Hexachlorobenzene ( HCB) As a Contaminant of Pentachlorophenol Ecological Hazard and Risk Assessment for the Pentachlorophenol Reregistration Eligibiltiy Decision ( RED) Document 3/ 4/ 05 i TABLE OF CONTENTS Page No.

LIST OF TABLES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ii LIST OF APPENDICES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ii I. EXECUTIVE SUMMARY . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1 II. BACKGROUND/ INTRODUCTION . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2 III. TOXICITY OF HEXACHLOROBENZENE ( HCB) TO AQUATIC ORGANISMS . . . 3 A. Fish, Aquatic Invertebrates, and Plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3 B. Sediment Toxicity of HCB . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6 C. Bioaccumulation of HCB . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6 D. Conclusions Concerning Aquatic Toxicity of HCB . . . . . . . . . . . . . . . . . . . . . . . . . . 8 IV. TOXICITY OF HCB TO TERRESTRIAL ORGANISMS . . . . . . . . . . . . . . . . . . . . . . . 8 A. Toxicity Assessment for Invertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9 B. Toxicity Assessment for Birds and Bird Eggs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9 C. Toxicity Assessment for Mammals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11 D Conclusions Concerning Terrestrial Toxicity of HCB . . . . . . . . . . . . . . . . . . . . . . . 12 V. ECOLOGICAL RISK ASSESSMENT . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13 A. Risk to Terrestrial Organisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15 B. Risk to Aquatic Organisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17 C. Risk to Plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19 VI. ENDANGERED SPECIES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19 VII. UNCERTAINTIES IN THIS ASSESSMENT . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20 VIII. REFERENCES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21 ii LIST OF TABLES Page No.

Table 1. Selected Aquatic Toxicity Values for HCB . . . . . . . . . . . . . . . . . . . . . . . . . . 4 Table 2. Whole Sediment Acute Toxicity Values for HCB . . . . . . . . . . . . . . . . . . . . . 6 Table 3. Selected Avian Toxicity Values . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9 Table 4. Selected Mammalian Toxicity Values . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11 Table 5. Summary of HCB Concentrations in Air, Water, Sediments, and Soils, as Calculated in U. S. EPA ( 2002) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16 Table 6. HCB Acute and Chronic RQs for Freshwater Fish . . . . . . . . . . . . . . . . . . . . 17 Table 7. HCB Acute and Chronic RQs for Freshwater Invertebrates . . . . . . . . . . . . . 17 Table 8. HCB Acute RQ for Estuarine/ Marine Fish . . . . . . . . . . . . . . . . . . . . . . . . . 18 Table 9. HCB Acute RQ for Aquatic Plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19 LIST OF APPENDICES Appendix 1. Toxicity Data for HCB ( EPA, 1985   Health Assessment for Chlorinated Benzenes) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29 Appendix 2. Toxicity Values for Aquatic Animals and Plants ( EPA, 1988) . . . . . . . . . . . 30 Appendix 3. Aquatic Toxicity Data from ECOTOX Database ( EPA, 2001) . . . . . . . . . . 33 Appendix 4. Additional Aquatic Toxicity Data  Miscellaneous Sources . . . . . . . . . . . . . 35 Appendix 5. Terrestrial Toxicity Data ( Invertebrates, Birds, Mammals)   Ecotox Database, EPA 2001 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 36 Appendix 6. Terrestrial Toxicity Data  EPA 1985, Health Assessment for Chlorobenzenes ( Mammals Only) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 38 Appendix 7. Terrestrial Toxicity Data ( Mammals)  ATSDR, 2000   Toxicity Profile for HCB . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 40 Appendix 8. Terrestrial Toxicity Data   Miscellaneous Sources . . . . . . . . . . . . . . . . . . . 42 Appendix 9. Terrestrial Exposure Assessment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 43 1 I. EXECUTIVE SUMMARY Hexachlorobenzene ( HCB) is formed as a by product during the manufacture of pentachlorophenol and other chlorine containing compounds. HCB released to the environment tends to persist in soil and surface water for a long time, with recorded half life values ranging from 3 6 years and longer. This compound also readily bioaccumulates in both aquatic and terrestrial organisms.

Data on the ecological effects of HCB are relatively limited. There are no Guideline studies for HCB, since it is not registered as a pesticide. The majority of data presented in this report are based on information compiled in EPA technical reports ( U. S. EPA, 1985; U. S EPA, 1988; U. S. EPA, 2001) and the Agency for Toxic Substances and Disease Registry's Toxicological Profiles for HCB ( ATSDR, 2000).

Reviewed scientific literature indicates that HCB has a limited potential to adversely affect aquatic organisms in the short term, primarily due to its very low solubility in surface water. The majority of the toxicological data for fish, benthic invertebrates and aquatic plants did not indicate any lethal or significant sublethal effects from acute or chronic exposure to HCB at its saturation concentration of 6

g/ L. Some studies indicate that HCB concentrations in the tissues of aquatic organisms appear to equilibrate very slowly with concentrations in the water. As a result, the reviewed chronic toxicity tests for fish species ( e. g., rainbow trout and fathead minnows) may not have been of sufficient duration to allow for the full equilibration of HCB in fish tissue with surface water concentrations. Also, due to the tendency of HCB to bioaccumulate in the aquatic foodweb, there is the potential for adverse effects to higher trophic level organisms from exposure to HCB in their diet.

