Drinking Water Criteria Document for Haloacetonitriles EPA/ OW/ OST/ HECD Final Draft V  1 Chapter V. Health Effects in Animals Limited toxicity testing data are available for the HANs. Tables V 15 through V 18 at the end of the chapter provide a summary of the toxicity studies for BCAN, DBAN, DCAN, and TCAN for noncarcinogenic endpoints.

A. Short Term Exposures Acute oral LD 50 values for DBAN, DCAN, and TCAN in rodents have been reported by several investigators, as summarized in Table V 1. Acute oral LD 50 values for DBAN and DCAN were reported to range from 245 to 361 mg/ kg in male and female CD rats and mice given single oral doses dissolved in corn oil. Ataxia, depressed respiration, depressed activity, and coma preceded death. No consistent, compound related, gross pathological effects were observed at necropsy ( Hayes et al., 1986). In a limited report on the acute toxicity of DBAN, groups of 20 rats and mice ( sex and strain not specified) were given single gavage doses of 10% DBAN in corn oil ranging from 25 to 1,600 mg/ kg in rats and from 25 to 3,200 mg/ kg in mice. The LD 50 was estimated as 50 to 100 mg/ kg in rats and 50 mg/ kg in mice with slight to moderate convulsions the only symptom reported ( Eastman Kodak Co., 1992). Smyth et al. ( 1962) reported an oral LD 50 of 0.25 mL/ kg ( 360 mg/ kg) in male Wistar rats for TCAN ( vehicle was not specified). Drinking Water Criteria Document for Haloacetonitriles EPA/ OW/ OST/ HECD Final Draft V  2 Table V 1. Acute Oral Lethality of Haloacetonitriles.

Reference Species LD50 ( mg/ kg) Male Female DBAN Hayes et al. ( 1986) Mouse 289 ( 253 324) a 303 ( 269 342)

Hayes et al. ( 1986) Rat 245 ( 210 286) 361 ( 320 410)

Eastman Kodak Co., 1992 Mouse 50 ( sex not specified)

Eastman Kodak Co., 1992 Rat 50  100 ( sex not specified)

DCAN Hayes et al. ( 1986) Mouse 270 ( 241  303) 279 ( 263  296)

Hayes et al. ( 1986) Rat 339 ( 298  387) 330 ( 300  500)

TCAN Smyth et al. ( 1962) Rat 360 a95% Confidence limits.

Lin and Guion ( 1989) reported clinical signs of toxicity in a mechanistic study to evaluate the ability of HANs to deplete liver GSH and inhibit GST activity. Male Fischer 344 rats were administered single gavage doses of 0.75 mmol/ kg BCAN ( 116 mg/ kg), DBAN ( 149 mg/ kg),

DCAN ( 82 mg/ kg), or TCAN ( 108 mg/ kg) in tricaprylin. These doses represented 10 to 30% of the reported LD 50 for the individual compounds. The incidence of deaths at 18 hours was 2 of 20 for TCAN and 1 of 10 for DBAN. No deaths were reported for the other HANs. Overt signs of acute toxicity, including gasping and salivation, were observed for TCAN and DBAN. Drinking Water Criteria Document for Haloacetonitriles EPA/ OW/ OST/ HECD Final Draft V  3 The acute toxicity of DBAN has also been evaluated by the inhalation route. Sprague

Dawley rats ( 10 animals/ sex/ dose) were exposed for 6 hours to air concentrations of 3.5, 9.5,

17.9, or 41 ppm DBAN ( 28.5, 77.3, 145.6, or 333.5 mg/ m3)( Dow Chemical Co., 1992). Slight transient eye and nasal irritation was noted at the lowest concentration and increased in severity with increasing concentration. In addition, respiratory tract irritation was observed at 9.5 ppm and increased in severity with increasing concentration. Three males and two females did not survive following exposure to 9.5 ppm; all males and 6 females in the 17.9 ppm exposure group died. The LC 50 was reported as 9.6 ppm for males and 14.4 ppm for females, respectively,

calculated using a moving average method. In the low exposure group, no treatment related effects were observed at necropsy. In the 9.5 ppm group, nasal irritation and distended stomach were the principal findings in the animals that did not survive; no effects were observed in the surviving animals. In the two high exposure groups, necropsy revealed distended stomachs, and congested liver and lungs. No treatment related lesions were observed following gross pathological examination in the four females that survived the 17.9 ppm exposure. In the study by Smyth et al. ( 1962), no mortality was observed following a single 4 hour exposure to 125 ppm ( 738 mg/ m3) of TCAN in a group of six albino rats ( sex and strain not specified), indicating that the 4 hour LC 50 would be well above 125 ppm. Data were not reported for end points other than mortality in this study. Although the different exposure durations and strains used preclude a direct comparison, the finding that the 4 hour LC 50 for TCAN was likely much higher than the 6

hour LC 50 for DBAN suggests that DBAN has higher acute inhalation toxicity. No inhalation studies were identified for BCAN or DCAN. Drinking Water Criteria Document for Haloacetonitriles EPA/ OW/ OST/ HECD Final Draft V  4 In a poorly reported skin irritation study, undiluted DBAN was applied to the skin of guinea pigs ( 3 animals/ dose) at doses ranging from 0.01 to 1.0 mL/ kg ( 23 to 2,300 mg/ kg)( Eastman Kodak Co., 1992). The estimated LD 50 was 0.1 to 1.0 mL/ kg ( 230 to 2,300 mg/ kg). Irritant symptoms were recorded as moderate to gross edema, with necrosis of the entire patch area and hemorrhagic periphery at 24 hours. At 1 week, the authors reported heavy eschar that was broken at the edges with secondary eschar beneath. At 2 weeks, small secondary eschar was noted, with heavy scarring and alopecia. The LD 50 value for a single dermal application of TCAN to male New Zealand rabbits was reported to be 0.90 mL/ kg ( 1,296 mg/ kg)( Smyth et al.,

1962). TCAN is extremely irritating; 0.5 mL of a 1% solution produced a severe eye burn in rabbits, and 0.01 mL of the material caused necrosis when applied to the clipped skin of rabbits.

The study descriptions did not indicate if the TCAN was undiluted or applied in a solvent vehicle.

No dermal studies were identified for BCAN or DCAN.

In addition to acute lethality and irritation, other effects of short term exposure to DBAN and DCAN have been evaluated. The evaluation of several endpoints for DBAN toxicity have been conducted or are being planned. DBAN has been selected for a neurotoxicity study ( NTP,

2002), has been tested in a short term reproductive and developmental toxicity screening study ( R. O. W. Sciences, 1997), and has been evaluated in 14 day dose range finding studies ( NTP,

2002). These latter two studies are described in this section. Drinking Water Criteria Document for Haloacetonitriles EPA/ OW/ OST/ HECD Final Draft V  5 The short term effects of DBAN were evaluated as part of dose range finding studies conducted for a reproductive and developmental toxicity screening study ( R. O. W. Sciences,

1997). In the initial dose range finding study, male and female Sprague Dawley rats ( 5/ sex/ dose)

were exposed to drinking water containing 0, 250, 500, 1000, or 2000 ppm DBAN ( doses resulting from these exposures were not reported by the study authors). Significant decreases in body weight relative to controls and decreased food consumption were observed within 4 days of exposure, beginning at 250 ppm in males and 500 ppm in females. Water consumption was significantly decreased at the lowest concentration tested ( 250 ppm) in males and females. The onset of these effects led the authors to terminate the study, since the lowest concentration used in this first range finding study was judged to be too high for use in the main study. Therefore,

the rats were allowed to recover to control body weights, and were exposed to a lower set of range finding concentrations of 0, 7, 20, 70, or 200 ppm in drinking water for 2 weeks. The rats were evaluated for clinical signs of toxicity, body weight, and feed and water consumption.

Based on measured water consumption and body weights, the estimated dose levels in the second range finding study were 0, 0.7, 2.2, 5.8, and 13.2 mg/ kg/ day for the males and 0, 0.8, 2.4, 6.8,

and 17.9 mg/ kg/ day for the females. No clinical signs of toxicity were observed in any of the exposure groups. A significant decrease in body weights of roughly 10% was observed in males exposed to 20 or 200 ppm DBAN after 4 days of treatment. However, no similar decrease was observed in the males exposed to 70 ppm, and the mean final weight and total weight gain at the end of the study were not affected in any dose group. No significant effects on feed consumption were observed. Water consumption was significantly decreased ( p

0.05) on day 11 in females Drinking Water Criteria Document for Haloacetonitriles EPA/ OW/ OST/ HECD Final Draft V  6 exposed to 70 ppm ( to 71% of controls). Water consumption was significantly decreased at each time point examined ( study days 4, 8, 11, and 15) for males and females in the 200 ppm group.

The maximum decrease in water consumption was to 50% of controls for males and to 67% of controls for females. The absence of clinical signs of toxicity or body weight changes indicates that the highest concentration tested ( 200 ppm in the second range finding study) is a study NOAEL. Decreased water consumption is not judged to be an adverse toxicological effect, as this result may reflect poor palatability of the DBAN containing water. Based on these considerations, the study NOAEL is the highest dose tested ( 13.2 mg/ kg/ day in males; 17.9 mg/ kg/ day in females) and no LOAEL was determined.

In the main study ( described in more detail in the Reproductive and Developmental Toxicity Section), Sprague Dawley rats were given 0, 15, 50, or 150 ppm DBAN in their drinking water. Male rats ( 10 animals/ dose) were given treated water on study days 6 through 35, and were then examined for clinical pathology ( hematology and clinical chemistry), body and organ weight changes ( liver, right kidney, spleen, thymus, right testis, right epididymis, right cauda epididymis), sperm analyses, and histopathology. Estimated DBAN doses in males were 0, 1.4,

3.3, and 8.2 mg/ kg/ day as calculated by the study authors from body weight and water consumption data. Separate groups of female rats were exposed to the same concentrations as males on study days 1 34 ( Group A) or from gestation day 6 to postnatal day 1 ( Group B).

Estimated DBAN doses were 0, 1.8, 5.1, and 10.9 mg/ kg/ day for Group A females, and 0, 1.9,

5.3, and 10.8 mg/ kg/ day for Group B females. For both groups of females, clinical signs of Drinking Water Criteria Document for Haloacetonitriles EPA/ OW/ OST/ HECD Final Draft V  7 toxicity, body weight, and feed and water consumption were determined at various intervals, and general pathology was evaluated at necropsy. No compound related effect was observed for any of these parameters. The only biologically significant effect that was treatment related for males or females was a decrease in water consumption in the mid and high concentration groups, which might have been related to decreased water palatability. Since the males were examined for more sensitive endpoints than females ( i. e. clinical chemistry and histopathology were evaluated), the NOAEL for systemic effects in this study was 8.2 mg/ kg/ day, the highest dose level in males, and no LOAEL was identified.

The subacute toxicity of DBAN has also been evaluated by the NTP ( 2002) in B6C3F1 mice and F344 rats as part of initial dose range finding studies in support of chronic exposure studies that are currently in progress. For the mouse study, DBAN was administered in drinking water for 14 days to male and female B6C3F1 mice ( 5/ sex/ dose) at concentrations of 0, 12.5, 25,

50, 100, or 200 mg/ L. The corresponding doses reported by the study authors were 0, 2.1, 4.3,

8.2, 14.7, and 21.4 mg/ kg/ day for males and 2.0, 3.3, 10.0, 13.9, and 21.6 mg/ kg/ day for females.

Animals were observed for clinical signs of toxicity, as well as body weight, and organ weight and pathology. In addition, liver glutathione S transferase ( GST) activity was measured. The only treatment related effect was a decrease in water consumption in both males and females. The decrease in water consumption was concentration related, decreasing to 58% of controls for males and 54% controls for the 200 mg/ L group mice. Since the only effect observed was a concentration related decrease in water intake, which could reflect poor water palatability, the Drinking Water Criteria Document for Haloacetonitriles EPA/ OW/ OST/ HECD Final Draft V  8 high doses of 21.4 mg/ kg/ day for males and 21.6 mg/ kg/ day for females is considered a NOAEL.

No LOAEL is identified.

For the rat study, DBAN was administered in drinking water for 14 days to male and female F344 rats ( 5/ sex/ dose) at concentrations of 0, 12.5, 25, 50, 100, or 200 mg/ L. The corresponding doses reported by the study authors were 0, 2, 3, 7, 12, and 18 mg/ kg/ day for males and 2, 4, 7, 12, and 19 mg/ kg/ day for females. Animals were observed for clinical signs of toxicity, as well as body weight, and organ weight and pathology. In addition, liver glutathione

S transferase ( GST) activity was measured. A concentration related decrease in water consumption was observed for both males and females. Water consumption decreased to 60% of controls for males and 61% controls for females in the 200 mg/ L group mice. No effects other than decreased water consumption were noted in females. However, in males DBAN exposure caused a decrease in body weight gain and terminal body weight that was judged to be toxicologically significant only at the high dose. The reported body weight gains for males were 61.1, 66.3, 66.0, 65.2, 56.4, and 34.0 grams for the control and increasing dose groups,

respectively. Terminal body weights as percent of controls were 100, 104.5, 104.3, 101.3, 98.7,

82.7% for the control and increasing dose groups, respectively. Significantly decreased testes weights were observed in the high dose group males. This finding was accompanied by testicular atrophy in 2 of 5 males in this dose group. Elevated liver GST activity ( 126% of controls) was also reported for high dose males. Based on decreased body weight and decreased testes weight Drinking Water Criteria Document for Haloacetonitriles EPA/ OW/ OST/ HECD Final Draft V  9 and pathology in males, the NOAEL for this study is 12 mg/ kg/ day and the LOAEL is 18 mg/ kg/ day.

Hayes et al. ( 1986) investigated the subacute ( 14 day) toxicity of DBAN in adult CD rats.

The chemical was administered daily by gavage in corn oil at doses of 0, 23, 45, 90, or 180 mg/ kg/ day ( 10 animals/ sex/ dose). Endpoints assessed included mortality, body weight, organ weight, serum chemistry, hematology, urinalysis, and gross necropsy, as shown in Tables V 2 and V 3. Histopathological studies were not conducted. Also, food and water consumption rates were not measured. There was 100% mortality at 180 mg/ kg/ day. At 90 mg/ kg/ day, 40% of the males and 20% of the females were dead by day 14. No consistent, compound related and dosedependent adverse effects were apparent in any of the serum chemistry, hematological, or urinary parameters measured, although several serum chemistry parameters were significantly different than controls, particularly at the high dose. The authors reported a trend toward higher values for hemoglobin, total red blood cell and white blood cell counts, and fibrinogen in all treated animals ( no data reported). No remarkable findings were observed at necropsy ( gross observation).

Sporadic relative or absolute organ weight changes were reported, mostly at the high dose, but these were not considered as compound related. The only organ weight change that showed a clear dose dependence at multiple doses was for absolute liver weight in females, which was significantly increased 12% over controls ( p

0.05) beginning at 23 mg/ kg/ day, up to 22%

above controls at 90 mg/ kg/ day. Liver weights relative to body weight and brain weight were also affected, but only at higher doses. However, no corresponding significant elevation of serum Drinking Water Criteria Document for Haloacetonitriles EPA/ OW/ OST/ HECD Final Draft V  10 levels of hepatic enzymes was observed in females, and in the absence of histopathology data, it is unclear if the increased liver weight in females is an adverse response. The authors could not identify specific target organs for DBAN toxicity, and concluded that decreased body weight was the most sensitive indicator of toxicity. In males, no decrease in body weight was observed at 23 mg/ kg/ day; body weight was decreased by more than 20% in a dose related manner at 45 and 90 mg/ kg/ day ( p

0.05). No effect on body weight was noted in females. The authors stated that the NOAEL for DBAN was 45 mg/ kg/ day, but the decreased body weight in male rats exposed to 45 mg/ kg/ day suggests that the NOAEL in this study is 23 mg/ kg/ day and the LOAEL is 45 mg/ kg/ day. The data sets for body weight in males and relative liver weight in females were further analyzed to determine benchmark doses ( BMDs) according to draft EPA Guidance ( U. S.

