dose. Similarly, using the BMDL based cyanide RfD,

cyanide would be sufficiently health protective as a surrogate unless direct conversion to thiocyanate were more than 1/ 2 of the cyanogen chloride dose. Given that conversion of cyanogen chloride to cyanide is dose related, and 80% of cyanogen chloride is converted to cyanide at low doses, it is likely that < 20% of cyanogen chloride is converted directly to thiocyanate at environmentally relevant doses, and plausible that this direct conversion accounts for < 10% of the cyanogen chloride dose. In light of the greater uncertainty involved in using thiocyanate as a surrogate, and given that the factor of 2 difference between the 20% and 10%

conversion is well within the uncertainty of the method, cyanide is preferred as the surrogate for derivation of the cyanogen chloride RfD.

Therefore, the quantitative assessments developed for cyanide are applied to cyanogen chloride by adjusting the value by the ratio of the molecular weights of cyanogen chloride and cyanide, 61.5: 26, ( approximately 2.36). Confidence in the use of cyanide as the surrogate RfD is reduced because the critical effect for cyanide, male reproductive toxicity, was not evaluated in the limited studies on cyanogen chloride. However, it is plausible that male reproductive toxicity could be observed following long term exposure to cyanogen chloride, since cyanide is a major metabolite. It should be noted that there is considerable uncertainty in using cyanide as a surrogate for cyanogen chloride, due to the incompleteness of the data on cyanogen chloride toxicity and metabolism. Indeed, the database for cyanogen chloride is far less complete than the minimal database required for deriving even a " low confidence" RfD. Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 37 Final draft B. 3.1 One day Health Advisory for Cyanogen Chloride No suitable studies are available. In the absence of adequate data, the Ten day HA value is recommended as a conservative estimate of an appropriate One day HA value.

B. 3.2 Ten day Health Advisory for Cyanogen Chloride The Ten day Health Advisory for cyanide is based on a LOAEL of 14 mg CN/ kg day for increased liver weight in rats that ingested cyanide in drinking water for 21 days ( Palmer and Olson, 1979); a NOAEL in feed in the same study was not used, due to uncertainties in the analysis of the cyanide in feed. Similarly, a BMDL of 2.1 mg CN/ kg day identified for decreased plasma TSH and increased thyroid weight in a rat dietary study ( Kreutler et al., 1978) was not used for the development of the Health Advisory, due to the large degree of extrapolation below the data. An uncertainty factor of 1000 was used, based on factors of 10 each for interspecies extrapolation, protective of sensitive subpopulations, and extrapolation from a LOAEL to a NOAEL. This value is used for cyanogen chloride, after adjustment for the relative molecular weights of cyanide and cyanogen chloride. ( See Table VIII 7.)

Ten day HA for child = ( 0.14 mg/ L) * ( 61.5/ 26) = 0.33 mg/ L ( rounded to 0.3 mg/ L)

where:

0.14 mg/ L = child Ten day HA for cyanide ( prior to rounding)

61.5/ 26 = the molecular weight ratio of cyanogen chloride to cyanide Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 38 Final draft To determine whether systemic acidosis resulting from HCl production may be a concern at this exposure level, the number of mmoles of HCl that could be produced from this dose of cyanogen chloride is determined:

Dose = ( 0.3 mg/ L) ( 1 L/ day) = 0.0008 mmol HCl/ kg day ( 10 kg) ( 36.5)

where:

0.3 mg/ kg day = Ten day Health Advisory 10 kg = assumed body weight of a child. 1 L/ day = assumed daily water consumption by a 10 kg child. 36.5 = molecular weight of HCl For comparison, the average net ( systemic) acid production for adults is about 60 milliequivalents ( 60 mmol) per day, corresponding to approximately 0.9 mmol/ kg day ( assuming a body weight of 70 kg). Thus, the amount of HCl produced from cyanogen chloride at a drinking water concentration corresponding to the Ten day Health Advisory is less than 1/ 1000 the daily acid production, and would be toxicologically inconsequential. This means that systemic acidosis would not be a concern.