HCB is slightly to moderately toxic to terrestrial birds and mammals. LD50 values for evaluated avian species ranged from 568 mg/ kg for Japanese quail ( exposed to HCB over a 5 day period) to greater than 5000 mg/ kg for mallard ducks ( exposed to HCB for 8 days). LD50 values for mammals ranged from 710 mg/ kg for deer mice ( exposed to HCB over a three day period) to 4,000 mg/ kg for mice ( species and exposure duration not reported).

For chronic exposure, the lowest reported NOAELs for birds and mammals were 1 mg/ kg in Japanese quail ( over a 90 day period) and 1.6 mg/ kg BW per day in rats ( over a duration greater than 247 days), respectively. Sublethal effects reported for birds and mammals were often similar and included hepatic toxicity, porphyria, reproductive effects and reduced birth weight.

Hexachlorobenzene accumulated in birds is excreted into the egg, which results in uptake by the embryo. HCB concentrations measured in the eggs of sea birds and raptors from a number of locations from around the world approach those associated with reduced embryo weights in herring gulls ( 1.5 mg/ kg), suggesting that HCB has the potential to harm embryos of avian species. For mammals, a sensitive endpoint for chronic HCB exposure is the reduction of birth weight and increased mortality in mink offspring exposed to 1 ppm HCB ( 0.16 mg/ kg BW day) for 47 weeks. This observation is ecologically significant because field studies have observed HCB concentrations in fish tissue at a number of sites worldwide that are within an order of 2 magnitude of the dietary toxicity level of 1 ppm. This suggests that HCB has the potential to cause adverse effects in mink and perhaps other fish eating mammals, especially given HCB's tendency to bioaccumulate.

Reviewed literature indicate that HCB does have the potential to adversely affect both aquatic and terrestrial organisms, due to its persistence in the environment and its ability to readily accumulate in the aquatic and terrestrial foodwebs.

II. BACKGROUND/ INTRODUCTION HCB is a very stable chlorinated aromatic compound that was commonly used as a pesticide until 1965. Currently, there are no commercial uses of the substance in the United States. HCB may be formed as a byproduct during the manufacture of chemicals used as solvents, pesticides and other chlorine containing compounds. Small amounts of this compound can also be produced during combustion processes such as burning of city wastes.

HCB is widely distributed throughout the global ecosystem because if its mobility and resistance to degradation. It has been detected in all environmental media and in numerous types of living organisms including insects, aquatic biota, birds and mammals. HCB has also been shown to bioaccumulate in both aquatic and terrestrial organisms.

This report presents a compilation and evaluation of the available ecological effects data for HCB on aquatic life ( e. g., fish, arthropods, etc.), terrestrial invertebrates, birds and mammals ( e. g., rats, monkeys). The majority of the data presented in this report are based on compilations of data in U. S. EPA ( 1985), U. S. EPA ( 1988), the EPA Ecotox database ( U. S. EPA, 2001) and ATSDR ( 2000). Toxicity endpoint data have been provided in tabular format and a brief discussion of certain studies is also provided. In some cases, data from the same study are presented in multiple tables. The reader is referred to U. S. EPA ( 1985; 1988; 2001) and ATSDR ( 2000) for a more detailed discussion of the toxicity studies used therein. 3 III. TOXICITY OF HEXACHLOROBENZENE TO AQUATIC ORGANISMS A. Fish, Aquatic Invertebrates, and Plants While there are no EPA Guideline studies available for HCB, literature data provide some studies with comparable endpoints. The primary sources of aquatic toxicity data used for this report are described below.

EPA's final report entitled " Health Assessment Document for Chlorinated Benzenes" published in 1985 summarized the current knowledge of chlorinated benzenes in the environment ( including fate, transport and distribution) and current literature regarding toxicological effects of chlorinated benzenes on laboratory animals and humans. The report included an ecological section focused on the effects of chlorinated benzenes on aquatic organisms including fish, crustaceans and plants. Laboratory test concentrations and effects data reported in U. S. EPA ( 1985) are summarized for HCB in Appendix 1.

EPA's Draft 1988 Report entitled " Ambient Aquatic Life Water Quality Criteria for Hexachlorobenzene" developed by the Office of Research and Development, Environmental Research Laboratory evaluated the current scientific literature regarding acute and chronic toxicity data for HCB on aquatic animals and plants to develop a national water quality criteria for hexachlorobenzene, specifically for the protection of aquatic organisms. Acute and chronic toxicity data reported in U. S. EPA ( 1988) are summarized in Appendix 2.