EPA, 2000c) to identify alternative critical effect levels. The results of the modeling are described in detail in Appendix A. A BMDL of 16 mg/ kg/ day for decreased body weight in males was selected as the most appropriate modeling result to serve as the basis for the quantitative doseresponse assessment. Drinking Water Criteria Document for Haloacetonitriles EPA/ OW/ OST/ HECD Final Draft V  11 Table V 2. Organ weights and ratios of CD rats exposed to dibromoacetonitrile by gavage for 14 days. a Parameterb Vehicle ( corn oil)

Male 23 mg/ kg/ day Male 45 mg/ kg/ day Male 90 mg/ kg/ day Male Vehicle ( corn oil)

Female 23 mg/ kg/ day Female 45 mg/ kg/ day Female 90 mg/ kg/ day Female Body Weight 275.5 ± 3.2 274.6 ± 5.5 243.7 ± 4.6* 209.4 ± 10.1* 176.5 ± 6.8 184.9 ± 2.4 181.1 ± 2.8 165.9 ± 5.4 Brain % body weight 1.81 ± 0.05 0.66 ± 0.02 1.77 ± 0.06 0.65 ± 0.03 1.76 ± 0.03 0.73 ± 0.02* 1.66 ± 0.05 0.80 ± 0.06* 1.67 ± 0.04 0.96 ± 0.04 1.68 ± 0.04 0.91 ± 0.03 1.64 ± 0.04 0.91 ± 0.03 1.69 ± 0.06 1.03 ± 0.04 Liver % body weight 12.91 ± 0.34 4.69 ± 0.15 11.60 ± 1.24 4.25 ± 0.45 11.47 ± 1.2 4.71 ± 0.49 10.18 ± 1.89 4.72 ± 0.83 8.59 ± 0.26 4.89 ± 0.12 9.65 ± 0.25*

5.22 ± 0.11 9.73 ± 0.20*

5.39 ± 0.15 10.50 ± 0.44*

6.34 ± 0.24*

Spleen % body weight 0.64 ± 0.03 0.23 ± 0.01 0.59 ± 0.05 0.21 ± 0.02 0.56 ± 0.05 0.23 ± 0.02 0.42 ± 0.05*

0.20 ± 0.02 0.48 ± 0.03 0.27 ± 0.01 0.45 ± 0.02 0.24 ± 0.01 0.52 ± 0.03 0.29 ± 0.02 0.42 ± 0.04 0.26 ± 0.02 Lungs % body weight 1.69 ± 0.10 0.61 ± 0.03 1.49 ± 0.09 0.54 ± 0.04 1.44 ± 0.07 0.59 ± 0.03 1.30 ± 0.14*

0.62 ± 0.06 1.58 ± 0.19 0.88 ± 0.08 1.42 ± 0.09 0.77 ± 0.05 1.25 ± 0.06 0.69 ± 0.04 1.09 ± 0.06*

0.66 ± 0.04*

Thymus % body weight 0.51 ± 0.03 0.19 ± 0.01 0.52 ± 0.04 0.19 ± 0.01 0.46 ± 0.02 0.19 ± 0.01 0.24 ± 0.06*

0.11 ± 0.03* 0.40 ± 0.03 0.23 ± 0.02 0.44 ± 0.03 0.24 ± 0.02 0.40 ± 0.01 0.22 ± 0.01 0.21 ± 0.02*

0.13 ± 0.01*

Kidneys % body weight 2.50 ± 0.10 0.91 ± 0.05 2.39 ± 0.09 0.87 ± 0.03 2.08 ± 0.06*

0.85 ± 0.02 2.13 ± 0.15*

1.02 ± 0.06 1.56 ± 0.06 0.89 ± 0.02 1.64 ± 0.04 0.89 ± 0.02 1.62 ± 0.04 0.89 ± 0.03 1.53 ± 0.05 0.92 ± 0.03 Testes/ Ovaries % body weight 2.70 ± 0.09 0.99 ± 0.04 2.71 ± 0.07 0.99 ± 0.03 2.67 ± 0.06 1.10 ± 0.03 2.69 ± 0.06 1.30 ± 0.07* 0.11 ± 0.01 0.06 ± 0.00 0.14 ± 0.01 0.07 ± 0.01 0.14 ± 0.01 0.08 ± 0.01 0.12 ± 0.02 0.07 ± 0.01 Adapted from Hayes et al. ( 1986).

a All data expressed as mean ± SEM.

b All absolute weights are presented in grams.

* Significantly different from vehicle control ( p

0.05). Drinking Water Criteria Document for Haloacetonitriles EPA/ OW/ OST/ HECD Final Draft V  12 Table V 3. Serum Chemistry Values for CD Rats Exposed to Dibromoacetonitrile ( DBAN) for 14 days. a Parameter Vehicle ( corn oil)

Male 23 mg/ kg/ day Male 45 mg/ kg/ day Male 90 mg/ kg/ day Male Vehicle ( corn oil)

Female 23 mg/ kg/ day Female 45mg/ kg/

day Female 90 mg/ kg/ day Female Serum Glutamate Pyruvate Transaminase ( IU/ L) 168 ± 55 75 ± 9 76 ± 11 77 ± 19 74 ± 6 62 ± 8 74 ± 14 68 ± 11 Serum Glutamate Oxaloacetic Transaminase ( IU/ L) 339 ± 105 142 ± 14 172 ± 26 273 ± 81 178 ± 32 138 ± 9 182 ± 21 161 ± 15 Alkaline Phosphatase ( IU/ L) 360 ± 41 466 ± 33 247 ± 29 187 ± 8* 236 ± 26 299 ± 29 260 ± 24 179 ± 34 5' Nucleotidase ( IU/ L) 18 ± 3 17 ± 1 13 ± 2 14 ± 1 25 ± 4 22 ± 2 28 ± 2 18 ± 4 Protein ( g/ dL) 6.3 ± 0.1 6.5 ± 0.1 6.2 ± 0.1 4.7 ± 0.4* 6.7 ± 0.1 6.2 ± 0.3 6.2 ± 0.1 5.3 ± 0.2*

Albumin ( g/ dL) 4.0 ± 0.1 4.0 ± 0.0 4.0 ± 0.1 2.9 ± 0.3* 4.3 ± 0.2 4.0 ± 0.1 4.2 ± 0.1 3.5 ± 0.1*

Globulin ( g/ dL) 2.3 ± 0.1 2.5 ± 0.1 2.2 ± 0.1 1.8 ± 0.1* 2.5 ± 0.1 2.2 ± 0.2 2.1 ± 0.1 1.8 ± 0.1*

Alb/ globulin ratio 1.8 ± 0.1 1.6 ± 0.1 1.8 ± 0.1 1.6 ± 0.2 1.7 ± 0.1 1.8 ± 0.1 2.1 ± 0.1 2.0 ± 0.1 Glucose ( mg/ dL) 265 ± 47 212 ± 41 217 ± 40 130 ± 1 212 ± 17 171 ± 18 197 ± 22 138 ± 17 Cholesterol ( mg/ dL) 66 ± 2 64 ± 2 74 ± 4 95 ± 5* 66 ± 4 58 ± 6 66 ± 6 71 ± 3 Bilirubin ( mg/ dL) 0.3 ± 0.0 0.4 ± 0.0 0.2 ± 0.0 0.3 ± 0.1 0.3 ± 0.0 0.2 ± 0.0 0.3 ± 0.0 0.2 ± 0.0 BUN ( mg/ dL) 15 ± 1 13 ± 1 10 ± 1 18 ± 6 15 ± 2 13 ± 2 13 ± 1 15 ± 2 Creatinine ( mg/ dL) 1.3 ± 0.2 0.8 ± 0.1 0.9 ± 0.1 0.7 ± 0.0* 1.2 ± 0.1 1.0 ± 0.1 1.0 ± 0.1 0.8 ± 0.0*

BUN/ creatinine ratio 13 ± 2 16 ± 2 12 ± 1* 24 ± 8 12 ± 2 15 ± 4 14 ± 2 19 ± 3 Calcium ( mg/ dL) 11.3 ± 0.4 12.5 ± 0.6 11.8 ± 0.4 9.6 ± 0.6 11.9 ± 0.4 11.1 ± 0.3 12.1 ± 0.5 11.1 ± 0.3 Phosphorus ( mg/ dL) 12.0 ± 0.4 11.2 ± 0.4 11.3 ± 0.5 7.6 ± 0.7* 11.6 ± 0.1 10.1 ± 0.6 10.8 ± 0.5 9.3 ± 0.7*

Chloride ( mEq/ L) 99 ± 2 99 ± 2 99 ± 1 100 ± 2 98 ± 1.4 102 ± 0.7 100 ± 1.2 101 ± 0.7 Adapted from Hayes et al. ( 1986).

a All data expressed as mean ± SEM.

* Significantly different from vehicle control ( p

0.05). Drinking Water Criteria Document for Haloacetonitriles EPA/ OW/ OST/ HECD Final Draft V  13 Table V 4. Body Organ Weights and Ratios of CD rats Exposed to Dichloroacetonitrile by Gavage for 14 days. a Parameterb Vehicle ( corn oil)

Male 12 mg/ kg/ day Male 23 mg/ kg/ day Male 45 mg/ kg/ day Male 90 mg/ kg/ day Male Vehicle ( corn oil)

Female 12 mg/ kg/ day Female 23 mg/ kg/ day Female 45 mg/ kg/ day Female 90 mg/ kg/ day Female Body Weight 157 ± 5 170 ± 7 147 ± 5 137 ± 5 115 ± 4 148 ± 6 147 ± 6 143 ± 4 146 ± 5 113 ± 5 Brain % body weight 1.68 ± 0.09 1.06 ± 0.03 1.63 ± 0.12 0.99 ± 0.09 1.66 ± 0.03 1.14 ± 0.04 1.56 ± 0.04 1.15 ± 0.04 1.52 ± 0.05*

1.34 ± 0.06* 1.61 ± 0.04 1.09 ± 0.04 1.52 ± 0.08 1.04 ± 0.04 1.47 ± 0.06 1.03 ± 0.03 1.61 ± 0.05 1.11 ± 0.07 1.45 ± 0.04 1.30 ± 0.06*

Liver % body weight 8.15 ± 0.31 5.19 ± 0.10 9.96 ± 0.58 5.85 ± 0.16* 9.71 ± 0.64 6.57 ± 0.30* 10.11 ± 0.52*

7.37 ± 0.18* 8.63 ± 0.48 7.50 ± 0.26* 8.06 ± 0.53 5.40 ± .20 8.49 ± 0.35 5.81 ± 0.15 10.56 ± 0.50*

7.37 ± 0.27* 11.14 ± 0.57*

7.59 ± 0.23* 7.78 ± 0.92 7.07 ± 0.82*

Spleen % body weight 0.54 ± 0.04 0.34 ± 0.02 0.62 ± 0.07 0.36 ± 0.04 0.46 ± 0.03 0.31 ± 0.01 0.47 ± 0.05 0.34 ± 0.03 0.35 ± 0.03*

0.30 ± 0.02 0.44 ± 0.03 0.30 ± 0.02 0.43 ± 0.04 0.30 ± 0.03 0.40 ± 0.02 0.28 ± 0.02 0.43 ± 0.02 0.29 ± 0.02 0.29 ± 0.02*

0.26 ± 0.01 Lungs % body weight 1.27 ± 0.07 0.81 ± 0.04 1.37 ± 0.11 0.82 ± 0.07 1.26 ± 0.08 0.85 ± 0.05 1.17 ± 0.07 0.86 ± 0.05 1.02 ± 0.09 0.89 ± 0.07 1.35 ± 0.09 0.91 ± 0.06 1.17 ± 0.09 0.80 ± 0.05 1.13 ± 0.07 0.79 ± 0.05 1.23 ± 0.10 0.84 ± 0.06 0.96 ± 0.07*

0.86 ± 0.07 Thymus % body weight 0.43 ± 0.04 0.27 ± 0.02 0.55 ± 0.05 0.32 ± 0.03 0.38 ± 0.02*

0.26 ± 0.02 0.33 ± 0.02*

0.24 ± 0.02 0.27 ± 0.02*

0.24 ± 0.02* 0.48 ± 0.03 0.33 ± 0.02 0.42 ± 0.04 0.29 ± 0.03 0.38 ± 0.03 0.27 ± 0.02 0.42 ± 0.02 0.29 ± 0.02 0.32 ± 0.02*

0.29 ± 0.02 Kidneys % body weight 1.53 ± 0.05 0.98 ± 0.03 1.85 ± 0.08*

1.10 ± 0.04* 1.52 ± 0.05 1.03 ± 0.02 1.48 ± 0.06 1.08 ± 0.02 1.38 ± 0.07 1.20 ± 0.04* 1.52 ± 0.05 1.03 ± 0.03 1.42 ± 0.06 0.97 ± 0.03 1.47 ± 0.04 1.03 ± 0.03 1.51 ± 0.06 1.03 ± 0.02 1.28 ± 0.05*

1.14 ± 0.04 Testes/ Ovaries % body weight 1.54 ± 0.13 0.97 ± 0.07 1.74 ± 0.15 1.00 ± 0.06 1.54 ± 0.12 1.03 ± 0.06 1.26 ± 0.13 0.90 ± 0.09 1.25 ± 0.14 1.07 ± 0.11 0.11 ± 0.01 0.08 ± 0.01 0.10 ± 0.01 0.07 ± 0.01 0.10 ± 0.01 0.07 ± 0.01 0.10 ± 0.01 0.07 ± 0.01 0.08 ± 0.01 0.07 ± 0.01 Adapted from Hayes et al. ( 1986).

a All data expressed as mean ± SEM.

b All absolute weights are presented in grams.

* Significantly different from vehicle control ( p

0.05). Drinking Water Criteria Document for Haloacetonitriles EPA/ OW/ OST/ HECD Final Draft V  14 Table V 5. Serum Chemistry Values for CD Rats Exposed to Dichloroacetonitrile ( DCAN) for 14 daysa.

Parameter Vehicle ( corn oil)

Male 12 mg/ kg/ day Male 23 mg/ kg/ day Male 45 mg/ kg/ day Male 90 mg/ kg/ day Male Vehicle ( corn oil)

Female 12 mg/ kg/ day Female 23 mg/ kg/ day Female 45 mg/ kg/ day Female 90 mg/ kg/ day Female Serum Glutamate Pyruvate Transaminase ( IU/ L) 70 ± 3 71 ± 6 89 ± 22 87 ± 15 88 ± 14 57 ± 6 66 ± 7 63 ± 4 63 ± 7 134 ± 38*

Serum Glutamate Oxaloacetic Transaminase ( IU/ L) 167 ± 13 266 ± 22 213 ± 53 276 ± 40 233 ± 27 205 ± 8 198 ± 6 201 ± 33 181 ± 28 340 ± 96 Alkaline Phosphatase ( IU/ L) 457 ± 63 459 ± 39 384 ± 21 398 ± 50 712 ± 84* 261 ± 24 381 ± 48 352 ± 45 384 ± 35* 651 ± 159*

5' Nucleotidase ( IU/ L) 14 ± 1 16 ± 1 14 ± 2 17 ± 2 23 ± 2* 23 ± 3 20 ± 2 16 ± 1 18 ± 1 24 ± 2 Protein ( g/ dL) 5.8 ± 0.2 6.0 ± 0.2 5.7 ± 0.2 5.5 ± 0.2 6.0 ± 0.1 6.4 ± 0.3 6.2 ± 0.2 6.4 ± 0.3 5.9 ± 0.2 6.1 ± 0.4 Albumin( g/ dL) 4.4 ± 0.1 4.3 ± 0.1 4.3 ± 0.1 4.3 ± 0.1 4.8 ± 0.1 4.6 ± 0.1 4.4 ± 0.1 4.5 ± 0.1 4.7 ± 0.1 4.8 ± 0.2 Globulin ( g/ dL) 1.4 ± 0.1 1.7 ± 0.1 1.4 ± 0.2 1.2 ± 0.2 1.2 ± 0.2 1.8 ± 0.2 1.9 ± 0.2 1.9 ± 0.2 1.2 ± 0.1 1.3 ± 0.3 Alb/ globulin ratio 3.1 ± 0.1 2.6 ± 0.1* 3.4 ± 0.5 3.8 ± 0.5 4.5 ± 0.7 2.8 ± 0.4 2.4 ± 0.2 2.4 ± 0.2 3.9 ± 0.3 4.0 ± 0.7 Glucose ( mg/ dL) 154 ± 11 129 ± 16 138 ± 6 134 ± 10 112 ± 16 191 ± 19 162 ± 30 142 ± 17 160 ± 6 137 ± 15 Cholesterol ( mg/ dL) 90 ± 5 93 ± 2 91 ± 6 80 ± 10 81 ± 11 84 ± 3 86 ± 4 99 ± 6 82 ± 8 73 ± 8 Bilirubin ( mg/ dL) 0.2 ± 0.0 0.2 ± 0.0 0.3 ± 0.0 0.4 ± 0.0* 0.6 ± 0.1 0.2 ± 0.0 0.2 ± 0.0 0.2 ± 0.0 0.4 ± 0.0* 0.7 ± 0.1*

BUN ( mg/ dL) 17 ± 2 17 ± 2 15 ± 2 17 ± 2 15 ± 1 17 ± 1 17 ± 1 15 ± 2 20 ± 3 23 ± 2 Creatinine ( mg/ dL) 1.7 ± 0.2 1.4 ± 0.2 1.4 ± 0.1 1.5 ± 0.1 1.4 ± 0.1 1.6 ± 0.2 1.3 ± 0.1 1.2 ± 0.1 1.6 ± 0.1 1.6 ± 0.2 BUN/ creatinine ratio 11 ± 3 12 ± 1 11 ± 1 12 ± 2 11 ± 1 11 ± 1 13 ± 1 13 ± 2 12 ± 1 14 ± 1 Calcium ( mg/ dL) 11.4 ± 0.1 11.0 ± 0.2 11.3 ± 0.2 9.8 ± 1.5 11.7 ± 0.3 11.7 ± 0.5 11.1 ± 0.3 11.3 ± 0.2 11.9 ± 0.3 11.8 ± 0.5 Phosphorus ( mg/ dL) 12.3 ± 0.6 12.0 ± 0.8 12.5 ± 0.6 11.7 ± 0.5 12.9 ± 0.4 11.4 ± 0.9 10.9 ± 0.5 11.4 ± 0.9 12.7 ± 0.1 9.9 ± 0.4 Chloride( mEq/ L) 100 ± 1 100 ± 1 101 ± 1 101 ± 2 101 ± 1 100 ± 1 99 ± 1 100 ± 1 101 ± 1 101 ± 2 Adapted from Hayes et al. ( 1986). aAll data expressed as mean ± SEM * Significantly different from vehicle control at ( p

0.05) Drinking Water Criteria Document for Haloacetonitriles V  15 EPA/ OW/ OST/ HECD Final Draft In a similar 14 day repeated dosing study with DCAN, Hayes et al. ( 1986) administered gavage doses of 0, 12, 23, 45, or 90 mg/ kg/ day in corn oil to CD rats ( 10 animals/ sex/ dose).