B. 3.3 Longer term Health Advisory for Cyanogen Chloride The Longer term Health Advisory for cyanide is based on the finding of male reproductive effects in rats at a NOAEL of 4.5 mg/ kg day ( NTP, 1993). An uncertainty factor of 1000 was used, based on factors of 10 each for interspecies extrapolation, protection of sensitive subpopulations, and for an incomplete database ( including the lack of a 2 generation reproduction study, and lack of developmental toxicity studies, particularly the absence of information on neurodevelopmental effects). An alternative Longer term Health Advisory for cyanide was developed based on a BMDL of 0.79 mg/ kg day for decreased epididymis weight in the NTP Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 39 Final draft ( 1993) study, using the same uncertainty factors. The Longer term Health Advisory based on the NOAEL, as well as the one based on the BMDL, were used for cyanogen chloride, after adjusting for the relative molecular weights of cyanide and cyanogen chloride. An additional modifying factor to account for production of HCl is not needed, because the critical effect occurs in the male reproductive tract, while the liver is the site where HCl formed from cyanogen chloride would be formed and would be the target of most concern for HCl exposure.

Derivation of the Longer term HA based on the study NOAEL Longer term HA for child = ( 0.045 mg/ L) ( 61.5/ 26) = 0.106 mg/ L ( rounded to 0.1 mg/ L) where: 0.045 mg/ L = child Longer term HA for cyanide ( prior to rounding) 61.5/ 26 = the molecular weight ratio of cyanogen chloride to cyanide Longer term HA for adult = ( 0.158 mg/ L) ( 61.5/ 26) = 0.37 mg/ L ( rounded to 0.4 mg/ L) where: 0.158 mg/ L = adult Longer term HA for cyanide ( prior to rounding) 61.5/ 26 = the molecular weight ratio of cyanogen chloride to cyanide Derivation of the Longer term HA based on the study BMDL Longer term HA for child = ( 0.0079 mg/ L) ( 61.5/ 26) = 0.019 mg/ L ( rounded to 0.02 mg/ L) where: 0.0079 mg/ L = child Longer term HA for cyanide ( prior to rounding) 61.5/ 26 = the molecular weight ratio of cyanogen chloride to cyanide Longer term HA for adult = ( 0.028 mg/ L) ( 61.5/ 26) = 0.066 mg/ L ( rounded to 0.07 mg/ L) where: 0.028 mg/ L = adult Longer term HA for cyanide ( prior to rounding) 61.5/ 26 = the molecular weight ratio of cyanogen chloride to cyanide Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 40 Final draft B. 3.4 Reference Dose, Drinking Water Equivalent Level, and Lifetime Health Advisory for Cyanogen Chloride The RfD for cyanide is based on the finding of male reproductive effects ( decreased epididymis and testes weight, decreased spermatid counts) in rats at a NOAEL of 4.5 mg/ kg day ( NTP, 1993). An uncertainty factor of 1000 was used, based on factors of 10 each for interspecies extrapolation and protection of sensitive subpopulations, and a factor of 10 for subchronic to chronic extrapolation and database considerations combined. An alternative RfD for cyanide was developed based on a BMDL of 0.79 mg/ kg day for decreased epididymis weight in the NTP ( 1993) study, using the same uncertainty factors. The RfD based on the NOAEL, as well as the one based on the BMDL, were used for cyanogen chloride, after adjusting for the relative molecular weights of cyanide and cyanogen chloride. An additional modifying factor to account for production of HCl is not needed, because the critical effect occurs in the male reproductive tract, while the liver is the site where HCl formed from cyanogen chloride would be formed and would be the target of most concern for HCl exposure.

Derivation of the Lifetime HA based on the study NOAEL Step 1: Determination of RfD for Cyanogen chloride.