EPA also maintains the ECOTOX database, which represents an integration of the AQUIRE, PHYTOTOX and TERRETOX databases that contain toxicological data for aquatic life, terrestrial plants and terrestrial wildlife, respectively. Summary toxicity data from ECOTOX regarding the effects of HCB on aquatic organisms are presented in Appendix 3. The data included in the appendix are only those data where a toxicity endpoint was provided. This database includes toxicity studies conducted through 2000, and therefore has more recent toxicity data than U. S. EPA 1985 and 1988.

In addition, aquatic toxicity data for HCB collected from other sources, such as scientific journals and technical reports, were also reviewed, and those data that were not already included in Appendices 1 3 are provided in Appendix 4.

Most of the reviewed aquatic toxicity studies for HCB have been conducted as acute or subchronic exposures. Selected acute and chronic endpoint values from Appendices 1 4 have been summarized and provided in Table 1. 4 Table 1. Selected Aquatic Toxicity Values for Hexachlorobenzene Species Test Type/ Duration Endpoint Concentration/ Effect Source Acute and Subchronic Toxicity Channel catfish ( Ictalurus punctatus) Static   24 hrs. LC50   6400

g/ L Mayer and Ellersieck, 1986 Bluegill ( Lepomis macrochirus) Static   duration not reported LC50   12,000

g/ L Johnson and Finley, 1980 Largemouth bass ( Micropterus salmoides) Static   96 hrs. LC50   12,000

g/ L Johnson and Finley, 1980 Coho salmon Static   96 hrs. LC50   50,000

g/ L Johnson and Finley, 1980 Flounder ( Platichthys flesus) 96 hrs. LC50   200

g/ L Furay and Smith, 1995 Sole ( Solea solea) 96 hrs. LC50   140

g/ L Furay and Smith, 1995 Largemouth bass ( Micropterus salmoides) 10 days 3.5

g/ L ( liver and kidney damage) Laseter et al., 1976 Diatom ( Cyclotella meneghiniana) Static  48 hrs. EC50  2

g/ L ( genetic effect) Figueroa et al., 1991 Freshwater green algae ( Selanastrum capricornutum) Static  96 hours EC50 > 30

g/ L Calamari et al., 1983 Chronic Toxicity Water flea ( Daphnia magna) 14 days EC50   16

g/ L ( 50percent reduction in fertility) Calamari et al., 1983 Cladoceran, ( Ceriodaphnia dubia) Life cycle/ partial life cycle   7 days NOEC > 7

g/ L Spehar, 1986 Rainbow trout ( Salmo gairdneri) Early life stage   90 days NOEC > 3.68

g/ L Spehar, 1986 Fathead minnow ( Pimephales promelas) Early life stage   32 days NOEC > 4.8f

g/ L Ahmad et al., 1984: Carlson and Kosian, 1987 Snail ( Lymnaea palustris) 10 12 weeks 0.5

g/ L, 5

g/ L ( significantly reduced growth rate) Baturo et al., 1995 Crab ( Portunus pelagicus) 6 weeks 5

g/ L ( reduction in carapace growth rate) Mortimer and Connell, 1995 HCB has a very low solubility in water ( 6

g/ L) ( ATSDR, 2000). As a result, the toxicity of HCB to aquatic organisms is generally considered a function of its solubility. Literature studies indicate that the only experimentally derived acute toxicity values for HCB were at concentrations that exceeded the solubility limit for this compound, in some cases by several orders of magnitude. For example, Mayer and Ellersieck ( 1986) determined that the LC50 for channel catfish exposed to HCB for 24 hours ranged between 6,400 and 16,200

g/ L. Similarly, Johnson and Findley ( 1980) reported that largemouth bass and coho salmon exposed 5 to HCB for 96 hours had LC50 values of 12,000

g/ L and 50,000

g/ L, respectively ( Table 1). They noted that the HCB concentrations for the experiments were obtained using a cosolvent to increase HCB's solubility.

Acute toxicity data compiled in Appendices 1 4 appear to support the conclusion that HCB concentrations that are within an order of magnitude of the saturation concentration of 6

g/ L do not appear to result in mortality to aquatic organisms. The lack of mortality is evident in the Appendices where the LC50s are consistently higher ( e. g., ">") than the maximum concentration of HCB evaluated in the study, with study concentrations ranging between 3.7

g/ L and 100

g/ L. The lowest identified LC50 concentration for acute toxicity at 96 hours exposure was approximately 200

g/ L for flounder and 140

g/ L for sole ( Furay and Smith, 1995) as shown in Table 1. These data indicate that ambient concentrations of HCB near the solubility limit are not anticipated to result in acute toxicity to the test species.

Some studies of acute exposures were continued longer than 96 hours with no resultant mortalities. For example, adult crayfish ( Procambarus clarki) were unaffected over a period of 10 days at a mean HCB concentration of 27.3

g/ L, and largemouth bass ( Micropterus salmoides) were unaffected over a period of 10 days at 25.8

g/ L ( Laseter et al., 1976 and Laska et al., 1978) ( Appendix 2). Two studies did, however, indicate sublethal effects for acute and subchronic exposure. As shown in Table 1, Laseter et al. ( 1978) found that largemouth bass exposed to a HCB concentration of 3.5

g/ L for 10 days did exhibit some liver and kidney damage. Also, Figueroa et al. ( 1991) found that diatoms exposed to HCB for 48 hours had an EC50 of 2

g/ L for genetic effects ( Table 1).