Food and water consumption rates were not measured. No mortality was reported for any treatment group. In males, depression in body weight to 94%, 87%, and 73% of body weight in control animals was observed at 23, 45, and 90 mg/ kg/ day, respectively, as shown in Table V 4.

In females, body weight was decreased to 76% of controls at 90 mg/ kg/ day. Several serum markers for organ toxicity were increased in treated animals as shown in Table V 5. Significantly increased levels of serum glutamic pyruvate transaminase ( SGPT) in females at 90 mg/ kg/ day, and alkaline phosphatase ( ALP) levels at 90 mg/ kg/ day in males and at 45 and 90 mg/ kg/ day in females were reported. Although the authors did not consider these changes to be compoundrelated adverse effects, they did not provide a reason. We considered these changes to be adverse, based on the magnitude of the change, and the supporting data for DCAN in females in the subchronic study. No other consistent dose dependent adverse effects were observed in any of the other serum chemistry, hematological, or urinary parameters measured. The authors reported a trend toward elevated red blood cell and white blood cell counts in all treated animals,

but no data were provided. No remarkable findings were observed at necropsy ( gross observation). Relative liver weight was significantly increased ( p

0.05) in male and female rats.

The relative liver weights in males were 13%, 26%, 42%, and 45% greater than controls at doses of 12, 23, 45, and 90 mg/ kg/ day, respectively. Absolute liver weight, in contrast, was significantly increased ( p

0.05) only in the 45 mg/ kg/ day dose group. In female rats, both relative and absolute liver weights were elevated beginning at 23 mg/ kg/ day, with relative liver weights 36%, 40%, and 31% greater than controls at 23, 45, and 90 mg/ kg/ day, respectively. Drinking Water Criteria Document for Haloacetonitriles V  16 EPA/ OW/ OST/ HECD Final Draft Liver weight changes would be considered as potentially adaptive in the absence of other signs of hepatic injury. However, the observed increase in serum levels of hepatic enzyme activity at higher doses than those associated with liver weight gives greater weight to the potential toxicological significance of the liver weight changes, even though the absence of histopathology data makes it difficult to determine conclusively if the effects were adverse at low doses. Based on this uncertainty, both decreased body weight and increased relative liver weight are considered toxicologically relevant responses. The more sensitive of these endpoints was selected as the critical effect. Therefore, the lowest dose tested of 12 mg/ kg/ day is the study LOAEL for increased relative liver weight in males, and no NOAEL is determined. The data sets for body weight and relative liver weight in males and females were further analyzed to determine benchmark doses ( BMDs) according to draft EPA Guidance ( U. S. EPA, 2000c) to identify alternative critical effect levels. The results of the modeling are described in detail in Appendix A.

A BMDL of 5 mg/ kg/ day for relative liver weight in males was selected as the most appropriate modeling result to serve as the basis for the quantitative dose response assessment.

B. Long Term Exposures No studies were identified that evaluated the toxicity of long term exposure to BCAN or TCAN by any route. No studies were identified that evaluated the toxicity of long term exposure to DBAN or DCAN by the inhalation or dermal routes.

The subchronic toxicity of DBAN has been evaluated by the NTP ( 2002) in B6C3F1 mice and F344 rats as part of initial dose range finding studies for chronic exposure studies that are Drinking Water Criteria Document for Haloacetonitriles V  17 EPA/ OW/ OST/ HECD Final Draft currently in progress. For the mouse study, DBAN was administered in drinking water for 13 weeks to male and female B6C3F1 mice ( 10/ sex/ dose) at concentrations of 0, 12.5, 25, 50, 100,

and 200 mg/ L. The corresponding doses reported by the study authors were 0, 1.6, 3.2, 5.6,

10.7, and 17.9 mg/ kg/ day for males and 0, 1.6, 3.0, 6.1, 11.1, and 17.9 mg/ kg/ day for females.

Animals were observed for clinical signs of toxicity, as well as body weight, organ weight and pathology, hematology, and clinical chemistry. A separate set of animals ( 10/ sex/ dose) were exposed to the same concentrations as the main study groups for 26 days, but were co exposed to DBAN and 5 bromo 2 deoxyuridine during the last five days of this period. These animals were used to collect tissue sample for analysis of induce cell proliferation. Decreased water consumption and decreased body weight were the only effects related to DBAN treatment.

Decreased water consumption was observed in both males and females at DBAN concentrations of 50 mg/ L and higher. A slight and transient decrease in body weight gain was observed;

terminal body weights were 94% of controls in high dose males and 96% of controls in high dose females. These small changes are not judged as toxicologically significant. Based on the minimal effects observed for DBAN in this study, the NOAEL is 17.9 mg/ kg/ day for males and females.

No LOAEL was identified.

For the rat study, DBAN was administered in drinking water for 13 weeks to male and female F344 rats ( 10/ sex/ dose) at concentrations of 0, 12.5, 25, 50, 100, and 200 mg/ L. The corresponding doses reported by the study authors were 0, 0.9, 1.8, 3.3, 6.2, and 11.3 mg/ kg/ day for males and 0, 1.0, 1.9, 3.8, 6.8, and 12.6 mg/ kg/ day for females. Animals were observed for clinical signs of toxicity, as well as body weight, organ weight and pathology, hematology, and Drinking Water Criteria Document for Haloacetonitriles V  18 EPA/ OW/ OST/ HECD Final Draft clinical chemistry. A separate set of animals ( 10/ sex/ dose) were exposed to the same concentrations as the main study groups for 26 days, but were co exposed to DBAN and 5

bromo 2 deoxyuridine during the last five days of this period. These animals were used to collect tissue samples for analysis of cell proliferation. Decreased water consumption and decreased body weight were the only effects related to DBAN treatment. Slight changes in clinical chemistry and hematology findings were considered by the study authors to be related to decreased water consumption. Decreased water consumption was observed in males at DBAN concentrations of 50 mg/ L and higher and in females at the two highest concentrations. A slight decrease in body weight gain was observed for high dose males and females. Terminal body weights were 94% of controls in high dose males and 95% of controls in high dose females. These small changes are not judged as toxicologically significant. Based on the minimal effects observed for DBAN in this study, the NOAEL is 11.3 mg/ kg/ day for males and 12.6 mg/ kg/ day for females. No LOAEL was identified.

In a 90 day study of DBAN toxicity, Hayes et al. ( 1986) administered doses of 0, 6, 23,

or 45 mg/ kg/ day in corn oil by gavage to groups of CD rats ( 20 animals/ sex/ dose). No compound related deaths occurred during this study. Body weights were depressed to 79% of controls at the end of the study in males, but not in females, at the highest dose tested ( 45 mg/ kg/

day). Food and water consumption rates were not measured. Observed changes in the serum chemistry, hematological, urinary parameters, and organ weight were generally not dose related and were not considered to be compound related ( Tables V 6 and V 7). The only exceptions were significantly increased ALP in females at 45 mg/ kg/ day and a significant increase in relative Drinking Water Criteria Document for Haloacetonitriles V  19 EPA/ OW/ OST/ HECD Final Draft liver weight ( but not absolute weight) in males at 45 mg/ kg/ day. Interim serum biochemistry analyses at one and two months of exposure also revealed no treatment related effects. No remarkable findings were apparent at gross necropsy and the authors did not identify a specific target tissue for DBAN. The finding that effects on the liver in this subchronic study were comparable to the effects following the 14 day exposure at the same doses suggests that the liver weight changes seen in the 14 day study were adaptive, or at least that tolerance developed to the repeated dosing. In addition, no clear increase in serum enzyme markers of hepatic injury were observed in males, even though they had a greater increase in relative liver weight than females.

Based on this discordance between serum biochemistry and liver weight findings, and in light of the results from the 14 day study, the observed liver weight changes were not judged sufficiently adverse to serve as the basis for the dose response assessment. Based on decreased body weight in males as the most sensitive endpoint, the NOAEL for this study is 23 mg/ kg/ day and the LOAEL is 45 mg/ kg/ day.

The body weight data in males were further analyzed to determine benchmark doses ( BMDs) according to draft EPA Guidance ( U. S. EPA, 2000c) to identify alternative critical effect levels. The results of the modeling are described in detail in Appendix A. A BMDL of 20 mg/ kg/ day for decreased body weight in males was selected as the most appropriate modeling result to serve as the basis for the quantitative dose response assessment. Drinking Water Criteria Document for Haloacetonitriles V  20 EPA/ OW/ OST/ HECD Final Draft Table V 6. Body and Organ Weights for CD Rats Exposed to Dibromoacetonitrile ( DBAN) by Gavage for 90 Daysa.

Parameterb Vehicle ( corn oil) Male 6 mg/ kg/ day Male 23 mg/ kg day Male 45 mg/ kg/ day Male Vehicle ( corn oil) Female 6 mg/ kg/ day Female 23 mg/ kg/ day Female 45 mg/ kg/ day Female Body Weight 556.0 ± 12.7 545.9 ± 13.5 523.1 ± 11.4 438.6 ± 7.5* 292.9 ± 5.8 279.6 ± 7.3 291.7 ± 9.5 274.9 ± 7.2 Brain % body weight 1.96 ± 0.04 0.36 ± 0.01 1.98 ± 0.05 0.37 ± 0.01 1.95 ± 0.04 0.38 ± 0.01 1.93 ± 0.4 0.44 ± 0.01* 1.83 ± 0.03 0.63 ± 0.02 1.81 ± 0.04 0.65 ± 0.02 1.76 ± 0.04 0.61 ± 0.01 1.76 ± 0.04 0.64 ± 0.02 Liver % body weight 21.85 ± 0.99 3.93 ± 0.15 21.70 ± 0.68 3.98 ± 0.08 22.39 ± 0.84 4.27 ± 0.12 19.7 ± 0.67 4.44 ± 0.13* 11.62 ± 0.37 3.98 ± 0.12 11.1 ± 0.36 3.98 ± 0.10 12.31 ± 0.33 4.25 ± 0.11 11.95 ± 0.39 4.36 ± 0.12 Spleen % body weight 0.84 ± 0.04 0.15 ± 0.01 0.84 ± 0.03 0.15 ± 0.01 0.78 ± 0.04 0.15 ± 0.01 0.72 ± 0.04 0.16 ± 0.01 0.51 ± 0.03 0.17 ± 0.01 0.52 ± 0.02 0.19 ± 0.01 0.55 ± 0.03 0.19 ± 0.01 0.57 ± 0.04 0.21 ± 0.01 Lungs % body weight 3.22 ± 0.14 0.58 ± 0.03 2.95 ± 0.09 0.54 ± 0.02 3.09 ± 0.11 0.59 ± 0.02 2.73 ± 0.09* 0.63 ± 0.03 2.27 ± 0.15 0.78 ± 0.05 2.34 ± 0.16 0.85 ± 0.07 2.35 ± 0.14 0.81 ± 0.05 2.47 ± 0.17 0.90 ± 0.06 Thymus % body weight 0.55 ± 0.03 0.10 ± 0.01 0.50 ± 0.03 0.09 ± 0.01 0.53 ± 0.02 0.10 ± 0.00 0.43 ± 0.03* 0.10 ± 0.01 0.40 ± 0.03 0.14 ± 0.01 0.40 ± 0.03 0.15 ± 0.01 0.34 ± 0.02 0.12 ± 0.01 0.26 ± 0.03* 1.10 ± 0.01*

Kidneys % body weight 3.67 ± 0.10 0.66 ± 0.02 3.56 ± 0.09 0.66 ± 0.02 3.62 ± 0.10 0.69 ± 0.01 3.16 ± 0.10* 0.72 ± 0.02* 2.10 ± 0.05 0.72 ± 0.02 1.98 ± 0.06 0.71 ± 0.02 2.01 ± 0.70 0.70 ± 0.02 1.98 ± 0.07 0.72 ± 0.02 Testes/ Ovaries % body weight 3.58 ± 0.05 0.65 ± 0.01 3.76 ± 0.07 0.69 ± 0.02 3.49 ± 0.60 0.67 ± 0.02 3.51 ± 0.15 0.80 ± 0.03* 0.17 ± 0.01 0.06 ± 0.00 0.17 ± 0.01 0.06 ± 0.00 0.16 ± 0.01 0.05 ± 0.00 0.17 ± 0.01 0.06 ± 0.60 Adapted from Hayes et al. ( 1986).

aAll data expressed as mean ± SEM bAll absolute weights are presented in grams. * Significantly different from vehicle control ( p

0.05). Drinking Water Criteria Document for Haloacetonitriles EPA/ OW/ OST/ HECD Final Draft V  21 Table V 7. Serum Chemistry Values for CD Rats Exposed to Dibromoacetonitrile ( DBAN) by Gavage for 90 days. a Parameter Vehicle ( corn oil)

Male 6 mg/ kg/ day Male 23 mg/ kg/ day Male 45 mg/ kg/ day Male Vehicle ( corn oil)

Female 6 mg/ kg/ day Female 23 mg/ kg/ day Female 45 mg/ kg/ day Female Serum Glutamate Pyruvate Transaminase ( IU/ L) 44 ± 4 69 ± 3 42 ± 4 44 ± 5 46 ± 3 35 ± 3 37 ± 3 32 ± 3*

Serum Glutamate Oxaloacetic Transaminase ( IU/ L) 176 ± 12 245 ± 52 190 ± 30 181 ± 13 207 ± 31 169 ± 31 174 ± 10 161 ± 11 Alkaline Phosphatase ( IU/ L) 191 ± 16 171 ± 28 192 ± 23 157 ± 16 134 ± 11 130 ± 9 138 ± 19 208 ± 32*

5' Nucleotidase ( IU/ L) 15 ± 1 17 ± 2 16 ± 1 13 ± 1 26 ± 2 26 ± 2 24 ± 2 19 ± 1*

Protein ( g/ dL) 8.2 ± 0.4 8.2 ± 0.3 8.1 ± 0.2 7.0 ± 0.2* 7.3 ± 0.2 7.5 ± 0.2 7.1 ± 0.2 6.9 ± 0.2 Albumin( g/ dL) 5.2 ± 0.1 5.3 ± 0.2 5.8 ± 0.1* 5.8 ± 0.1* 6.1 ± 0.1 6.6 ± 0.1* 6.1 ± 0.1 5.6 ± 0.1*

Globulin ( g/ dL) 2.9 ± 0.5 2.9 ± 0.4 2.3 ± 0.2 1.2 ± 0.1* 1.2 ± 0.1 0.9 ± 0.2 1.1 ± 0.1 1.3 ± 0.2 Alb/ globulin ratio 2.5 ± 0.6 2.3 ± 0.3 2.8 ± 0.4 5.8 ± 0.8 5.7 ± 0.6 9.0 ± 1.3* 6.3 ± 0.8 5.3 ± 0.8 Glucose ( mg/ dL) 148 ± 7 139 ± 7 143 ± 8 134 ± 5 145 ± 8 147 ± 5 135 ± 10 130 ± 7 Cholesterol ( mg/ dL) 73 ± 5 70 ± 5 69 ± 4 77 ± 5 80 ± 4 80 ± 6 78 ± 4 64 ± 3*

Bilirubin ( mg/ dL) 0.6 ± 0.1 0.9 ± 0.1 0.8 ± 0.1 0.7 ± 0.0 0.5 ± 0.0 0.5 ± 0.0 0.4 ± 0.1 0.7 ± 0.0*

BUN ( mg/ dL) 17 ± 1 15 ± 1 17 ± 1 18 ± 1 18 ± 1 18 ± 1 18 ± 1 15 ± 1 Creatinine ( mg/ dL) 1.0 ± 0.1 0.9 ± 0.1 1.3 ± 0.0 1.0 ± 0.0 1.2 ± 0.0 1.3 ± 0.1 1.0 ± 0.1 1.0 ± 0.0*

BUN/ creatinine ratio 17 ± 1 17 ± 2 14 ± 0 19 ± 1 15 ± 1 14 ± 1 18 ± 2 15 ± 1 Calcium ( mg/ dL) 11.3 ± 0.5 11.7 ± 0.2 12.8 ± 0.2* 10.9 ± 0.2 10.9 ± 0.2 11.4 ± 0.2 11.1 ± 0.2 10.7 ± 0.3 Phosphorus ( mg/ dL) 6.4 ± 0.4 6.9 ± 0.4 6.7 ± 0.2 6.5 ± 0.2 4.6 ± 0.2 4.8 ± 0.1 5.7 ± 0.2* 6.2 ± 0.3*

Chloride ( mEq/ L) 102 ± 1 101 ± 1 102 ± 1 102 ± 1 102 ± 1 101 ± 1 102 ± 1 101 ± 1 Adapted from Hayes et al. ( 1986).

a All data expressed as mean SEM * Significantly different from vehicle control ( p

0.05). Drinking Water Criteria Document for Haloacetonitriles V  22 EPA/ OW/ OST/ HECD Final Draft In a 90 day study of DCAN toxicity, Hayes et al. ( 1986) administered doses of 0, 8, 33,

or 65 mg/ kg/ day by gavage in corn oil to groups of CD rats ( 20 animals/ sex/ dose). Food and water consumption data were not reported. At 65mg/ kg/ day, 50% of males and 25% of females had died by the completion of the study; at 33 mg/ kg/ day, 10% of males and 5% of females had died; and at 8 mg/ kg/ day, 5% of males had died. Supplementary information provided by the authors indicated that there were one to three deaths ( 5% to 15%) in the control groups, and that some of the deaths in the high dose groups were due to gavage error. Most of the deaths that were judged to be compound related occurred in weeks 9 to 10. Body weight was significantly depressed in male and female rats at 65 mg/ kg/ day ( to 73% of controls) and in males at 33 mg/ kg/ day ( to 81% of controls). Most of the observed changes in the serum chemistry,

hematological, and urinary parameters did not appear to be compound related ( Tables V 8 and V

9). The exception was alkaline phosphatase, which was significantly increased in males and females at the high dose, and in males also at 33 mg/ kg/ day. Sporadic organ weight changes were observed, mostly at the high dose. Of these, a dose dependent increase was seen only for relative liver weights. Relative liver weight ( relative to body weight) was significantly increased ( p

0.05)

in males beginning at 33 mg/ kg/ day ( 60% increase), and in females beginning at 8 mg/ kg/ day ( 17% increase). The relative liver weight was also increased in males ( by 12%) at 8 mg/ kg/ day.