RfD = ( 0.0045 mg/ kg day) ( 61.5/ 26) = 0.011 mg/ kg day ( rounded to 0.01 mg/ kg day)

where: 0.0045 mg/ kg day = the RfD for cyanide ( before rounding) 61.5/ 26 = the molecular weight ratio of cyanogen chloride to cyanide Step 2: Determination of a Drinking Water Equivalent Level ( DWEL) for Cyanogen chloride DWEL = ( 0.011 mg/ kg day) ( 70 kg) = 0.39 mg/ L ( rounded to 0.4 mg/ L) ( 2 L/ day) Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 41 Final draft where:

0.011 mg/ kg day = RfD ( prior to rounding) 70 kg = assumed body weight of an adult 2 L/ day = assumed drinking water consumption of a 70 kg adult Step 3: Determination of Lifetime HA for cyanogen chloride Lifetime HA = ( 0.39 mg/ L) ( 20%) = 0.078 mg/ L ( rounded to 80

g/ L)

where: 0.39 mg/ L = DWEL ( before rounding) 20% = assumed relative source contribution from water Derivation of the Lifetime HA based on the study BMDL Step 1: Determination of RfD for Cyanogen chloride.

RfD = ( 0.00079 mg/ kg day) ( 61.5/ 26) = 0.0019 mg/ kg day ( rounded to 0.002 mg/ kg day)

where: 0.00079 mg/ kg day = the RfD for cyanide ( before rounding) 61.5/ 26 = the molecular weight ratio of cyanogen chloride to cyanide Step 2: Determination of a Drinking Water Equivalent Level ( DWEL) for Cyanogen chloride DWEL = ( 0.0019 mg/ kg day) ( 70 kg) = 0.067 mg/ L ( rounded to 0.07 mg/ L) ( 2 L/ day)

where:

0.0019 mg/ kg day = RfD ( prior to rounding) 70 kg = assumed body weight of an adult 2 L/ day = assumed drinking water consumption of a 70 kg adult Step 3: Determination of Lifetime HA for cyanogen chloride Lifetime HA = ( 0.067 mg/ L) ( 20%) = 0.013 mg/ L ( rounded to 10

g/ L)

where: 0.067 mg/ L = DWEL ( before rounding) 20% = assumed relative source contribution from water Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 42 Final draft C. Carcinogenic Effects Since there are no cancer bioassays of cyanogen chloride, the carcinogenicity assessment for this chemical is also conducted by evaluation of its known and putative metabolites. There are no cancer bioassays for cyanide or cyanate. There are no well conducted standard cancer bioassays of thiocyanate. Two oral carcinogenicity studies of thiocyanate in male Fischer rats were conducted in the same laboratory ( Lijinsky and Reuber, 1982; Lijinsky and Kovatch, 1989).

The only effect in the first study was an increase in liver tumors, and this was not confirmed at the higher dose tested in the second study, although a non statistically significant increase in thyroid tumors was observed. However, these studies suffer from a number of limitations. Only a single dose was tested in each study, only 20 rats/ group were tested, and the assay continued until all of the animals died ( increasing the background incidence of age related tumors, and thus decreasing the sensitivity of the bioassay), and results were not fully reported.

As described in Appendix D, one standard cancer bioassay on cyanamide was located.

NCI ( 1979) evaluated the carcinogenicity of calcium cyanamide in drinking water to F344 rats and B6C3F1 mice. No tumors related to treatment were observed in rats. Hemangiosarcomas were observed in male mice with a significant dose related trend. However, the tumor incidence was not statistically significant in either dose group when compared with controls. Malignant lymphomas were observed in female mice. Statistical analysis indicated a significant dose related trend for these tumors and a significant increase in incidence in the high dose group compared with controls. However, NCI ( 1979) reported that the historical control for this tumor type in B6C3F1 mice was 21%; therefore, the incidence of 5% in the matched control group may have Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 43 Final draft been abnormally low. NCI ( 1979) concluded that neither of these tumors was clearly related to administration of calcium cyanamide.

No data are available on the genotoxicity of cyanogen chloride, although a QSTR analysis predicted that cyanogen chloride was negative for carcinogenicity in rats and mice of both sexes ( Moudgal et al., 2000). Although the data for some known or potential metabolites of cyanogen chloride are limited, the available data suggest, overall, that none of the metabolites is genotoxic.