Chronic aquatic toxicity of HCB has also been investigated. Rainbow trout ( Salmo gairdneri) were exposed to HCB in a 90 day early life stage test ( Spehar, 1986). As shown in Table 1, no adverse effects on hatching, survival, or growth were observed at the highest tested concentration of 3.68

g/ L. Similarly, juvenile fathead minnows ( Pimephales promelas) were not affected at HCB exposures up to 4.8

g/ L in a 32 day early life stage test ( Ahmad et al., 1984; Carlson and Kosian, 1987) ( Table 1). Analysis determined that survival at hatch, survival at 32 days, and wet weight for the fathead minnows exposed to HCB were equivalent to the survival and wet weight for the control individuals. In a 7 day life cycle test with the Cladoceran ( Ceriodaphnia dubia), exposure to HCB did not result in any measurable effect upon survival or reproduction at concentrations as high as 7.0

g/ L ( Spehar, 1986) ( Table 1). These studies indicate that a HCB concentration of 3.68

g/ L does not appear to cause chronic toxicity in any of the tested species. However, based on review of the literature, two studies did indicate adverse effects on aquatic organisms from chronic exposure to HCB at concentrations below the solubility limit. Baturo et al. ( 1995) noted that snails ( Lymnaea palustris) exposed to HCB concentrations of 0.5

g/ L and 5

g/ L for a duration of 10 12 weeks had a significantly reduced growth rate compared with snails in a control group ( although the percentage difference in growth between the two groups was not reported). Similarly, Mortimer and Connell ( 1995) noted that crab carapace growth was reduced for juveniles exposed to HCB concentrations of 5

g/ L for a duration of 6 weeks compared with individuals in a control group, however, they did not estimate an EC10 or EC50 because a significant departure from the control group was only observed at one concentration. In 6 addition, Calamari et al. ( 1983) determined that water fleas ( Daphnia magna) exposed to HCB for 14 days had an EC50 of 16

g/ L for a reduction in fertility, although it should be noted that this concentration is higher than the solubility limit.

B. Sediment Toxicity of Hexacholorobenzene:

Although toxicity of HCB in surface water has been extensively studied, the toxicity of HCB absorbed to sediment is less clearly understood ( Barber et al., 1997). Organisms can be exposed to HCB through sediment ingestion, as well as through direct contact with surface water ( Kolok et al., 1996). Barber et al. ( 1997) evaluated potential effects to the amphipod ( Hyalella azteca) and midge ( Chironomus tentans) from exposure to HCB in sediment. The study results indicated that there were no adverse effects to survival or growth of the test species at the highest sediment concentration of HCB tested ( 42 mg/ kg). The authors noted that the lack of toxicity in this study even at the maximum HCB concentration is consistent with available literature on aquatic toxicity of this chemical. Fuchsman et al. ( 1998) also found that no significant toxicity occurred for the test species ( amphipod ( Hyalella azteca) and midge ( Chironomus tentans)) at a HCB sediment concentration of 60 mg/ kg, adding to the weight of evidence of limited potential for HCB related sediment toxicity to benthic invertebrates. They concluded that sediment quality guidelines such as the Ontario Ministry of the Environment's lowest effect level for HCB ( 0.02 mg/ kg) is likely to be overly conservative with respect to effects on benthic macroinvertebrates. A summary of sediment toxicity endpoints is provided in the table below.

Table 2. Whole Sediment Acute Toxicity of HCB to Aquatic Invertebrates Substance/ % Active Ingredient ( AI) Organism Endpoints/ Results ( mg/ kg) Comments Reference HCB Amphipod ( Hyalella azteca) NOEC = 42 14 day test duration ( spiked sediment) Barber et al., 1997 HCB Midge ( Chironomus tentans) NOEC = 42 14 day test duration ( spiked sediment) Barber et al., 1997 HCB Freshwater amphipod ( Hyalella azteca) NOEC = 60 10 day test duration ( spiked sediment) Fuchsman et al., 1998 HCB Midge ( Chironomus tentans) NOEC = 60 10 day test duration ( spiked sediment) Fuchsman et al., 1998 HCB Estuarine amphipod ( Leptochirus plumulosus) NOEC = 60 10 day test duration ( spiked sediment) Fuchsman et al., 1998 C. Bioaccumulation of Hexachlorobenzene:

Although the endpoint toxicity data for HCB in abiotic media appear somewhat limited, there is some evidence that bioaccumulation in lower trophic level species may result in effects to higher trophic level species. For example, Munoz et al. ( 1996) evaluated the effects of HCB 7 accumulation in algal tissue on the water flea ( Daphnia magna). They noted that uptake of HCB in algae ( Chlorella sp.) was rapid with an equilibrium time of 2 4 hours. The algae then had a HCB concentration of approximately 4.4 mg/ kg dry weight. Adult Daphnia fed the contaminated algae for six days had a hexachlorobenzene concentration of 1.7 mg/ kg dry weight. Although there was no mortality observed from HCB exposure, there was an effect on the reproductive capability in the Daphnia, with a 49 percent and 41 percent reduction in the numbers of juveniles produced and the total biomass of juveniles, respectively, that were produced during the time span of the experiment. Niimi and Cho ( 1980) similarly noted that uptake via diet might be more important than direct uptake for higher trophic level organisms, particularly in waters where HCB concentrations are low. Based on HCB's ability to bioaccumulate in aquatic organisms, Fuchsman et al. ( 1998) suggested that concentrations of HCB in surface water and sediment may pose a risk to higher trophic level organisms such as piscivorous wildlife and recommended that HCB studies focus on higher level species.