Although the 12% increase in relative liver weight in males was not statistically significant, it is judged to be biologically significant, based on the magnitude of the change and the observation in the 14 day study that males were more sensitive to liver weight changes than females. Liver weight changes would be considered as potentially adaptive in the absence of other signs of hepatic injury. The observed increase in serum levels of ALP activity gives greater weight to the Drinking Water Criteria Document for Haloacetonitriles V  23 EPA/ OW/ OST/ HECD Final Draft potential toxicological significance of the liver weight changes, although ALP is not a liverspecific enzyme, and no corresponding increases in the liver specific enzymes SGPT or serum glutamate oxaloacetic transaminase ( SGOT) were observed at study termination in the subchronic study. However, the toxicological relevance of the ALP results following subchronic dosing is supported by the increase in both ALP and SGPT observed in the 14 day study. The absence of histopathology data makes it difficult to determine conclusively if the effects were adverse at low doses. Based on this uncertainty, both decreased body weight and increased relative liver weight are considered toxicologically relevant responses. The more sensitive of these endpoints was selected as the critical effect. Therefore, the lowest dose tested of 8 mg/ kg/ day is the study LOAEL for increased relative liver weight in males and females, and no NOAEL is determined.

The body weight and relative liver weight data in males and females were further analyzed to determine benchmark doses ( BMDs) for these endpoints according to draft EPA Guidance ( U. S. EPA, 2000c) to identify alternative critical effect levels. The results of the modeling are described in detail in Appendix A. A BMDL of 4 mg/ kg/ day for increased relative liver weight in males was selected as the most appropriate modeling result to serve as the basis for the quantitative dose response assessment. Drinking Water Criteria Document for Haloacetonitriles EPA/ OW/ OST/ HECD Final Draft V  24 Table V 8. Body and Organ Weights for CD Rats Exposed to Dichloroacetonitrile ( DCAN) by Gavage for 90 daysa.

Parameterb Vehicle ( corn oil)

Male 8 mg/ kg/ day Male 33 mg/ kg/ day Male 65 mg/ kg/ day Male Vehicle ( corn oil)

Female 8 mg/ kg/ day Female 33 mg/ kg day Female 65 mg/ kg/ day Female Body Weight 541.4 ± 13.1 572.1 ± 14.5 438.0 ± 16.6* 285.6 ± 20.6* 309.0 ± 6.91 282.5 ± 17.4 280.6 ± 10.7 225.5 ± 8.9*

Brain % body weight 1.86 ± 0.06 0.35 ± 0.01 1.84 ± 0.07 0.33 ± 0.02 1.86 ± 0.06 0.44 ± 0.02* 1.77 ± 0.06 0.63 ± 0.03* 1.68 ± 0.07 0.55 ± 0.02 1.69 ± 0.07 0.57 ± 0.02 1.58 ± 0.06 0.58 ± 0.03 1.46 ± 0.08 0.66 ± 0.05 Liver % body weight 21.43 ± 0.73 4.0 ± 0.10 25.83 ± 0.104 4.5 ± 0.13 27.26 ± 1.06 6.4 ± 0.43* 17.91 ± 1.60 6.4 ± 0.92* 12.10 ± 0.42 4.0 ± 0.14 14.13 ± 0.87 4.7 ± 0.29* 16.83 ± 0.58*

6.1 ± 0.22* 13.88 ± 0.705 6.1 ± 0.15*

Spleen % body weight 0.72 ± 0.03 0.13 ± 0.03 0.73 ± 0.03 0.13 ± 0.02 0.63 ± 0.03 0.15 ± 0.04 0.50 ± 0.06*

0.18 ± 0.03 0.54 ± 0.02 0.18 ± 0.01 0.56 ± 0.03 0.19 ± 0.01 0.50 ± 0.02 0.18 ± 0.01 0.450 ± 0.020 0.20 ± 0.01 Lungs % body weight 2.98 ± 0.13 0.56 ± 0.02 2.77 ± 0.13 0.49 ± 0.02* 2.63 ± 0.10 0.61 ± 0.02 1.83 ± 0.06*

0.65 ± 0.04 2.04 ± 0.10 0.67 ± 0.03 2.20 ± 0.15 0.74 ± 0.05 2.22 ± 0.16 0.79 ± 0.03* 1.83 ± 0.10 0.82 ± 0.05*

Thymus % body weight 0.60 ± 0.03 0.12 ± 0.01 0.79 ± 0.04*

0.14 ± 0.01 0.62 ± 0.03 0.15 ± 0.01* 0.31 ± 0.03*

0.11 ± 0.02 0.50 ± 0.03 0.16 ± 0.01 0.54 ± 0.03 0.18 ± 0.01 0.42 ± 0.02*

0.15 ± 0.01 0.375 ± 0.02 0.17 ± 0.01 Kidneys % body weight 3.71 ± 0.07 0.69 ± 0.01 3.88 ± 0.12 0.68 ± 0.02 3.45 ± 0.12 0.81 ± 0.04* 2.77 ± 0.19*

1.0 ± 0.12* 2.23 ± 0.06 0.72 ± 0.01 2.38 ± 0.06 0.80 ± 0.02 2.20 ± 0.08 0.80 ± 0.03 2.25 ± 0.12 1.00 ± 0.02*

Testes/ Ovaries % body weight 3.42 ± 0.08 0.63 ± 0.02 3.66 ± 0.06 0.65 ± 0.02 3.40 ± 0.07 0.80 ± 0.04* 2.91 ± 0.15*

1.0 ± 0.12* 0.16 ± 0.01 0.05 ± 0.004 0.14 ± 0.01 0.05 ± 0.002 0.14 ± 0.01 0.05 ± 0.003 0.10 ± 0.02*

0.05 ± 0.01 Adapted from Hayes et al. ( 1986).

All data expressed as mean ± SEM.

b All absolute weights are presented in grams.

* Significantly different from vehicle control ( p

0.05). Drinking Water Criteria Document for Haloacetonitriles EPA/ OW/ OST/ HECD Final Draft V  25 Table V 9. Serum Chemistry Values for CD Rats Exposed to Dichloroacetonitrile ( DCAN) for 90 days. a Parameter Vehicle ( corn oil)

Male 8 mg/ kg/ day Male 33 mg/ kg/ day Male 65 mg/ kg/ day Male Vehicle ( corn oil)

Female 8 mg/ kg/ day Female 33 mg/ kg/ day Female 65 mg/ kg/ day Female Serum Glutamate Pyruvate Transaminase ( IU/ L) 53 ± 1 45 ± 2 116 ± 60 53 ± 4 51 ± 3 32 ± 2* 27 ± 2* 35 ± 3*

Serum Glutamate Oxaloacetic Transaminase ( IU/ L) 195 ± 28 140 ± 13 265 ± 96 163 ± 18 120 ± 14 127 ± 15 124 ± 14 126 ± 16 Alkaline Phosphatase ( IU/ L) 222 ± 20 320 ± 31 471 ± 46* 603 ± 79* 228 ± 29 226 ± 25 286 ± 30 499 ± 46*

5' Nucleotidase ( IU/ L) 19 ± 2 14 ± 1 22 ± 3 21 ± 1 30 ± 2 19 ± 1* 16 ± 1* 17 ± 1*

Protein ( g/ dL) 7.5 ± 0.1 7.2 ± 0.1 7.0 ± 0.1* 6.5 ± 0.1 7.1 ± 0.1 7.3 ± 0.1 7.1 ± 0.1 6.6 ± 0.1*

Albumin ( g/ dL) 5.6 ± 0.1 6.0 ± 0.1* 5.6 ± 0.1* 5.4 ± 0.1 6.2 ± 0.2 6.0 ± 0.1 6.0 ± 0.1 5.5 ± 0.1*

Globulin ( g/ dL) 1.9 ± 0.1 1.2 ± 0.1 1.4 ± 0.1 1.0 ± 0.1 0.9 ± 0.1 1.3 ± 0.1 1.1 ± 0.1 1.1 ± 0.1 Alb/ globulin ratio 3.1 ± 0.4 5.3 ± 0.4 4.0 ± 0.4 5.0 ± 1 7.0 ± 1 5.1 ± 0.4 6.0 ± 0.4 5.0 ± 1 Glucose ( mg/ dL) 145 ± 4 143 ± 3 147 ± 4 109 ± 4* 150 ± 6 127 ± 5* 140 ± 5 120 ± 6*

Cholesterol ( mg/ dL) 81 ± 6 92 ± 6 71 ± 5 55 ± 4* 73 ± 3 96 ± 6* 80 ± 7 54 ± 4*

Bilirubin ( mg/ dL) 0.5 ± 0.0 0.5 ± 0.0 0.5 ± 0.1 0.6 ± 0.1 0.5 ± 0.0 0.4 ± 0.0 0.4 ± 0.0 0.5 ± 0.0 BUN ( mg/ dL) 14 ± 1 13 ± 1 14 ± 1 16 ± 2 15 ± 1 16 ± 1 15 ± 1 16 ± 1 Creatinine ( mg/ dL) 1.1 ± 0.0 1.1 ± 0.0 1.2 ± 0.0 1.2 ± 0.0 1.1 ± 0.0 1.2 ± 0.1 1.2 ± 0.1 1.4 ± 0.1*

BUN/ creatinine ratio 13.8 ± 1 11.7 ± 0.7 11.4 ± 0.7 13.2 ± 1 14 ± 1 13 ± 1 12 ± 1 12 ± 1 Calcium ( mg/ dL) 10.1 ± 0.2 9.4 ± 0.1* 9.6 ± 0.2 9.2 ± 0.3* 9.7 ± 0.1 10.1 ± 0.2 9.8 ± 0.2 9.8 ± 0.2 Phosphorus ( mg/ dL) 5.7 ± 0.3 5.1 ± 0.2 6.1 ± 0.4 5.9 ± 0.3 5.0 ± 0.2 5.1 ± 0.2 5.4 ± 0.2 5.0 ± 0.2 Chloride ( mEq/ L) 105 ± 1 103 ± 1 103 ± 1 107 ± 1 107 ± 1 106 ± 2 105 ± 1 106 ± 1 Adapted from Hayes et al. ( 1986).

a. All data expressed as mean ± SEM.

* Significantly different from vehicle control at p

0.05. Drinking Water Criteria Document for Haloacetonitriles V  26 EPA/ OW/ OST/ HECD Final Draft C. Reproductive and Developmental Effects No multigeneration studies were located on the reproductive effects of any of the haloacetonitriles ( HANs). No reproductive or developmental studies were identified for any of the HANs following inhalation or dermal exposure.

Meier et al. ( 1985) evaluated the in vivo genotoxicity of several drinking water disinfectants and their by products in mice. As part of this study, the ability of a series of disinfectants to induce sperm head shape abnormalities was examined as a measure of germ cell mutagenicity. Of the disinfectants tested, only hypochlorite induced a dose related increase in the percent of abnormal sperm heads. To investigate whether this positive finding with hypochlorite might be due to the in vivo formation of haloacetonitriles from hypochlorite, groups of 8 to11

week old male B6C3F 1 mice ( 10/ dose group) were administered 0, 12.5, 25, or 50 mg/ kg/ day of BCAN, DBAN, DCAN, or TCAN in water by gavage for 5 days. The authors indicated that the highest total dose of 250 mg/ kg ( 5 days x 50 mg/ kg/ day) was selected to approximate the reported LD 50 values for these compounds. Positive controls received five daily doses of 200 mg/ kg ethylmethanesulfonate administered intraperitoneally. Control and test animals were sacrificed three or five weeks after the last treatment and sperm recovered from the caudae epididymides were examined for abnormal sperm head morphology. The study authors reported no effects on sperm head shape abnormalities for any of the HANs at doses up to 50 mg/ kg/ day.

The positive control yielded an increase in sperm head abnormalities. The NOAEL for effects on sperm head abnormalities in this study is 50 mg/ kg/ day for all of the HANs that were tested. The Drinking Water Criteria Document for Haloacetonitriles V  27 EPA/ OW/ OST/ HECD Final Draft HANs were not tested for their potential to cause micronuclei or chromosome aberrations, since no positive results were seen for these end points with the initial disinfectants.

R. O. W. Sciences ( 1997) conducted a reproductive and developmental toxicity screening study for DBAN. Based on the results of the dose range finding studies ( described in detail in the section on shorter term toxicity studies), the main reproductive and developmental toxicity study included concentrations of 0, 15, 50, or 150 ppm DBAN in the drinking water of Sprague Dawley rats ( 10 animals/ dose). Male rats, 11 weeks of age on study day 1, were given treated water on study days 6 through 34 or 35 ( the day of necropsy). The estimated doses resulting from exposure to 0, 15, 50, or 150 ppm DBAN were 0, 1.4, 3.3, and 8.2 mg/ kg/ day ( calculated by the study authors from body weight and water consumption data). Males were examined for clinical signs of toxicity and body weight at various intervals during the study. At study termination,

clinical pathology ( hematology and clinical chemistry), body and organ ( liver, right kidney, spleen,

thymus, right testis, right epididymis, right cauda epididymis) weight measurements, and gross and histopathology ( of the same array of organs for which organ weight was determined) were conducted. Sperm analyses for control and high dose males included sperm motility, spermatid head counts, sperm morphology, and chromatin structure. Sperm density was assessed for all necropsied males. All parameters evaluated were within normal limits, unless described below.

The only concentration related effects in males were a slight ( 4%) body weight decrease at 150 ppm that was not statistically significant, and statistically significant decreases in water consumption at 50 ppm and 150 ppm. Decreased water consumption ( reported on the basis of grams/ kg body weight/ day) ranged from 68% to 78% for study days 8, 10, 21, and 33 in the 50 Drinking Water Criteria Document for Haloacetonitriles V  28 EPA/ OW/ OST/ HECD Final Draft ppm group, and from 53% to 67% in the 150 ppm exposure group. Decreased food consumption ( reported on the basis of grams/ kg body weight/ day) was observed at 150 ppm on study days 8 ( 85% of controls) and 10 ( 89% of controls), but not on study days 21 or 33. A 14% decrease in the absolute weight of the right cauda epididymis was reported for the 50 ppm males. However,

the absence of a statistically significant effect at the highest concentration, coupled with the absence of effects on any of the measured sperm parameters or on epididymal histopathology suggests that this result is not biologically significant.