Overall, cyanide has been negative in bacterial mutagenicity assays ( De Flora, 1981; De Flora et al., 1984; NTP, 1993) and assays of DNA damage and repair ( De Flora et al., 1984; Painter and Howard, 1982), although a positive result was obtained in one strain of Salmonella with and without S9 activation ( Kushi et al., 1983). No assays of mammalian gene mutation or chromosome aberration are available. Genotoxicity data on thiocyanate are limited to marginal results in a S. typhimurium mutagenicity assay ( Kier, 1988, as reported by Rosenkranz and Klopman, 1990). Potassium thiocyanate does not have any structural alerts for genotoxicity ( Rosenkranz and Klopman, 1990). Cyanate was negative in the only study located, a mammalian gene mutation assay ( Melzer et al., 1983). Except for a positive result in Salmonella strain TA 1535 ( Zeiger, 1987), cyanamide was negative in bacterial gene mutation assays ( Loveless et al.,

1954; Zeiger, 1987), a Drosophila sex linked recessive assay ( Yoon et al., 1985), an assay of DNA damage in mammalian cells ( Sina et al., 1983), and a micronucleus assay in mice ( Menargues et al., 1984).

Based on these considerations, cyanogen chloride, cyanide, thiocyanate, cyanate, and cyanamide are all classified as Group D, Not Classifiable as to Human Carcinogenicity, using the Drinking Water Criteria Document for Cyanogen Chloride and Potential Metabolites EPA/ OW/ OST/ HECD VIII 44 Final draft U. S. EPA ( 1986) guidelines. Using the U. S. EPA ( 1999) Draft Guidelines for Carcinogen Risk Assessment, the data are inadequate for an assessment of the human carcinogenic potential of these compounds.

D. Characterization of Uncertainties and Data Gaps Very little toxicity information exists for cyanogen chloride, and the existing data on cyanogen chloride were developed primarily prior to the advent of modern methods of toxicology.

This assessment assumed that, based on the available data on cyanogen chloride metabolism,

structural similarity, and qualitative similarity of observed endpoints, it is appropriate to use cyanide as a surrogate for cyanogen chloride toxicity. However, that assumption has not been tested directly. Ideally, sufficient data from long term studies would be obtained so that the toxicity values can be derived from data on cyanogen chloride, rather than a surrogate. Failing that, data from short term toxicity studies can be used to identify targets of toxicity and to test the hypothesis that cyanide is an appropriate surrogate. High quality, in vivo metabolism studies ( or at least in vitro metabolism studies using cell extracts) would also be useful to determine which of the predicted metabolites actually are formed, and to quantitate their production. The NTP has recently selected cyanogen chloride for testing; such testing would significantly reduce the uncertainties in this assessment.

There are also uncertainties in the assessments for some of the metabolites, as discussed in the context of the database uncertainty factor for each RfD, and as discussed in Chapter 9.

Confidence in the calculated health advisories is also discussed in Chapter 9.


[Federal Register Volume 85, Number 24 (Wednesday, February 5, 2020)]
[Notices]
[Pages 6570-6571]
[FR Doc No: 2020-02237]

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DEPARTMENT OF HOMELAND SECURITY

Coast Guard

[Docket No. USCG-2019-0749]

National Boating Safety Advisory Committee; Initial Solicitation
for Members

AGENCY: U.S. Coast Guard, Department of Homeland Security.

ACTION: Request for applications.

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SUMMARY: The Coast Guard is requesting applications from persons
interested in serving as a member of the National Boating Safety
Advisory Committee (``Committee''). This recently established Committee
will advise the Secretary of the Department of Homeland Security on
matters relating to national boating safety. Please read this notice
for a description of the 21 Committee positions we are seeking to fill.

DATES: Your completed application should reach the Coast Guard on or
before April 6, 2020.

ADDRESSES: Applicants should send a cover letter expressing interest in
an appointment to the National Boating Safety Advisory Committee and a
resume detailing the applicant's experience. We will not accept a
biography. Applications should be submitted via one of the following
methods:
     By Email<EMAIL_ADDRESS>(preferred).
     By Mail: Commandant (CG-BSX-2), Attn: NBSAC ADFO, U.S.
Coast Guard Stop 7501, 2703 Martin Luther King Jr. Avenue SE,
Washington, DC 20593-7501.