Several studies have been conducted to determine the equilibration time for HCB in aquatic organisms relative to the surrounding surface water concentration. According to EPA ( 1988), HCB concentrations in the tissues of many aquatic organisms appear to equilibrate very slowly with concentrations in the surrounding water. Oliver and Niimi ( 1983) reported that equilibrium had not yet occurred after 119 day exposure of rainbow trout to HCB at a concentration of 0.00032

g/ L, at which time the bioconcentration factor ( BCF) was 12,000. HCB uptake in fathead minnows was rapid during the first week of exposure, then residues in the minnows gradually increased through 115 days. HCB residues in the minnows represented a BCF of 16,200 at 32 days exposure and a BCF of 45,700 at 115 days exposure ( Veith et al., 1979). According to EPA ( 1988), HCB has been shown to have a relatively long half life in salmonids, particularly the liver. Rainbow trout that were fed HCB and maintained at a water temperature of 15oC had a half life ranging from 224 to 770 days ( EPA, 1988). The high accumulation levels of HCB in the fish tissue can be explained by the efficient uptake of HCB from water and food due to its high octanol water partition coefficient combined with the slow rate of conversion of HCB in fish tissue to water soluble metabolites and subsequent elimination.

The equilibration data provided in EPA ( 1988) are not consistent with Nebeker et al.' s ( 1989) investigation. Nebeler et al. ( 1989) that found that the equilibration time for HCB was somewhat shorter in duration. Nebeker et al. ( 1989) determined that a period of 28 days was adequate for the equilibration of HCB in the water flea and fathead minnow tissue relative to ambient water concentrations. They also noted that the tissue concentration in the test organisms dropped rapidly following removal of the HCB from the water. Also, Koneman and Van Leeuwen ( 1980) found an apparent steady state between HCB residues in guppy tissue and surrounding water concentration that occurred within 7 days ( with a BCF of approximately 16,000). Differences in ambient water concentrations of HCB and other test conditions may have contributed to the difference in HCB equilibration time observed for fathead minnows by Veith et al. ( 1979) of 115 days compared with Nebeker et al.' s ( 1989) result of 28 days. 8 If the equilibration time for HCB in certain organisms is of long duration as suggested by Veith et al. ( 1979), there is a possibility that certain chronic toxicity studies listed in Table 1 such as Spehar's ( 1986) 90 day study of rainbow trout and Ahmad et al ( 1984) and Carlson and Kosian's ( 1987) 32 day study of the fathead minnow may not be of sufficient duration to determine accurate chronic threshold values. Chronic toxicity tests with longer observation periods are needed to ensure that the test organisms are in equilibrium with the concentration of HCB in the surrounding surface water and sediment.

D. Conclusions Concerning Aquatic Toxicity of Hexachlorobenzene Based on reviewed scientific literature, surface water concentrations of HCB of 6

g/ L ( the solubility limit for this compound) do not appear to result in mortality to any tested species. Surface water concentrations of HCB at or below the solubility limit also have a limited potential to cause sublethal toxic effects in aquatic organisms for both subchronic and chronic exposure. Although certain studies did identify adverse effects from exposure to HCB at concentrations less than 6

g/ L, including liver damage in largemouth bass ( Laseter et al., 1976) and reduced growth rates in snails and crabs ( Baturo et al., 1995; and Mortimer and Connell, 1995, respectively), the majority of the data compiled in Appendices 1 4 do not indicate adverse effects to aquatic organisms for this range of concentrations. Similarly, toxicity studies did not identify sublethal adverse effects to benthic macroinvertebrates from exposure to HCB in sediment.

Some studies indicate that HCB concentrations in the tissues of aquatic organisms appear to equilibrate very slowly with concentrations in the water. Oliver and Niimi ( 1983) and Veith et al. ( 1979) found that HCB residue concentrations continued to increase in rainbow trout and fathead minnows, respectively, following 100 days of exposure to the compound. If HCB does require a relatively long duration to reach equilibrium, it is possible that chronic toxicity tests for certain species ( e. g., rainbow trout and fathead minnow) may not be of sufficient duration to accurately determine threshold concentrations. Also, because HCB tends to bioaccumulate in aquatic organisms, the potential for adverse effects to higher trophic level species, such as piscivorous wildlife, is a concern.