Female rats were divided into two groups to separately evaluate toxicity during conception and early gestation ( designated as Group A) versus toxicity during gestation through parturition ( designated as Group B). Female rats for both groups were approximately 11 weeks of age on study day 1. Group A females ( 10 animals/ dose) were exposed on study days 1 through 34, cohabitated with treated males on days 13 through 17, and examined on day 34 ( the day of necropsy). The estimated doses resulting from exposure to 0, 15, 50, or 150 ppm DBAN were 0,

1.8, 5.1, and 10.9 mg/ kg/ day ( calculated by the study authors from body weight and water consumption data). Group A females were removed from cohabitation upon detection of vaginal sperm or a copulatory plug, or after five days in the absence of mating. Clinical signs of toxicity,

body weight, and feed and water consumption were determined at various intervals. At study termination ( corresponding to gestation day 16 20), gross necropsy was performed and the rats were evaluated for pregnancy status, number and position of live and dead fetuses, number and position of early and late resorptions, and number of corpora lutea. Mating, pregnancy, and fertility indices, total number of implants, and pre and post implantation losses were calculated. Drinking Water Criteria Document for Haloacetonitriles V  29 EPA/ OW/ OST/ HECD Final Draft No treatment related changes were observed in any of the mating, fertility, pregnancy, or developmental endpoints that were examined. All of the maternal parameters evaluated were within normal limits, except for a 7% decrease in terminal body weight ( non statistically significant) observed at 150 ppm. Feed and water consumption was also significantly decreased in this concentration group. Feed consumption was statistically significantly decreased only on study day 3 ( 88% of controls) in the 150 ppm exposure group; slight decreases in food consumption on other study days were not statistically significant. Water consumption was statistically significantly decreased in the 150 ppm exposure group on all four study days for which the data were summarized. The decrease in water consumption relative to controls was 52% on day 3, 72% on day 5, 67% on day 21, and 71% on day 33. Slight decreases in water consumption in the 50 ppm exposure group were not statistically significant.

Group B females ( 13 animals/ dose) were cohabitated with males beginning on study day 1,

were separated as soon as they were sperm positive, had a copulatory plug, or on study day 5,

and then were exposed on gestation day 6 through postnatal day ( PND) 1. The estimated doses resulting from exposure to 0, 15, 50, or 150 ppm DBAN were 0, 1.9, 5.3, and 10.8 mg/ kg/ day ( calculated by the study authors from body weight and water consumption data). The group B females were examined on PND 5, and pups were examined on PND 1, 3, and 5. Clinical signs of toxicity, body weight, and feed and water consumption were determined at various intervals during the study. At study termination ( PND 5), females were examined for terminal body weight, gross necropsy was performed, and uterine evaluation was conducted with the number of implantation sites and resorptions recorded. On PND 1, 3, and 5, pup weights, number of live Drinking Water Criteria Document for Haloacetonitriles V  30 EPA/ OW/ OST/ HECD Final Draft and dead pups, and number of male and female pups were recorded. The anogenital distance of pups was also measured on PND 1. All parameters were within normal control limits except for decreased water consumption in the 50 and 150 ppm exposure groups on study days 8, 14, and 21, but not on study day 6. In the 50 ppm exposure group water consumption decreases ranged from 72% to 83% of controls. In the 150 ppm exposure group water consumption decreases ranged from 47% to 60%.

This study suggests that DBAN is not a reproductive or developmental toxicant at the highest dose tested, since no treatment related effects on reproductive or developmental parameters were observed for males or Group A or Group B females. However, definitive conclusions regarding the potential for DBAN to induce reproductive or developmental effects is hampered by the fact that this was a screening study that was not designed to evaluate the full spectrum of endpoints of interest. For example, males were not exposed during all stages of spermatogenesis, and the pups were not evaluated for possible visceral or skeletal malformations.

The NOAEL for male reproductive effects in this study is 8.2 mg/ kg/ day. The NOAEL for female reproductive and developmental effects is 10.8 mg/ kg/ day ( the lower of the calculated doses for Group A and Group B females exposed to the highest DBAN concentration). No LOAEL was identified for male or female reproductive or developmental toxicity. Slight ( non statistically significant) body weight decreases were observed in high concentration group males and Group A females. However, these results were not considered biologically significant because of their limited magnitude, and the fact that they might be secondary to decreased water consumption,

rather than due to a toxicological effect. The systemic effects evaluated in males were more Drinking Water Criteria Document for Haloacetonitriles V  31 EPA/ OW/ OST/ HECD Final Draft extensive than in females, and yield a critical effect level with the greatest degree of certainty.

Therefore, the NOAEL of 8.2 mg/ kg/ day for males is selected as the overall study NOAEL for systemic toxicity. No LOAEL was identified.

Smith and colleagues have investigated the reproductive and developmental effects of BCAN, DBAN, DCAN, and TCAN through a series of studies, first testing different HANs at the maximum tolerated dose, and then conducting dose response studies with the individual compounds. These studies were followed by investigation of the potentially confounding role of the tricaprylin solvent vehicle in the observed developmental toxicity. NOAELs and LOAELs are reported for the initial studies, conducted in tricaprylin vehicle, for completeness. However, as discussed below these NOAELs are not adequate to serve as the basis for the dose response assessment, since tricaprylin itself contributes to the developmental toxicity of the HANs, and the effects due to tricaprylin cannot be separated from those that are due to the test chemical.

In the initial developmental toxicity screening studies ( Smith et al., 1986; Smith et al.,

1987) sexually mature Long Evans rats ( age not provided, 20 to 26 sperm positive rats per dose group) were administered gavage doses on gestation days 7 to 21 at the maximum tolerated dose for each chemical dissolved in tricaprylin: 55 mg/ kg/ day for BCAN, DCAN, and TCAN; 50 mg/ kg/ day for DBAN. Control rats received tricaprylin alone. No food and water consumption data were reported. Duration of pregnancy, litter size, sex ratios, and litter weights were determined at birth. Dams not delivering by day 23 of gestation were sacrificed, and the status of their pregnancies was determined. Survival and weight gain of offspring to PND 4 were Drinking Water Criteria Document for Haloacetonitriles V  32 EPA/ OW/ OST/ HECD Final Draft measured. At PND 6, litters were randomly culled to a consistent number. At weaning, the number of pups per litter were further randomly reduced to a consistent number. The remaining pups were monitored for growth and health until study day 41 42. At this time, all surviving pups were sacrificed and subjected to gross necropsy, with livers, kidneys, spleens, lungs, and gonads removed and weighed. The following parameters were evaluated: maternal toxicity ( maternal weight gain and mortality), reproductive success ( percent pregnant, percent delivering viable litters), and growth and viability of pups ( live pups/ litter; postnatal survival; mean birth weight;

survival, weight gain and development through necropsy; gross necropsy and organ weight results). All parameters evaluated were within normal limits, unless described below.

Treatment related increases in maternal deaths were observed for DBAN ( 15% of dams)

and for TCAN in one study group ( 20% of dams), although the result for TCAN was not replicated in a second study group receiving the same dose. The study authors could not explain the disparate results in duplicate dose groups for this endpoint, or other endpoints noted below.

Maternal deaths observed in the BCAN and DCAN groups were attributed to intubation error.

No maternal deaths were reported in the vehicle controls. All four HANs caused decreased maternal weight gain during gestation, although the decrease was not statistically significant for BCAN. Since the maternal weights were measured prior to delivery, decreased unadjusted maternal weights are affected by litter resorptions and decreased pup weights. As a result, the observed decreases in maternal weight gain cannot be attributed solely to maternal toxicity. Both DCAN and TCAN decreased the percentage of sperm positive females that became pregnant, but the decrease in apparent pregnancy rate was observed only in one of two duplicate dose groups Drinking Water Criteria Document for Haloacetonitriles V  33 EPA/ OW/ OST/ HECD Final Draft for each compound. According to the study authors, the apparent decrease in pregnancy rate could reflect late preimplantation losses or very early resorptions. This conclusion was based on the expected stage of pregnancy at the initiation of dosing on gestation day ( GD) 7 and the absence of detectable ammonium sulfide staining decidua ( which would indicate implantation sites). DCAN and TCAN ( in only one of two duplicate dose groups for TCAN) also decreased the percentage of females delivering viable litters and increased the percentage of litters totally resorbed ( p

0.05). Mean birth weights were reduced for all four compounds ( p

0.05), and postnatal survival on day 4 was significantly reduced ( p

0.05) in pups from dams exposed to DCAN ( in only one of two duplicate dose groups) and TCAN. Cursory visual inspection of nonviable pups did not reveal gross terata, although the authors noted that the current study protocol was not expected to provide a sensitive assessment of malformations because of the embryolethality of the single doses that were tested. No pups died after the culling of litters on day 6 until weaning, but after weaning some pups ( number not reported) from the DCAN replicate died because they were too small to reach the water source. Weight gain to PND 4 was significantly decreased for male and female pups for BCAN and DCAN ( in one of two duplicate dose groups), and in male pups for DBAN. Measurement of pup weights at weaning ( days 21

22) and adolescence ( days 41 42) revealed significant decreases in pup body weight at both time points for TCAN and at adolescence for DCAN ( in one of two duplicate dose groups). Postnatal weight gain from weaning until puberty and sacrifice was reduced in both male and female pups in all groups administered HANs, but these effects were statistically significant only in males and females receiving TCAN and females receiving BCAN. A scattering of effects in the relative organ to body weight ratios obtained at the day 41 42 sacrifice was reported across all dose Drinking Water Criteria Document for Haloacetonitriles V  34 EPA/ OW/ OST/ HECD Final Draft groups. However, these effects are judged to be without biological significance because no uniform or consistent patterns of change were noted.

This study identifies a frank effect level ( FEL) of 50 mg/ kg/ day for DBAN and 55 mg/ kg/ day for TCAN, based on increased maternal deaths. A maternal LOAEL of 55 mg/ kg/ day is assigned for BCAN and DCAN based on decreased maternal weight gain.

However, the contribution of developmental effects ( e. g. decreased litter sizes, fetal weight) to the observed decrease in weight gain cannot be ruled out, and therefore it is not known if the observed effect was due to systemic toxicity, severe developmental toxicity, or both. The LOAEL for developmental effects was 50 to 55 mg/ kg for all four compounds. For DCAN and TCAN, these LOAELs were based on severe effects ( increased percent of early resorptions or litters totally resorbed). Maternal and developmental NOAELs were not identified due to the limited number of test doses used in this screening assay.

In a follow up study evaluating the dose response for TCAN, Smith et al. ( 1988)

administered TCAN to sperm positive Long Evans rats aged 65 80 days ( 19 to 21 per dose group of TCAN treated animals, 30 rats in the vehicle control group, and 10 rats in water controls) by gavage in tricaprylin at doses of 0, 1.0, 7.5, 15, 35, or 55 mg/ kg/ day on gestation days 6 to 18.

Dams were sacrificed on day 21 of gestation, their uterine horns were examined for number and location of fetuses or resorption sites, and the fetuses were removed and examined. Two thirds of each litter was fixed for dissection and one third stained for bone and cartilage examination.

The parameters evaluated in this study were maternal toxicity ( maternal weight gain and Drinking Water Criteria Document for Haloacetonitriles V  35 EPA/ OW/ OST/ HECD Final Draft mortality), embryolethality ( number and location of fetuses or resorption sites, number of viable litters, litter sizes), and the following fetal effects: weights, sex ratios, and structural abnormalities including external, visceral, and skeletal effects. All parameters were within normal limits, unless described below ( statistical comparisons are made to the tricaprylin control unless noted otherwise). The high dose was lethal in 4 out of 19 dams, and maternal weight gain was decreased beginning at 15 mg/ kg/ day, but maternal weight gain adjusted for fetal weight and excluding females with full litter resorptions was significantly decreased only at 55 mg/ kg/ day.

The primary developmental effects were on fetal viability, malformations, and fetal body weight.

There was a dose related increase in full litter resorptions ( compared to both water and tricaprylin control groups) at 7.5 mg/ kg/ day and higher, affecting 2/ 3 of the surviving dams at the high dose.

Fetal weight was significantly decreased only at 35 mg/ kg/ day, while post implantation loss ( as percent resorptions per litter) was significantly elevated at doses of 15 mg/ kg/ day and higher.

Post implantation losses were significantly higher and male fetal body weight was significantly lower ( p

0.05) in the tricaprylin controls compared to the water controls. The authors noted that the value obtained for post implantation loss for water controls in this study was lower than the historical laboratory control levels for this endpoint, which might have enhanced the observed effect of the tricaprylin control. While no malformations were observed in the water controls,

soft tissue ( fetal incidence of 3.8% and litter incidence of 6/ 30) and skeletal ( fetal incidence of 13.3% and litter incidence of 7/ 30) malformations were observed in tricaprylin controls. The percent of fetuses affected per litter with soft tissue malformations was dose dependent, ranging from 18% at 7.5 mg/ kg/ day to 35% at 35 mg/ kg/ day. This value decreased to 22% at the high dose. The soft tissue malformation frequency at the low dose of 1.0 mg/ kg/ day ( 8.4% of fetuses Drinking Water Criteria Document for Haloacetonitriles V  36 EPA/ OW/ OST/ HECD Final Draft affected per litter) was not statistically different from vehicle controls ( 3.8%), although the authors expressed concern that this level of malformations could be of biological significance.

However, consideration of the data in terms of the more appropriate unit ( U. S. EPA, 1991) of percent of litters affected indicated no effect of TCAN compared to the tricaprylin control ( 4/ 20 litters affected at 1 mg/ kg/ day and 6/ 30 litters affected in tricaprylin controls). The incidence of pups with cardiovascular malformations was increased at 15 mg/ kg/ day and urogenital malformations were significantly increased at both 15 mg/ kg/ day and 35 mg/ kg/ day ( p

0.05); the percentage of litters with cardiovascular malformations was increased beginning at 7.5 mg/ kg/ day ( 8/ 18 litters affected versus 6/ 30 litters in tricaprylin controls). No increase in the incidence of external or skeletal malformations was reported, and the study authors do not appear to have reported on the incidences of external, skeletal, or internal variations ( which are more subtle effects than malformations).

Since the apparent increase in soft tissue malformations at 1.0 mg/ kg/ day was not statistically significant when compared to tricaprylin controls, and there was no effect on the percentage of affected litters, a dose of 1.0 mg/ kg/ day of TCAN is the NOAEL for developmental toxicity, and 7.5 mg/ kg/ day is the LOAEL, although this conclusion is limited by the absence of reported data on the incidence of variations. The maternal NOAEL for this study is 35 mg/ kg/ day. The next higher dose of 55 mg/ kg/ day is a FEL, based on increased maternal deaths.

Although significant decreases in overall maternal body weight gain were observed beginning at 15 mg/ kg/ day, adjusted body weight gain was decreased only at 55 mg/ kg/ day. Adjusted body weight gain is a more appropriate indicator of maternal toxicity than overall body weight gain, Drinking Water Criteria Document for Haloacetonitriles V  37 EPA/ OW/ OST/ HECD Final Draft since the latter would be affected by developmental toxicity as well as maternal toxicity.

Regardless of the selection of the critical effect levels, the results of this study are not appropriate for dose response analysis due to the confounding effects of the tricaprylin solvent vehicle as evidenced by the differences between water and tricaprylin controls in this study, and as described in more detail in the study by Christ et al. ( 1996) for TCAN.

In a second follow up study evaluating the dose response for DCAN, Smith et al. ( 1989)

administered DCAN in tricaprylin to pregnant Long Evans rats aged 65 80 days ( 22 to 24 rats per dose group) by gavage at doses of 0, 5, 15, 25, or 45 mg/ kg/ day on gestation days 6 to 18.

Tricaprylin served as the vehicle control and distilled water served as an additional control group.

Dams were sacrificed on day 20 of gestation. Their livers, spleens, and kidneys were removed and weighed, and their uterine horns were examined for number and location of fetuses or resorption sites. Fetuses were removed and examined. Two thirds of each litter was fixed for dissection and one third stained for bone and cartilage examination. The parameters evaluated in this study were maternal toxicity ( maternal body weight gain, liver, spleen, and kidney weight, and mortality), embryolethality ( number and location of fetuses or resorption sites, number of viable litters, litter sizes), and the following fetal effects: weights, crown rump lengths, sex ratios, and structural abnormalities including external, visceral, and skeletal effects. All parameters were within normal limits, unless described below ( statistical comparisons are made to the tricaprylin control unless noted otherwise). Two of 22 dams died in the high dose group ( 45 mg/ kg/ day).

No maternal deaths were observed in any other dose group, or in either of the control groups.

Maternal body weight gain and adjusted maternal body weight gain were significantly decreased Drinking Water Criteria Document for Haloacetonitriles V  38 EPA/ OW/ OST/ HECD Final Draft ( p

0.05) at 45 mg/ kg/ day. Surviving dams in the 45 mg/ kg/ day dose group showed elevated spleen and kidney weights ( p

0.05). Liver weight was significantly increased at 25 mg/ kg/ day,

but not at the high dose.

DCAN treatment affected a number of developmental toxicity parameters.

Postimplantation loss was significantly increased beginning at 25 mg/ kg/ day. At 45 mg/ kg/ day, the following developmental effects were observed: increased number of totally resorbed litters,

decreased number of viable litters, decreased fetal weight in males and females, and decreased crown rump length in male and females. The incidence of malformations increased in a doserelated manner. Significant increases ( p

0.05) in total soft tissue, cardiovascular, urogenital system, and skeletal malformations were observed in fetuses from dams exposed to 45 mg/ kg/ day.