FOR FURTHER INFORMATION CONTACT: Mr. Jeff Decker, Alternate Designated
Federal Officer of the National Boating Safety Advisory Committee;
Telephone<PHONE_NUMBER> or Email at<EMAIL_ADDRESS>
SUPPLEMENTARY INFORMATION: The National Boating Safety Advisory
Committee is a federal advisory committee. It will operate under the
provisions of the Federal Advisory Committee Act, 5 U.S.C. Appendix,
and the administrative provisions in Section 601 of the Frank LoBiondo
Coast Guard Authorization Act of 2018 (specifically, 46 U.S.C. 15109).
    The Committee was established on December 4, 2018, by the Frank
LoBiondo Coast Guard Authorization Act of 2018, which added section
15105, National Boating Safety Advisory Committee, to Title 46 of the
U.S. Code (46 U.S.C. 15105). The Committee will advise the Secretary of
Homeland Security on matters relating to national boating safety.
    We expect the Committee will hold meetings at least twice a year,
but it may meet even more frequently. The Committee is required to meet
at least once a year in accordance with 46 U.S.C. 15109(a). The
meetings are held at a location selected by the U.S. Coast Guard.
    All members serve at their own expense and receive no salary or
other compensation from the Federal Government. Members may be
reimbursed, however, for travel and per diem in accordance with Federal
Travel Regulations.
    Under provisions in 46 U.S.C. 15109(f)(6), if you are appointed as
a member of the Committee, your membership term will expire on December
31 of the third full year after the effective date of your appointment.
The Secretary may require an individual to have passed an appropriate
security background examination before appointment to the Committee, 46
U.S.C. 15109(f)(4). In this initial solicitation for Committee members,
we will consider applications for all 21 positions:
     Seven members shall represent State officials responsible
for State boating safety programs;
     Seven members shall represent recreational boat and
associated equipment manufacturers;
     Seven members shall represent the general public or
national recreational boating organizations and, of the seven, at least
five shall represent national recreational boating organizations.
    Each member of the Committee must have particular expertise,
knowledge, and experience in matters relating to the function of the
Committee, which is to advise the Secretary of Homeland Security on
matters related to national boating safety.
    If you are selected as a member drawn from the general public, you
will be appointed and serve as a Special Government Employee as defined
in Title 18, U.S.C section 202(a). Applicants for appointment as a
Special Government Employee are required to complete a Confidential
Financial Disclosure Report (OGE Form 450) for new entrants and if
appointed as a member must submit Form 450 annually. The Coast Guard
may not

release the reports or the information in them to the public except
under an order issued by a Federal Court or as otherwise provided under
the Privacy Act (5 U.S.C 552a). Only the Designated U.S. Coast Guard
Ethics Official or his or her designee may release a Confidential
Financial Disclosure Report. Applicants can obtain this form by going
to the website of the Office of Government Ethics (www.oge.gov), or by
calling or emailing the individual listed above in the FOR FURTHER
INFORMATION CONTACT section. Applications for member drawn from the
general public must be accompanied by a completed OGE Form 450.
    Registered lobbyists are not eligible to serve on Federal Advisory
Committees in an individual capacity. See ``Revised Guidance on
Appointment of Lobbyists to Federal Advisory Committees, Boards and
Commissions'' (79 FR 47482, August 13, 2014). Registered lobbyists are
``lobbyists,'' as defined in 2 U.S.C. 1602, who are required by 2
U.S.C. 1603 to register with the Secretary of the Senate and the Clerk
of the House of Representatives.
    The Department of Homeland Security does not discriminate in
selection of Committee members on the basis of race, color, religion,
sex, national origin, political affiliation, sexual orientation, gender
identity, marital status, disabilities and genetic information, age,
membership in an employee organization, or any other non-merit factor.
The Department of Homeland Security strives to achieve a widely diverse
candidate pool for all of its recruitment selections.
    If you are interested in applying to become a member of the
Committee, send your cover letter and resume to Mr. Jeff Decker,
Alternate Designated Federal Officer of the National Boating Safety
Advisory Committee via one of the transmittal methods in the ADDRESSES
section by the deadline in the DATES section of this notice.
    If you send your application to us via email, we will send you an
email confirming receipt of your application.

    Dated: January 31, 2020.
David C. Barata,
Captain, U.S. Coast Guard, Director of Inspections and Compliance.
[FR Doc. 2020-02237 Filed 2-4-20; 8:45 am]
 BILLING CODE 9110-04-P