IV. TOXICITY OF HEXACHLOROBENZENE TO TERRESTRIAL ANIMALS Adverse effects to individual wildlife species from HCB have been documented in laboratory studies. Using the results of these studies to estimate effects on wild populations has limitations because the exposure route, medium of administration, and exposure duration of HCB for wild animals usually will differ from that of laboratory animals. Using these studies to assess effects on wild species assumes that the wild species are comparably sensitive to HCB as laboratory species; however, some species may be significantly more or less sensitive than laboratory species. The methodologies for predicting the effects of chemicals on terrestrial wildlife populations and ecosystems are still in development. Thus, in the absence of actual predictive methods for the population level, measures of the effects of HCB on reproduction 9 are currently the most useful indicators of possible effects on the populations of the species in the wild.

Toxicity data for terrestrial animals were compiled primarily from EPA's Ecotox database ( EPA, 2001), EPA's Health Assessment Document for Chlorinated Benzenes ( EPA, 1985), the Agency of Toxic Substances and Disease Registry Toxicological Profile for hexachlorobenzene ( ATSDR, 2000) and miscellaneous sources such as journal articles and technical reports. The toxicity data from these four general sources have been compiled in Appendices 5,6,7, and 8, respectively.

A. Toxicity Assessment for Invertebrates Toxicity data for terrestrial invertebrates were somewhat limited. Neuhauser ( 1985) conducted an acute toxicity test involving exposure of earthworms ( Eisenia fetida) to HCB applied in a water carrier ( single dose) with a test duration of 48 hours. As shown in Appendix 5, test results indicated that the LC50 exceeded 1000

g/ cm2. Similarly, Ballhorn ( 1984) found that earthworms exposed to HCB in artificial soil for a duration of 28 days had an LC50 greater than 1000 mg/ kg ( Appendix 5).

B. Toxicity Assessment for Birds and Bird Eggs The acute toxicity of HCB in birds was evaluated in several studies and toxicity data for certain endpoints ( LC50, NOEC) are provided in Table 2 below.

Table 3. Selected Avian Toxicity Values Species Test Method/ Duration Endpoint   Concentration Reference Acute Toxicity Ring necked Pheasant ( Phasianus colchicus) 5 Days LC50   617 ppm Hill et al., 1975 Japanese Quail 5 day LC50   568 ppm Hill, E. and M. Camardese, 1986 Bobwhite Quail NR LD50  575 mg/ kg HSDB, 1995 Mallard Duck NR LD50  1450 mg/ kg HSDB, 1995 Mallard Duck ( Anas platyrhynchos) 8 Days LC50 > 5000 ppm Hill et al., 1975 Herring gulls ( Larus argentatus) Egg injection study LD50   4300

g/ kg ( embryo mortality) Boersma et al., 1986 Herring gulls ( Larus argentatus) Egg injection study 1500

g/ kg   significant reduction in embryonic weight Boersma et al., 1986 Chronic Toxicity Japanese Quail ( Coturnix japonica) 90 Days LOEL   20 ppm ( egg volume) Vos, 1971 Japanese Quail ( Coturnix japonica) 90 Days NOEL   5 ppm ( egg volume) Vos, 1971 10 Japanese Quail ( Coturnix japonica) 90 Days LOEL   5 ppm ( morphology, organ weight) Vos, 1971 Japanese Quail ( Coturnix japonica) 90 Days NOEL   1 ppm ( morphology, organ weight) Vos, 1971 Hexachlorobenzene is considered slightly to moderately toxic to bird species. As shown in Table 2, LD50 values for evaluated avian receptors generally ranged between 568 mg/ kg and 1450 mg/ kg. Hill et al. ( 1975) found that mallard ducks exposed to HCB for 8 days did not exhibit any mortality at a concentration of 5000 ppm ( LC50 > 5000 ppm). This endpoint was somewhat higher than other reported values, possibly the result of interspecies variation or differences in specific test conditions. In regard to acute non lethal effects, Carpenter et al., ( 1985a) reported hepatic toxicity and porphyria in Japanese quail treated orally with 500 mg/ kg/ day hexachlorobenzene for 5 days ( Appendix 5), with most changes occurring after the first dose of hexachlorobenzene. Carpenter et al. ( 1985b) also conducted a sub chronic study of the effects of hexachlorobenzene Japanese quail. They determined that quail exposed to HCB for 17 days at a dose of 100 mg/ kg per day exhibited alterations in porphyrin level and changes in organ weight related to total body weight ( Appendix 5).

Chronic toxicity data for avian receptors in the reviewed literature were limited to Vos et al.' s ( 1971) investigation of Japanese quail ( Coturnix japonica) exposed to HCB. The investigation involved dietary administration of 0, 1, 5, 20 or 80 mg/ kg HCB in the diet for 90 days ( Vos. et al., 1971). They determined the NOEL to be 1 mg/ kg, at higher concentrations liver damage, porphyrin excretion and alteration of enzyme levels increased in a dose related fashion. Vos et al., ( 1971) also noted that there was also a dose related decrease in the hatchability of eggs, especially in groups treated at 20 mg/ kg for 90 days.