Increases in the incidence of litters affected was most apparent for total soft tissue malformations:

water control 0/ 19; tricaprylin control 4/ 19; 5 mg/ kg/ day 5/ 22; 15 mg/ kg/ day 5/ 16; 25 mg/ kg/ day 9/ 16; 45 mg/ kg/ day 7/ 7. The authors did not present a statistical analysis of the malformation data presented as the incidence of litters affected. Analysis of these data for preparation of this Criteria Document ( using Fischer's Exact Test) revealed that the litter incidence of malformations is significantly greater ( P< 0.05) in the 25 mg/ kg/ day and 45 mg/ kg/ day dose groups than in tricaprylin controls. The water controls were not significantly different than the tricaprylin controls ( P = 0.0525), based on the litter incidence, although the study authors reported a significant difference between water and tricaprylin controls on the basis of fetuses affected per litter. Drinking Water Criteria Document for Haloacetonitriles V  39 EPA/ OW/ OST/ HECD Final Draft The NOAEL for maternal toxicity for this study is 15 mg/ kg/ day and the LOAEL is 25 mg/ kg/ day based on increased liver weight. No additional measures of liver toxicity were included in this study, but the selection of increased liver weight as an adverse effect is supported by ability of DCAN to induce liver toxicity at similar doses as reported in the subacute and subchronic studies by Hayes et al. ( 1986). The NOAEL for developmental toxicity is 15 mg/ kg/ day and the LOAEL is 25 mg/ kg/ day, based on increased post implantation loss and malformations. No differences in fetal viability or size were noted between the water and tricaprylin control groups. However, the incidence of total soft tissue malformations was significantly lower in water ( 0/ 19 litters affected) versus tricaprylin controls ( 4/ 19 litters affected), suggesting that tricaprylin induces developmental toxicity, and that this effect is further potentiated following combined treatment with tricaprylin and DCAN ( as high as 9/ 16 litters affected). Such a finding is of considerable importance because, as mentioned in the paragraph above introducing the Smith studies, it raises the possibility that the fetal malformations attributed to the HANs in this study may result from the potentiation by the tricaprylin vehicle. Tricaprylin effects and their implications for assessing HAN toxicity are discussed further in the studies below.

The developmental toxicity of BCAN was evaluated in 120 to 150 day old Long Evans rats ( Christ et al., 1995) in a third study to evaluate the dose response for HANs. Pregnant rats were administered BCAN by gavage in tricaprylin on gestation days 6 to 18 at doses of 0, 5, 25,

45, or 65 mg/ kg/ day. Dams treated with tricaprylin only served as the vehicle control and dams treated with distilled water served as an additional control. The number of rats per treatment Drinking Water Criteria Document for Haloacetonitriles V  40 EPA/ OW/ OST/ HECD Final Draft group is unclear, because while the methods section indicates that between 17 and 23 animals were assigned per dose group, results for reproductive performance presented in the paper suggest group sizes of 20 to 28 dams. Dams were sacrificed on day 20 of gestation. Their livers,

spleens, and kidneys were removed and weighed, and their uterine horns were examined for number and location of fetuses or resorption sites. Fetuses were removed and examined.

Twothirds of each litter was fixed for dissection and one third stained for bone and cartilage examination. The parameters evaluated in this study were maternal toxicity ( maternal body weight gain, liver, spleen, and kidney weight, and mortality), embryolethality ( number and location of fetuses or resorption sites, number of viable litters, litter sizes), and the following fetal effects: weights, crown rump lengths, sex ratios, and structural abnormalities, including external,

visceral, and skeletal effects. All parameters were within normal limits, unless described below ( statistical comparisons are made to the tricaprylin control unless noted otherwise).

Treatment with BCAN in tricaprylin resulted in both maternal and developmental toxicity.

Mortality was statistically significantly increased in the high dose dams ( 3 of 26 treated dams)

compared with the water and tricaprylin control groups in which no treatment related deaths were observed. Dams in both the 45 and 65 mg/ kg/ day dose groups had significantly decreased percentage of body weight gain compared with the tricaprylin control group, and dams in the 65 mg/ kg/ day group had decreased body weight gain after adjusting for gravid uterine weight.

Kidney weights were significantly increased in dams at doses

25 mg/ kg/ day compared with tricaprylin controls. Liver and spleen weights were significantly increased only in the high dose Drinking Water Criteria Document for Haloacetonitriles V  41 EPA/ OW/ OST/ HECD Final Draft dams. There was no difference in any of these maternal toxicity parameters between tricaprylin and water controls.

In terms of reproductive and developmental endpoints, an increase in the number of litters totally resorbed and a decrease in the number of viable litters compared with tricaprylin controls were observed beginning at 45 mg/ kg/ day. The percent post implantation loss and the percent of resorbed fetuses per litter increased at doses of

45 mg/ kg/ day compared with the tricaprylin control group. Although the increase for post implantation loss was not statistically significant in the high dose group, this parameter was statistically significant in the second highest dose group,

and also substantially elevated, although not statistically, in the third highest dose group. Fetal crown rump length was significantly decreased in all the treated groups, and fetal weights were decreased beginning at 25 mg/ kg/ day. Statistical analysis of the malformation data was done in terms of the percent of fetuses affected per litter. However, examination of the data on the basis of the percent of litters affected appears to lead to similar conclusions with regard to selection of effect levels. A significant increase in cardiovascular malformations compared with tricaprylin controls was observed in all the dose groups. Urogenital malformations were significantly increased only at 45 mg/ kg/ day, although the percentage of litters affected appeared to be significantly greater at both 25 and 45 mg/ kg/ day. Total soft tissue malformations were increased beginning at 25 mg/ kg/ day and skeletal malformations were significantly increased beginning at 45 mg/ kg/ day. Drinking Water Criteria Document for Haloacetonitriles V  42 EPA/ OW/ OST/ HECD Final Draft In addition to the BCAN treated groups, tricaprylin vehicle alone had significant effects on embryotoxicity. Developmental endpoints affected by tricaprylin are noted as mean ± standard error tricaprylin versus water controls. Animals in the tricaprylin control group had significantly increased percent post implantation loss ( 15.5 ± 19.0 versus 6.7 ± 9.8); decreased fetal body weight ( grams) in males ( 3.19 ± 0.3 versus 3.61 ± 0.2); decreased fetal body weight ( grams) in females ( 2.90 ± 0.3 versus 3.41 ± 0.2); decreased crown rump length ( cm) in males ( 3.4 ± 0.2 versus 3.6 ± 0.2); and decreased crown rump length ( cm) in females ( 3.4 ± 0.2 versus 3.5 ± 0.1).

An increased incidence of urogenital malformations ( 0/ 19 litters versus 4/ 23 litters) was also induced by tricaprylin as compared to water controls.

The NOAEL for maternal toxicity in this study is 45 mg/ kg/ day and the high dose of 65 mg/ kg/ day is a FEL based on increased maternal deaths, and accompanied by decreased adjusted maternal weight gain, and organ weight changes. The increase in kidney weight at the lower doses was not used as the basis for assigning maternal toxicity effect levels because of the absence of data to determine whether this effect was adverse. Developmental effects, including decreased crown rump length and increased cardiovascular malformations were observed at the the low dose of 5 mg/ kg/ day. Therefore, 5 mg/ kg/ day is a developmental LOAEL for this study, and no NOAEL is identified. However, based on the observation of embryotoxicity of the tricaprylin vehicle in this study, and later work by this laboratory which suggests that tricaprylin may act synergistically with TCAN to enhance developmental toxicity ( Christ et al., 1996), use of this study for dose response assessment is not appropriate, because it may not accurately reflect the toxicity of BCAN in drinking water in the absence of the tricaprylin vehicle. Drinking Water Criteria Document for Haloacetonitriles V  43 EPA/ OW/ OST/ HECD Final Draft Based on the observed increased embryotoxicity in tricaprylin versus water treated controls in earlier studies, Christ et al. ( 1996) investigated the effect of solvent vehicle on the developmental toxicity of TCAN ( Table V 10). Groups of approximately 20 sperm positive Long Evans rats aged 65 80 days were treated with 15, 35, 55, or 75 mg/ kg/ day TCAN in corn oil, or 15 mg/ kg/ day TCAN in tricaprylin. In addition, water, corn oil, and tricaprylin were used as controls. Treatments were administered by oral gavage on gestation days 6 to 18 for all of the treatment groups. Dams were sacrificed on day 20 of gestation. Their livers, spleens, and kidneys were removed and weighed, and their uterine horns were examined for number and location of fetuses or resorption sites. Fetuses were removed and examined. Two thirds of each litter was fixed for dissection and one third stained for bone and cartilage examination. The parameters evaluated in this study were maternal toxicity ( maternal body weight gain, liver, spleen, and kidney weight, and mortality), embryolethality ( number and location of fetuses or resorption sites,

number of viable litters, litter sizes), and the following fetal effects: weights, crown rump lengths,

sex ratios, and structural abnormalities, including external, visceral, and skeletal effects. All parameters were within normal limits, unless described below. Of the 20 dams treated with 75 mg/ kg/ day TCAN in corn oil, five dams died, five were nonpregnant, and nine dams resorbed their entire litter, so that only one viable litter was produced. For this reason, other data for the 75 mg/ kg/ day group were not reported. No maternal deaths were reported in any of the other groups. Maternal weight gain was significantly decreased beginning at 15 mg/ kg/ day, and maternal weight gain, after adjusting for gravid uterine weight, was significantly decreased at 35 mg/ kg/ day and higher doses. Relative maternal liver weight was increased at

35 mg/ kg/ day, and liver, spleen, and kidney weights were significantly increased at 55 mg/ kg/ day as compared to Drinking Water Criteria Document for Haloacetonitriles V  44 EPA/ OW/ OST/ HECD Final Draft corn oil vehicle controls. Similarly, increased relative liver weights were observed in the animals given 15 mg/ kg/ day TCAN in tricaprylin as compared to 15 mg/ kg/ day TCAN administered in corn oil.

The percent post implantation loss was significantly increased, and the number of live fetuses per litter, fetal body weight, and crown rump length were all significantly decreased in the group administered 55 mg/ kg/ day TCAN in corn oil vehicle. Also in this dose group, the mean percentage of fetuses per litter with external malformations, skeletal malformations, and softtissue malformations was significantly increased. The incidence of cardiovascular and urogenital malformations was not increased at any dose, but other soft tissue malformations ( i. e. not classified as either cardiovascular or urogenital) were significantly increased in the 55 mg/ kg/ day dose group. Drinking Water Criteria Document for Haloacetonitriles EPA/ OW/ OST/ HECD Final Draft V  45 Table V 10. Reproductive and Developmental Toxicity of TCANa.

TCAN ( mg/ kg/ d ) TCAN ( mg/ kg/ d)

Water Corn oil 15 35 55 75 Tricaprylin 15 No. Sperm positive treated females 20 20 17 21 21 20 20 21 Nonpregnant females 2 3 0 0 3 5 1 2 Deaths 0 0 0 0 0 5b 0 0 Litters totally resorbed 0 0 0 0 1 9b 0 3 Viable littersc 18 17 17 21 17 1 19 16 Percent maternal weight gaind 45.5 ± 9.0 48.9 ± 8.1 44.2 ± 7.8b 36.9 ± 8.8b 23.7 ± 7.7b nde 41.6 ± 10.0 41.7 ± 10.5 Adjusted maternal weight gaind 18.6 ± 5.5 20.4 ± 4.8 17.6 ± 5.6 11.1 ± 5.7b 6.8 ± 5.4b nde 18.1 ± 5.8 17.3 ± 6.7 Total implants per litter 13.6 ± 1.6 14.2 ± 1.3 13.8 ± 1.3 13.4 ± 2.4 13.8 ± 1.7 12.3 ± 4.3 12.7 ± 2.5f 13.4 ± 3.0 Percent preimplantation lossg 3.3 ± 5.6 4.6 ± 7.5 3.5 ± 6.6 5.0 ± 13.8 3.8 ± 8.3 17.7 ± 28.1 14.0 ± 13.9h 11.6 ± 20.0 Percent post implantation lossi 12.1 ± 12.4 7.0 ± 7.6 6.3 ± 7.0 8.4 ± 11.1 29.7 ± 25.2b 98.8 ± 4.0b 15.9 ± 17.8 25.4 ± 34.8 Live fetuses per litter 11.9 ± 2.1 13.2 ± 1.6 12.9 ± 1.4 12.3 ± 2.8 10.4 ± 2.6b nd 10.7 ± 3.2f 12.4 ± 1.8 Fetal body weight ( g) ( male) 3.58 ± 0.25 3.42 ± 0.17 3.52 ± 0.20 3.38 ± 0.30 2.54 ± 0.46b nd 3.22 ± 0.28f 2.93 ± 0.37b, j Fetal body weight ( g) ( female) 3.41 ± 0.26 3.28 ± 0.19 3.26 ± 0.22 3.25 ± 0.30 2.39 ± 0.41b nd 3.01 ± 0.25h 2.76 ± 0.38b, j Crown rump length ( cm)

( male) 3.58 ± 0.08 3.57 ± 0.11 3.60 ± 0.13 3.54 ± 0.14 3.28 ± 0.23b nd 3.45 ± 0.13h 3.46 ± 0.16j Drinking Water Criteria Document for Haloacetonitriles Table V 10. Reproductive and Developmental Toxicity of TCANa.

TCAN ( mg/ kg/ d ) TCAN ( mg/ kg/ d)

Water Corn oil 15 35 55 75 Tricaprylin 15 EPA/ OW/ OST/ HECD Final Draft V  46 Crown rump length ( cm)

( female) 3.51 ± 0.09 3.52 ± 0.09 3.51 ± 0.17 3.45 ± 0.13 3.16 ± 0.19b nd 3.36 ± 0.13h 3.36 ± 0.18j Fetal malformations External 0 0 0 1.4 ± 6.2 ( 1/ 21) k 5.3 ± 8.8b ( 6/ 17) nd 0 0.5 ± 1.9 ( 1/ 16)

Total soft tissue ( visceral) 0 1.9 ± 4.3 ( 3/ 17) 0.6 ± 2.4 ( 1/ 17) 0 15.0 ± 17.1b ( 3/ 17) nd 2.4 ± 6.3 ( 3/ 19) 15.0 ± 21.2b, j ( 9/ 16)

Cardiovascular 0 0.6 ± 2.4 ( 1/ 17) 0.6 ± 2.4 ( 1/ 17) 0 4.5 ± 11.2 ( 3/ 17) nd 1.3 ± 5.7 ( 1/ 19) 12.0 ± 17.3b, j ( 7/ 16)

Urogenital 0 1.3 ± 3.8 ( 2/ 17) 0 0 0 nd 2.4 ± 6.3 ( 3/ 19) 3.0 ± 7.1 ( 4/ 16)

Other soft tissues 0 0 0 0 10.4 ± 16.3b ( 3/ 17) nd 0 0 Skeletal 0 0 0 1.6 ± 7.3 ( 1/ 21) 7.1 ± 14.4b ( 2/ 17) nd 0 0 a. Adapted from Christ et. al. ( 1996)

b. Significantly different from vehicle control, p

0.05.

c. Viable litters were those containing at least one live pup.

d. Weight gain analysis included only females with viable litters ( mean ± SD reported); adjusted maternal weight gain controls for the effect of gravid uterine weight.

e. Values for 75 mg/ kg/ day are not reported since there was only one dam with a viable litter.

f. Significantly different from corn oil control, p

0.05. g. Percent preimplantation loss = ( number of copora lutea  number of implants) / number of corpora lutea x 100.

h. Significantly different from water or corn oil control, p

0.05.

i. Percent post implantation loss = ( number of implants  number of live fetuses) / number of implants x 100.

j. Significantly different from TCAN 15/ mg/ kg/ day in corn oil.

k. ( number litters examined / number of litters affected) Drinking Water Criteria Document for Haloacetonitriles V  47 EPA/ OW/ OST/ HECD Final Draft No developmental effects were observed at the doses below 55 mg/ kg/ day of TCAN in corn oil. By contrast, fetal body weight was significantly decreased in the group dosed with 15 mg/ kg/ day TCAN in the tricaprylin vehicle and the percentage of fetuses per litter with total soft tissue and cardiovascular malformations also was significantly increased in this group ( both comparisons relative to both the the tricaprylin control and the same dose of TCAN in corn oil).

The authors noted a shift in the spectrum of soft tissue malformations from cardiovascular ( communication or vascular defects) and urogenital effects in the groups treated with TCAN in tricaprylin to external cranio facial malformations and positional cardiovascular malformations ( e. g., levocardia) in the groups treated with TCAN in corn oil. Comparing the two vehicles,

TCAN administered at 15 mg/ kg/ day in tricaprylin produced effects, including increased liver and kidney weights, decreased fetal weight, decreased crown rump length, and increased percent of fetuses with soft tissue malformations that were not observed when TCAN was administered at 15 mg/ kg/ day in corn oil. When the water, corn oil, and tricaprylin control groups were compared, no differences were observed between the water and corn oil groups. However, the tricaprylin group had the following statistically significant changes compared to water or corn oil groups: increased maternal kidney weight, decreased total implants per litter, increased preimplantation loss, decreased live fetuses/ litter, and decreased fetal weight and crown rump length.

The comparison of solvent vehicle effects in this study clearly shows that when TCAN is administered in tricaprylin, maternal and developmental toxicity are observed at lower doses and the spectrum of effects is changed as compared to TCAN administered in corn oil. Since corn oil Drinking Water Criteria Document for Haloacetonitriles V  48 EPA/ OW/ OST/ HECD Final Draft and water control responses were not different in this study, but tricaprylin controls showed increased toxicity for several endpoints, the data for TCAN dissolved in corn oil are judged to be most relevant for dose response assessment, and the data for TCAN in tricaprylin are judged as inadequate for use in dose response assessment. Based on this rationale, the NOAEL for maternal toxicity in this study is 15 mg/ kg/ day, and the LOAEL is 35 mg/ kg/ day for decreased maternal body weight after adjusting for gravid uterine weight. No developmental effects were observed below 55 mg/ kg/ day TCAN in corn oil. Therefore, for developmental effects, the NOAEL of TCAN in corn oil is 35 mg/ kg/ day and the LOAEL is 55 mg/ kg/ day. The data sets for maternal and developmental endpoints were further analyzed to determine benchmark doses ( BMDs) according to draft EPA Guidance ( U. S. EPA, 2000c) to identify alternative critical effect levels. The results of the modeling are described in detail in Appendix A. A BMDL of 17 mg/ kg/ day for decreased adjusted maternal body weight gain was selected as the most appropriate modeling result to serve as the basis for the quantitative dose response assessment.

Based on the effects of tricaprylin alone and its ability to potentiate the toxicity of TCAN,

the use of the data from the developmental toxicity studies using this vehicle is not appropriate for risk assessment. The mechanism responsible for the greater sensitivity and different pattern of malformations produced by TCAN when it is administered in tricaprylin instead of water or corn oil is not understood. However, an earlier abstract by this same laboratory ( Gordon et al., 1991)

suggested that multiple exposures to TCAN in tricaprylin results in different distribution of TCAN ( or the TCAN/ tricaprylin combination) to maternal tissues and embryos than occurs when TCAN is administered in corn oil. However, as described in more detail in Chapter III ( Toxicokinetics) Drinking Water Criteria Document for Haloacetonitriles V  49 EPA/ OW/ OST/ HECD Final Draft the data are not sufficient to determine whether the observed differences in response when tricaprylin was used as the solvent vehicle were due to a toxicokinetic or a toxicodynamic interaction.

D. Mutagenicity and Genotoxicity The genotoxicity of HANs has been tested in a variety of diverse assays. In this section we present the study results for the HANs by study type. The results of mutagenicity assays are described first, followed by measures of chromosome effects, and then assays of DNA damage.

Table V 12 provided at the end of this section summarizes the overall genotoxicity study results on a chemical by chemical basis.

The mutagenicity of HANs has been studied by several investigators using standard protocols or variations of the Salmonella typhimurium/ mammalian microsome mutagenesis assay.

In a report summarizing results of the USEPA Gene Tox program, Kier et al. ( 1986) presented the results of testing DCAN in strains TA1535, TA1537, TA1538, TA100, and TA98. These varying tester strains are used to identify a range of mutagenic target sites, where strains TA100 and TA1535 detect transitions and transversions, while strains TA98 and TA1538 detect frameshifts and small deletions/ insertions. ( The study results were initially reported by Simmon et al., 1977; and Simmon and Kauhanen, 1978). A positive result was defined as the generation of greater revertant counts than controls at two doses and at one dose the number of the revertants had to be 3 times greater than control for strains TA1535, TA1537, or TA1538, or twice the control value for TA100 or TA98. Positive mutagenic activity with or without addition of S9 Drinking Water Criteria Document for Haloacetonitriles V  50 EPA/ OW/ OST/ HECD Final Draft activation was reported in strains TA100 and TA1535, while findings were negative regardless of S9 activation in strains TA98, TA1537, and TA1538. Overall, DCAN was positive in this assay.

Bull et al. ( 1985) studied the mutagenic effects of BCAN, DBAN, DCAN, and TCAN in the Salmonella/ microsome assay using tester strains TA98, TA100, TA1535, TA1537, and TA1538. Cells were dosed with up to 5.44 µ mol/ plate BCAN, up to 0.58 µ mol/ plate DBAN, up to 12.4 µ mol/ plate of DCAN, and up to 11.7 µ mol/ plate TCAN. Wide ranges of doses were used and the highest concentration equaled or exceeded the LC 50 for each compound. In this assay,

BCAN was positive in strain TA100 (+ S9) and in TA1535 (+ S9 or  S9), while DCAN was positive in TA98, TA100, and TA1535 regardless of S9 activation. The positive results for BCAN and DCAN in strain TA1535 led the authors to conclude that both of these compounds can induce base pair substitutions. None of the HANs increased the number of revertants in strains TA1537 or TA1538. In addition, DBAN and TCAN failed to produce dose related increases in the frequency of histidine revertants in any strain.

The Ames fluctuation test was performed by exposing S. typhimurium strain TA100 to the HANs in liquid culture with and without the addition of S9 ( Le Curieux et al., 1995). The range of doses tested, in µ g/ mL, were as follows: BCAN, 0.03 10 ( S9), 0.3 100 (+ S9); DBAN, 0.03

10 ( S9), 0.1 30 (+ S9); DCAN, 0.3 300 ( S9), 0.3 1000 (+ S9); TCAN, 0.1 1000 (+/ S9). In all cases the tested dose range included cytotoxic concentrations ( defined as sufficient to reduce the number of positive wells compared to controls). All of the compounds except DBAN generated positive results ( i. e. generated a statistically significant increase in the number of positive wells Drinking Water Criteria Document for Haloacetonitriles V  51 EPA/ OW/ OST/ HECD Final Draft compared to controls). The effective doses ranged widely and no clear pattern of dependence on S9 activation was apparent. BCAN yielded a positive result at a concentration as low as 0.6 µ g/ mL ( S9), DCAN yielded a positive result at 10 µ g/ mL ( S9) and 300 µ g/ mL (+ S9), and TCAN at 30 µ g/ mL ( S9).

Gee et al. ( 1998) conducted a validation analysis for the Salmonella/ microsome assay by comparing the effects of a group of substances in base specific tester strains to mutagenic activity in traditional strains. Mutagenic activity of TCAN in the base specific strains TA7001, TA7002,

TA7003, TA7004, TA7005, and TA7006 and a mix of these six strains was contrasted to the mutagenic activity of TCAN in the traditional frameshift strains TA98 and TA1537. The assay was done using a liquid fluctuation exposure protocol in the presence and absence of S9 activation. For each combination of tester strains, four doses of TCAN ranging from 50 to 1000 µ g/ mL in dimethyl sulfoxide were used, plus solvent control and positive control groups. The test agent was considered mutagenic if any of the test doses were found to generate revertants at statistically significant levels compared to the control. The only positive result reported was for the base specific strain mix with S9 activation. This result must be interpreted with caution,

however, because the authors did not explain why a positive result was identified in the basespecific strain mix, when none of the base specific strains was positive when tested individually.

In addition, inspection of the raw data ( available from an online site provided in the paper) did not reveal any clear increase in the frequency of revertants in the mix versus the individual test strains. Drinking Water Criteria Document for Haloacetonitriles V  52 EPA/ OW/ OST/ HECD Final Draft Taken together, these mutagenicity assays indicate that BCAN and DCAN are mutagenic in S. typhimurium ( Bull et al., 1985; Kier et al., 1986; Le Curieux et al., 1995). TCAN has produced mixed results, with negative results reported in the standard assay protocol ( Bull et al.,

1985; Le Curieux et al., 1995; Gee et al., 1998). DBAN has yielded negative results in all of the assays reported ( Bull et al., 1985; Le Curieux et al., 1995). No studies testing any of these compounds in mammalian gene mutation assays were located.

Several investigators have tested the ability of HANs to induce chromosome damage, with mixed results. Bull et al. ( 1985) studied the ability of HANs to produce chromosomal damage or loss by examining micronuclei production in polychromatic erythrocytes in an in vivo assay in CD

1 mice ( 5 animals/ sex/ group). Animals were dosed by gavage with BCAN, DBAN, DCAN, or TCAN dissolved in 10% Emulphor at 0, 12.5, 25, or 50 mg/ kg/ day for five consecutive days and sacrificed 6 hours after the last dose. The highest dose was selected to generate a cumulative dose ( 5 doses x 50 mg/ kg/ day = 250 mg/ kg/ day) approximating the oral LD 50. No significant increases in micronuclei frequency were observed for any of the HANs tested. It is unclear,

however, whether sufficiently high doses were tested in this study. The study authors did not present the supporting data and did not report whether cytotoxicity of the target tissue occurred ( as evidenced by a change in the ratio of polychromatic to normochromatic erythrocytes). In addition, the shallow duration response curve seen for the 14 day versus 90 day toxicity ( Hayes et al., 1986) suggests that the cumulative dose is not an appropriate surrogate for a single dose at the LD 50. Drinking Water Criteria Document for Haloacetonitriles V  53 EPA/ OW/ OST/ HECD Final Draft In contrast to this result, other investigators have reported positive evidence for chromosome damage using less standard assay systems. Le Curieux et al. ( 1995) treated Pleurodeles waltl larvae ( newt) with HANs dissolved in the container water. Following 12 days of treatment, blood was collected and micronucleated erythrocytes were counted from a population of 1,000 erythrocytes. The range of concentrations tested, in µ g/ mL, were as follows:

BCAN, 0.0312 0.125; DBAN, 0.12 0.5; DCAN, 0.25 1; TCAN, 0.025 0.1. All the test compounds generated a positive result in this assay, with the lowest effective concentration for TCAN beginning at 0.1 µ g/ mL, for DBAN, and BCAN at 0.12 µ g/ mL, and for DCAN at 0.25 µ g/ mL. Although all of the compounds significantly increased micronuclei formation, the magnitude of this effect was characterized as relatively weak for DCAN and BCAN, since median increases in the number of micronucleated erythrocytes was on the order of 2 fold compared to controls. Based on results provided in a summary table, the maximum increase in the number of micronucleated erythrocytes for TCAN was also near 2 fold, while for DBAN increases were larger ( maximum of 6.17 fold).

In another assay for chromosome damage, DCAN at an inhalation concentration of 8.6 ppm induced aneuploidy in the offspring of female Drosophila melanogaster exposed for up to 45 minutes ( Osgood and Sterling, 1991). DBAN was highly toxic; a concentration of only 0.3 ppm killed 30 40% of flies after a 45 minute exposure, compared to 8.6 ppm for DCAN.

However, at a dose of 0.3 ppm, DBAN did not induce aneuploidy. Drinking Water Criteria Document for Haloacetonitriles V  54 EPA/ OW/ OST/ HECD Final Draft A variety of other studies have been conducted to test whether HANs can result in DNA damage. Bull et al. ( 1985) studied the ability of HANs to induce sister chromatid exchanges ( SCE), which measures a repair response to DNA damage. Chinese hamster ovary ( CHO) cells were treated with BCAN, DBAN, DCAN, or TCAN ( concentrations indicated in Table V 11)

without exogenous metabolic activation for 2 hours, at which time 5 bromo 2 deoxyuridine ( BrdU) was added to the medium and the incubation was continued for 26 to 32 hours. The cells with metabolic activation were treated with HANs for 2 hours in the presence of S9, at which time the cells were rinsed and fresh medium containing BrdU was added and the incubation was continued for 26 hours. The average SCE frequency was significantly elevated in the presence of nonactivated or S9 activated for BCAN, DBAN, DCAN, and TCAN. Comparisons of the potency of the HANs established the following: DBAN > BCAN > TCAN > DCAN.

Zimmermann et al. ( 1984) compiled and reviewed published results of genetic damage assays in Saccharomyces cerevisiae ( yeast). The only HAN compound for which data were presented was DCAN ( based on the study of Simmon et al., 1977; and Simmon and Kauhanen,

1978). Based on their analysis of the original report, Zimmerman and colleagues summarized the results of a homozygosity assay, which serves to indicate gene recombination and gene conversion events. The assay was conducted in a stationary culture of S. cerevesiae strain D3 with and without metabolic activation with liver microsomes. DCAN yielded a positive result in this assay Drinking Water Criteria Document for Haloacetonitriles V  55 EPA/ OW/ OST/ HECD Final Draft Table V 11. Induction of Sister Chromatid Exchange in Chinese Hamster Ovary Cells by BCAN, DBAN, DCAN, and TCAN. Drinking Water Criteria Document for Haloacetonitriles V  56 EPA/ OW/ OST/ HECD Final Draft regardless of metabolic activation. In a more complex assay system in yeast, Zimmermann and Mohr ( 1992) studied the effects of several agents, including DBAN, on mitotic chromosome loss and mitotic recombination. Diploid S. cerevisiae strain D61. M, heterozygous for three recessive alleles ( cyh2, leu1, and ade6) on chromosome VII, was used to test for effects of DBAN on chromosomal malsegregation or mitotic recombination. Chromosome loss was scored based on the number of colonies expressing all three recessive markers, and mitotic recombination was evaluated based on the expression of the chy2 and ade6 ( but not leu1 which was located between these two markers on the chromosome). Gene expression was identified by colony formation on the appropriate selective plates. The yeast cultures were treated with DBAN at concentrations ranging from 0 to 18.2 mg/ mL. Mitotic recombination was induced in a dose dependent fashion.

Chromosome loss was not the reason for the expression of the recessive markers. In contrast to this result, when yeast were treated with DBAN in combination with propionitrile, which is known to induce chromosome loss and enhances the sensitivity of the malsegregation analysis, the expected loss of chromosomes was observed. The authors speculated that failure to induce malsegregation with DBAN treatment alone reflects high toxicity at doses that would induce malsegregation.

The studies that have evaluated DNA damage at the chromosome level have resulted in inconsistent results. No increase in micronuclei was reported for any of the HANs in a standard assay for this end point ( Bull et al., 1985), but it is unclear whether high enough doses were tested. Positive results were reported for all four compounds in newt larvae, but this is not a wellcharacterized assay system. Positive results have generally been observed in assays that measure Drinking Water Criteria Document for Haloacetonitriles V  57 EPA/ OW/ OST/ HECD Final Draft responses to DNA damage, including sister chromatid exchange in CHO cells ( Bull et al., 1985)

and recombination studies in yeast ( Zimmermann et al., 1984; Zimmermann and Mohr, 1992).

Several studies have been conducted to evaluate DNA damage. Le Curieux et al. ( 1995)

studied the genotoxicity of BCAN, DBAN, DCAN, and TCAN in the SOS chromotest. This assay measures genotoxic activity, based on induction of the SOS DNA repair system ( measured by increased   galactosidase activity) in the Escherichia coli strain PQ37. The test was conducted with and without S9 activation. The range of doses tested, in µ g/ mL, was: BCAN, 3

3000 (+/ S9); DBAN, 3 1000 (+/ S9); DCAN 3 1000 (+/ S9); TCAN, 3 1000 (+ S9), 0.01 1000 ( S9). TCAN was negative up to cytotoxic concentrations. BCAN and DBAN were active beginning at 5 µ g/ mL and 10 µ g/ mL, respectively, in the absence of S9, but were inactive in the presence of S9 activation. DCAN generated a positive result in the presence of S9 beginning at 50 µ g/ mL, and was negative in the absence of S9. The results were dose dependent at the lower doses, but decreased at the higher concentrations, as the doses exceeded the threshold concentrations for cytotoxicity. The responses seen with BCAN, DBAN, and DCAN were considered weak, based on limited induction of   galactosidase activity.

Lin and colleagues in a series of papers have reported on the ability of HANs to induce direct DNA damage by measuring DNA strand breaks, the ability to bind to the nucleophilic agent ( 4 p nitrobenzyl pyridine), and formation of covalent DNA adducts ( reviewed in Lin et al., 1986).

Daniel et al. ( 1986) reported that HANs produced DNA strand breaks in cultured human lymphoblastic cells. The most potent initiator of DNA strand breaks compared to control cells Drinking Water Criteria Document for Haloacetonitriles V  58 EPA/ OW/ OST/ HECD Final Draft was TCAN which induced 2 fold more strand breaks than the positive controls. BCAN and DBAN had intermediate activity, while DCAN was described as having marginal activity. The protocol description did not clearly state the concentration of the HANs at which strand breaks were observed, but exposure to 50 µ M for 1 hour resulted in cell survival decreases ranging from 40% to 80% of control cells. The HANs also showed highly variable alkylation potential as measured by the potential to bind to 4 p nitrobenzyl pyridine ( Daniel et al., 1986). The relative alkylation potential of the HANs was DBAN> BCAN>> DCAN> TCAN. The range in reactivity toward 4 p nitrobenzyl pyridine varied by 627 fold among the HANs.