Notably, a significant portion of HCB residue in bird tissue is typically excreted in egg yolks following exposure to the compound ( Extension Toxicology Network, 2001). As a result, studies on the effects of HCB on wild birds have often focused on the accumulation of the compound in eggs and its subsequent effects on embryo survival and reproductive parameters. Egg hatchability and embryo survival appear to be two of the more sensitive endpoints for birds exposed to HCB. Boersma et al., ( 1986) conducted an egg injection study and found a significant increase in embryo mortality with increasing dose of HCB in herring gulls ( Larus argentatus), and calculated the LD50 to be 4300

g/ kg ( 4.3 mg/ kg) as shown in Table 2. Boersma et al. ( 1986) also noted that an egg concentration of 1500 µ g/ kg ( 1.5 mg/ kg) resulted in a significant reduction in embryonic weight ( Table 2).

Weseloh et al. ( 1983) also reported a link between HCB egg concentration and successful avian reproduction. They determined that 95 percent of Double breasted Comorants' eggs ( in a total sample size of 18 eggs) with a mean HCB concentration of 0.01 ppm wet weight failed to hatch, resulting in reproductive failure. Jarman et al. ( 1996) evaluated the relationship between organochlorine levels in prairie falcon eggs and egg hatchability for falcon eggs collected between 1989 and 1991. Their study results indicate that the high HCB levels observed in the falcon eggs combined with the possible additive effect of DDE may have had a deleterious effect on the hatching success of prairie falcons in portions of the study area. 11 C. Toxicity Assessment for Mammals The acute toxicity of HCB in mammals was evaluated in several studies and toxicity data for certain endpoints ( LC50, NOEC) are provided in Table 3 below.

Table 4. Selected Mammalian Toxicity Values Species Test Method/ Duration Endpoint   Concentration Reference Acute/ Subchronic Toxicity Deer Mouse ( Peromyscus maniculatus) 3 Days ( dosed one time per study period) LD50   710 mg/ kg Schafer and Bowles, 1985 Rat not reported LD50   3,500 mg/ kg Savitskii, 1964; 1965 Mouse not reported LD50   4,000 mg/ kg Savitskii, 1964; 1965 Rabbit not reported LD50   2,600 mg/ kg Savitskii, 1964; 1965 Cat not reported LD50   1,700 mg/ kg Savitskii, 1964; 1965 Rat Fo to F1a and F1b generations 21 day LD 50 values for pups were 100 and 140 mg/ kg for the F1a and F1b generations, respectively Kitchin et al., 1982 Rat 13 Weeks 19 mg/ kg/ day   lethality ( 4 of 9 individuals died) Den Besten et al., 1993 Rat 5 Weeks, ( dosed 1x per day, 5 days per week) LOAEL   1 mg/ kg/ day ( increased liver weight) Andrews et al., 1988 Rat 80 days LOAEL   100 mg/ kg/ day ( increased hepatic porphyrins) Cantoni et al., 1990 Rat 8 Weeks ( 1x per day) LOAEL   1000 mg/ kg/ day ( increased liver weight and hepatic porphyrins) Kleiman de Pisarev et al., 1990 Rat Chronic/ sub chronic   30 days LOAEL   1 g/ kg BW day ( 1000 mg/ kg BW day)   hepatotoxicity, porphyria, alteration of oestrus cycles. Alvarez, 2000 Chronic Toxicity Rat 2 Generations NOAEL   2 mg/ kg/ day ( reproductive effects) Arnold et al., 1985 Rat Chronic (> 247 days) NOAEL   1,600

g/ kg BW day ( reproductive effects) Grant et al., 1977 Rat 2 Generations LOAEL   0.016 mg/ kg/ day ( developmental effects, lymphocytosis and fibrosis of the liver) Arnold et al., 1985 Mink Chronic ( 47 weeks) LOAEL   1 ppm ( 0.16 mg/ kg BW day)   reduced birth weight and increased mortality Bleavins et al., 1984 12 The acute lethality of ingested HCB in terrestrial mammals is relatively low. As shown in Table 3, Savitskii ( 1964; 1965) determined that oral LD50 values for mammals exposed to HCB ranged from 1,700 mg/ kg for cats to 4,000 mg/ kg for mice; the duration for these studies was not reported. In a more recent study, Schafer and Bowles ( 1985) determined that the LD50 for deer mice exposed to HCB for 3 days was 710 mg/ kg.

Lethal levels in mammalian studies are progressively lower as the exposure duration is increased. For example, as shown in Table 3, Kitchin et al. ( 1982) found that rats exposed to HCB in utero and for 21 days following birth had an LD50 of 100 mg/ kg per day and Den Besten et al. ( 1993) found that 4 of 9 Wistar rats died following exposure to 19 mg/ kg hexachlorobenzene per day for a duration of thirteen weeks. Study data indicate that female animals are more susceptible to HCB induced mortality than males ( ATSDR, 2000).