Daniel et al. ( 1986) also reported the results of DNA adduct analysis. In a cell free system, [ 14C] DCAN was incubated with calf thymus DNA, the DNA was precipitated,

hydrolyzed and separated by HPLC. Similar elution peaks were observed between the hydrolyzed calf thymus DNA and incubations of 14[ C] DCAN with polyadenylic acid or polyguanylic acid,

suggesting that DCAN forms an adduct with these nucleotides. Oral administration of DCAN or DBAN to rats did not result in detectable adduct formation in liver DNA ( no supporting data were presented) ( Lin et al., 1986). In a subsequent study, Lin et al. ( 1992) reported the formation of DNA adducts following gavage dosing of radiolabeled TCAN to male F344 rats. A single oral gavage dose of either [ 1 14C] or [ 2 14C] TCAN ( in tricaprylin, 7.2   69.3 mg/ kg) was administered to male F344 rats and tissues were analyzed from 2 to 48 hours following dosing.

TCAN bound to both DNA and proteins and DNA binding was highest in the stomach, followed by liver and kidney. Drinking Water Criteria Document for Haloacetonitriles V  59 EPA/ OW/ OST/ HECD Final Draft Overall, the data suggest that HANs can directly damage DNA as evaluated by a wide array of assays ( summarized in Table V 12). The weight of the evidence varies for each compound. BCAN has yielded positive results in all assays tested. DBAN yielded negative results in S. typhimurium mutation assays and failed to form DNA adducts in vivo. DBAN appears to induce DNA strand breaks and yield positive results in assays that reflect responses to DNA damage ( i. e. SCE, gene conversion, and SOS assays). The results for DCAN and TCAN are less consistent. DCAN yielded positive results in S. typhimurium mutation assays and assays reflecting DNA recombination, but the reason for absence of significant effects in the DNA strand break assay is not clear. For TCAN, the weak responses in S. typhimurium mutation assays did not correspond well with the observed in vivo formation of DNA adducts, although the positive results in the DNA strand break assay and SCE assay were consistent.

Table V 12. Summary of Results of Genotoxicity and Tumor Screening Assays for Haloacetonitriles.

BCAN DBAN DCAN TCAN Mutation assays ( S. typhimurium) +  + ± Micronuclei ± ± ± ± Aneuploidy ( D. Melanogaster) nt  + nt Sister Chromatid Exchange + + + +

Gene Conversion/ Recombination ( S. cerevisiae) nt + + nt SOS chromotest + + +

DNA Strand Breaks + +  +

DNA adducts ( in vivo) nt   +

Lung tumors ( A/ J mice) +   +

Skin tumors ( Sencar mice) + +

nt = not tested Drinking Water Criteria Document for Haloacetonitriles V  60 EPA/ OW/ OST/ HECD Final Draft The evidence for the induction of chromosome damage by HANs is less compelling, due to the limited number of studies available for evaluation and the inconsistent results. In the single study that used a standard assay protocol to evaluate induction of micronuclei, no effect was observed for any of the HANs, although is was not clear that sufficiently high doses were used.

In contrast, positive results for micronuclei formation were reported for all four compounds in a less well characterized newt larvae system. DCAN, but not DBAN, induced aneuploidy in the Drosophila melanogaster assay system.

E. Carcinogenicity No 2 year carcinogenicity bioassays have been conducted for any of the HANs by any route of exposure. No alternative carcinogenicity studies were identified for any the HANs by the inhalation route. There are, however, several short term assays that can aid in hazard identification. In addition, DBAN is currently under test in a full cancer bioassay by NTP ( 2002).

In a published review paper, Bull and Robinson ( 1985) reported studies on the incidence of lung tumors in groups of 40 female A/ J mice ( 10 weeks of age) that were administered a single oral dose of 10 mg/ kg of BCAN, DBAN, DCAN, or TCAN, three times weekly for 8 weeks ( Table V 13). Control groups received the vehicle only ( 10% emulphor) or ethyl carbamate ( positive control). As discussed in Chapter VII, emulphor solutions have generally been deemed as more appropriate solvent vehicles than corn oil for disinfectant byproducts. All animals were sacrificed at 9 months of age, allowing for an approximately 6 months post exposure observation period. The incidence of lung tumors ( adenomas) was significantly increased in groups given Drinking Water Criteria Document for Haloacetonitriles V  61 EPA/ OW/ OST/ HECD Final Draft BCAN and TCAN ( p

0.05). DBAN and DCAN produced marginal, but nonsignificant ( p >

0.05) increases in lung tumors. The authors stated that the results should be interpreted with caution, since there is a relatively large variation in the background incidence of lung tumors in this strain of mice and the 10 mg/ kg dose level was considerably below the maximum tolerated dose, decreasing the reliability of the negative findings with DBAN and DCAN.

Table V 13. Effects of Orally Administered Haloacetonitriles on the Development of Lung Adenomas in Female A/ J Mice.

Number of % Animals Tumors/ Chemical Dosea animals necropsied w/ tumors animal Vehicle ( emulphor 10%) 0.2 mL/ mouse x24 31 10 0.10 BCAN 10 mg/ kg x24 32 31 0.34b DBAN 10 mg/ kg x 24 31 16 0.19 DCAN 10 mg/ kg x 24 30 23 0.23 TCAN 10 mg/ kg x 24 32 28 0.38b Ethyl carbamate ( positive control) 42 mg/ kg x 24 29 100 9.00 a. Forty female strain A/ J mice were administered the indicated doses of each chemical three times weekly for a period of 8 weeks. Treatment was begun at 10 weeks of age. Animals were sacrificed at nine months of age.

b. Significantly increased above controls at P< 0.05.

Adapted from Bull and Robinson ( 1985).

Bull et al. ( 1985) studied the ability of BCAN, DBAN, DCAN, and TCAN to induce tumors in mouse skin the ability of dermally applied HANs to act as tumor initiators was studied using a tumor initiation/ promotion protocol. Six topical doses of 0, 200, 400, or 800 mg/ kg HAN dissolved in acetone were applied to the shaved backs of female Sencar mice ( 40 animals/ dose group) over a two week period, for total doses of 0, 1200, 2400 and 4800 mg/ kg. Beginning two Drinking Water Criteria Document for Haloacetonitriles V  62 EPA/ OW/ OST/ HECD Final Draft weeks after the last HAN dose, 1.0 µ g of 12 O tetradecanoylphorbol 13 acetate ( TPA) was applied three times per week for 20 weeks. Papilloma incidence and regression were recorded on a weekly basis. Animals were then maintained for 1 year and sacrificed to determine the incidence of squamous cell carcinomas. The results from this initiation/ promotion study are presented in Table V 14. These data were compiled by the study author from three independent experiments as indicated in the column of the table labeled " Experiment No".

Table V 14. Histopathological Diagnosis of Tumors Resulting from Topical Treatment with Halogenated Haloacetonitriles in the Sencar Mouse Treatment Experiment No. Total Dosea ( mg/ kg) Nb Number of animals with squamous cell tumors Squamous cell tumors diagnosed % Animals bearing carcinomas Number % Papilloma Carcinoma Acetone 1 2 3 0.2 ml X 6 0.2 ml X 6 0.2 ml X 6 34 37 34 1 3 5 3 8 15 0 3 1 1 0 4 2.9 0 11.7 BCAN 2 2 2 1200 2400 4800 35 37 37 1 7 8 3 19 22* 0 0 3 1 7 6 2.9 18.9* 16.2*

DBAN 1 2 3 2 3 1200 2400 2400 4800 4800 36 35 35 37 37 8 17 16 7 3 22* 49** 46** 19 11 6 9 7 5 1 2 8 9 2 2 5.6 22.9** 25.7** 5.4 7.4 DCAN 1 2 2 1200 2400 4800 39 35 35 4 4 1 10 11 3 4 2 1 0 3 0 0 8.6 0 TCAN 1 2 3 2 3 1200 2400 2400 4800 4800 34 36 38 36 29 2 11 1 3 2 6 31** 3 8 7 1 5 0 2 1 1 6 1 1 1 2.9 16.7* 2.6 2.8 3.4 Drinking Water Criteria Document for Haloacetonitriles V  63 EPA/ OW/ OST/ HECD Final Draft a Total doses of haloacetonitrile shown were delivered topically in six equal doses over a 2 week period. Then 2 weeks later animals were treated with 1.0 µ g 12 O tetradecanoylphorbol 13 acetate ( TPA) in 1.2 mL acetone topically 3 times weekly for 20 weeks. b There were 40 animals initiated in each group, N = number available for histological examination. * Significantly different from the control, p < 0.05, Fisher exact test. ** Significantly different from control, p < 0.01, Fisher exact test.

Adapted from Bull et al. ( 1985).

Both BCAN and DBAN induced dose related increases in the percent of animals with squamous cell tumors ( i. e., combined papillomas and carcinomas) ( during the first 24 weeks after TPA application was started) and the percent of animals with carcinomas ( at 1.5 years)( Table V

14). The study authors indicated that the decreased tumor response for DBAN at the high dose as compared to the low and mid doses may have resulted from the severe irritation and ulcerations induced by this treatment. For TCAN, an increase in the percent of animals with squamous cell tumors and in carcinomas was observed at the mid dose ( in experiment 2), but this increase was not replicated in experiment 3 at this dose. The combined data for TCAN did not yield a significant increase in tumor response. No significant increase in papillomas or carcinomas was observed with DCAN. A similar pattern of results for each of the HANs was obtained when the number of tumors/ animal was evaluated as the response metric. Based on these results, the authors concluded that DBAN is the most potent mouse skin tumor initiator of the HANs tested.

DBAN is followed in potency by BCAN, while TCAN and DCAN were judged as ineffective inducers of skin tumors in this screening bioassay.

Bull et al. ( 1985) conducted a second series of studies, designed to assess the ability of orally administered HANs to act as tumor initiators. In this study, total oral doses of 50 mg/ kg were administered to female Sencar mice six times over a 2 week period. The promotion phase Drinking Water Criteria Document for Haloacetonitriles V  64 EPA/ OW/ OST/ HECD Final Draft of the study was conducted using the same procedures as for the dermal dosing initiation/ promotion study described above. No statistically significant increase in tumor yield or decrease in the time to tumor was observed for any of the HANs when the data across all oral experiments were combined. Sporadic increases in tumor yield at various times for individual HANs were noted ( data not shown by the authors), but were not judged by the study authors to be biologically significant. The positive control, 300 mg/ kg urethane, increased the yield of papillomas.

In a third cancer screening study presented in the paper by Bull et al. ( 1985), the ability of dermally applied HANs to act as complete carcinogens was assessed. For this study, BCAN,

DCAN, or TCAN ( 800 mg/ kg), or DBAN ( 400 mg/ kg) was applied to the skin of female Sencar mice 3 times per week for 24 weeks, and the number of squamous cell tumors recorded. None of the HANs induced skin tumors in this assay ( data not shown by the authors).

DBAN, DCAN, and TCAN were tested for initiating activity using the rat liver gamma glutamyltranspeptidase foci ( GGT foci) assay in F344 rats as an indicator of carcinogenicity ( Herren Freund and Pereira, 1986; Lin et al., 1986). The protocol for this assay consists of a two thirds partial hepatectomy followed 18 or 24 hours later by the administration of the initiator. The doses employed were 2.0 mmol/ kg ( 398 mg/ kg) for DBAN, 2.0 mmol/ kg ( 220 mg/ kg) for DCAN, and 1.0 mmol/ kg ( 144 mg/ kg) for TCAN. Seven days after initiation, the rats received the promoter ( 500 ppm sodium phenobarbital in their drinking water) for at least ten weeks. The halogenated acetonitriles were inactive as initiators in the GGT foci assay. Drinking Water Criteria Document for Haloacetonitriles V  65 EPA/ OW/ OST/ HECD Final Draft The existing data provide at best only marginal support for the conclusion that HANs are carcinogenic. The evidence is stronger for BCAN, which increased tumor yields in both lung tumor and dermal screening assays. DBAN was positive at non ulcerative doses in the dermal screening assay. TCAN was positive only in the lung assay, and DCAN treatment did not increase either lung or skin tumors. Opposing these positive findings are the negative results for DBAN, DCAN, and TCAN in the GGT foci assay. Overall, the data are insufficient to qualitatively or quantitatively assess the carcinogenic potential of any of the HANs. The positive results in two tumor screening assays, together with positive bacterial gene mutation results,

suggest that it would be worthwhile to conduct a full 2 year bioassay for BCAN. Results for the other HANs are mixed, with inconsistencies between the genotoxicity and tumor screening data.

F. Summary The toxicity data on the HANs are summarized in Tables V 15 through V 18. Overall,

very little data are available evaluating the non cancer effects of the HANs. Acute oral LD 50 values for DBAN, DCAN, and TCAN in rodents have been reported to range from 50 to 361 mg/ kg. DBAN and TCAN have been reported to be irritants. DBAN causes eye, nasal, and respiratory tract irritation following inhalation, and skin irritation following dermal exposure.

TCAN also causes skin irritation following dermal exposure. No data on the acute toxicity of BCAN are available, and no subacute or subchronic studies of either BCAN or TCAN are available. No chronic studies have been conducted on any of the HANs. Drinking Water Criteria Document for Haloacetonitriles V  66 EPA/ OW/ OST/ HECD Final Draft No target organ has been clearly established for HANs following oral exposure, although absolute and relative organ weight changes, including decreased testes weight ( NTP, 2002) and increased liver weight ( Hayes et al., 1986; Christ et al., 1996) have been reported. Fourteen day or longer systemic toxicity studies have been conducted in mice and rats. In 14 day and 90 day studies of DBAN ( NTP, 2002; Hayes et al., 1986), consistent, compound related,

dosedependent effects were limited to decreased water consumption, decreased body weight, and decreased testes weight and pathology. However, effects on the testes reported in the NTP ( 2002) study were observed only in rats in the 14 day study. No effects on the testes were observed in rats in the 13 week study, or in mice ( NTP, 2002). In addition, no effects were observed on the testes in rats in a 14 day or 90 day gavage study in rats, even at much higher doses ( Hayes et al., 1986). For DBAN, the observed liver weight increases reported in Hayes et al. ( 1986) were not supported by other measures of liver toxicity in the same study or observed in the more recent NTP study ( NTP, 2002), and therefore, this endpoint was not selected as the basis for the quantitative dose response assessment. Taken together, the data suggest that decreased body weight appears to be the primary indicator of toxicity for DBAN. Overall, male rats appear to be more sensitive than female rats for DBAN. For DCAN, consistent,

compoundrelated dose dependent effects were limited to decreased body weight and increased liver weight ( Hayes et al., 1986). In this case, the observed liver weight increases were supported by changes in serum biochemistry parameters suggestive of liver damage. No histopathological evaluation was done in the key study for DCAN ( Hayes et al., 1986), so the degree, if any, of liver damage can not be confirmed. The data for TCAN and BCAN are too limited to identify with confidence any potential target organs. Drinking Water Criteria Document for Haloacetonitriles V  67 EPA/ OW/ OST/ HECD Final Draft The data are inadequate to determine whether HANs are reproductive toxicants. No multigeneration reproductive toxicity study has been conducted. BCAN, DBAN, DCAN, and TCAN at doses of up to 50 mg/ kg/ day had no effect on sperm morphology ( Meier et al., 1985),

but the data on testes weight changes are mixed ( NTP, 2002; Hayes et al., 1986). DBAN at doses up to approximately 10 mg/ kg/ day had no effect on any male or female reproductive parameter evaluated in a screening assay ( R. O. W Sciences, 1997). A series of developmental toxicity studies in rats has also been conducted. Exposure to BCAN and DBAN on gestation days 7 to 21 resulted in reduced mean birth weight ( Smith et al., 1986; Smith et al., 1987). It addition to this effect, DCAN and TCAN decreased the percentage of females delivering viable litters and increased fetal resorptions ( Smith et al., 1986; Smith et al., 1987; Smith et al., 1988;

Smith et al., 1989; Christ et al., 1996). DCAN and TCAN also significantly increase the frequency of malformations in fetuses ( Smith et al., 1988; Smith et al., 1989; Christ et al., 1996).

These studies by Smith and colleagues on the developmental toxicity of HANs in rats were conducted using tricaprylin as a vehicle, because these compounds are very miscible in this vehicle. However, in these studies, comparison of tricaprylin versus water treated controls revealed increased embryotoxicity due to tricaprylin. A recent study by Christ et al. ( 1996)

indicates that tricaprylin also influences the pattern of malformations observed in fetuses caused by TCAN. For TCAN in corn oil, the malformations were primarily cranio facial in nature while for TCAN in tricaprylin the malformations were primarily cardiovascular and urogenital in nature.

Therefore, the use of data from studies in which tricaprylin was used as the vehicle is not appropriate for risk assessment purposes. In the one developmental toxicity study that used a Drinking Water Criteria Document for Haloacetonitriles V  68 EPA/ OW/ OST/ HECD Final Draft vehicle other than tricaprylin ( Christ et al., 1996), maternal toxicity was observed at lower doses than developmental effects.