Chronic toxicity effects of HCB in mammals have been provided in Appendices 6,7 and 8 and are summarized in Table 3. Some of the more conservative endpoints are associated with reproductive and developmental effects as shown in Table 3. Arnold et al. ( 1985) determined that 2 generations of rats exposed to HCB had a NOAEL of 2 mg/ kg/ day for reproductive effects. Similarly, Grant et al. ( 1977) found that rats exposed to HCB for a duration grater than 247 days had a NOAEL of 1.6 mg/ kg/ day for reproductive effects. For developmental effects, Arnold et al. ( 1985) found that rats exposed to HCB for 2 generations had a LOAEL of 0.016 mg/ kg/ day for lymphocytosis and fibrosis of the liver and Bleavins et al. ( 1984) determined that mink exposed to HCB for 47 weeks had a LOAEL of 0.16 mg/ kg/ day for reduced pup birth weight and increased mortality.

D. Conclusions Concerning Terrestrial Toxicity of Hexachlorobenzene HCB is considered slightly to moderately toxic to bird species. The lowest LC50 dose for avian receptors was 575 mg/ kg for bobwhite quail. Sublethal effects reported for HCB include hepatic toxicity, porphyria and alteration of enzyme levels. The lowest NOAEL for HCB exposure to birds is 1 ppm ( mg/ kg) for Japanese quail over a 90 day period.

Scientific literature indicates that HCB absorbed in avian tissue tends to be excreted in bird eggs. As a result, several laboratory and field investigations have focused on this toxicological endpoint. The lowest reported endpoint for HCB in bird eggs in laboratory studies was 1.5 mg/ kg in herring gull eggs, which caused a significant reduction in embryonic weight. Notably, HCB concentrations observed in the eggs of sea birds and raptors from a number of locations from around the world approach this level, suggesting that HCB could potentially cause an adverse affect on the reproduction of bird species.

The acute lethality of ingested HCB in terrestrial mammals is relatively low. Reviewed LD50 values range from 710 mg/ kg for deer mice ( over a 3 day exposure period) to 4,000 mg/ kg for mice ( species and exposure duration not reported). The toxicity of HCB in mammal studies increases with increasing exposure duration. Sublethal effects to mammals from HCB exposure 13 include liver damage, alteration of oestrus cycles and other reproductive effects and reduced birth weight.

For chronic exposure, the lowest reported NOAEL was 1.6 mg/ kg BW per day in rats ( over a duration greater than 247 days) for reproductive effects. Another sensitive endpoint is the reduction of birth weight and increased mortality in mink offspring following chronic exposure to 1 ppm HCB ( 0.16 mg/ kg BW day) for a duration of 47 weeks. Field observations of HCB concentrations in fish tissue at a number of sites worldwide that are within an order of magnitude of 1 ppm, suggesting that HCB has the potential to cause adverse effects in mink and perhaps other fish eating mammals.

V. Ecological Risk Assessment Risk assessment integrates the results of the exposure and ecotoxicity data to evaluate the likelihood of adverse ecological effects. One method of integrating the results of exposure and ecotoxicity data is called the quotient method. For this method, risk quotients ( RQs) are calculated by dividing exposure estimates by ecotoxicity values, both acute and chronic:

RQ = EXPOSURE/ TOXICITY RQs are then compared to AD's levels of concern ( LOCs). These LOCs are criteria used by OPP to indicate potential risk to nontarget organisms and the need to consider regulatory action. The criteria indicate that a pesticide used as directed has the potential to cause adverse effects on nontarget organisms. LOCs currently address the following risk presumption categories: ( 1) acute high  potential for acute risk is high regulatory action may be warranted in addition to restricted use classification; ( 2) acute restricted use  the potential for acute risk is high, but this may be mitigated through restricted use classification; ( 3) acute endangered species  the potential for acute risk to endangered species is high, and regulatory action may be warranted, and ( 4) chronic risk  the potential for chronic risk is high, and regulatory action may be warranted. Currently, AD does not perform assessments for chronic risk to plants, acute or chronic risks to nontarget insects, or chronic risk from granular/ bait formulations to mammalian or avian species.

The ecotoxicity test values ( i. e., measurement endpoints) used in the acute and chronic risk quotients are derived from the results of required studies. Examples of ecotoxicity values derived from the results of short term laboratory studies that assess acute effects are: ( 1) LC50 ( fish and birds) ( 2) LD50 ( birds and mammals ( 3) EC50 ( aquatic plants and aquatic invertebrates) and ( 4) EC25 ( terrestrial plants). Examples of toxicity test effect levels derived from the results of longterm laboratory studies that assess chronic effects are: ( 1) LOEC ( birds, fish, and aquatic invertebrates) ( 2) NOEC ( birds, fish and aquatic invertebrates) and ( 3) MATC ( fish and aquatic invertebrates). For birds and mammals, the NOEC value is used as the ecotoxicity test value in assessing chronic effects. Other values may be used when justified. Generally, the MATC ( defined as the geometric mean of the NOEC and LOEC) is used as the ecotoxicity test value in 14 assessing chronic effects to fish and aquatic invertebrates. However, the NOEC is used if the measurement endpoint is production of offspring or survival.

Risk presumptions, along with the corresponding RQs and LOCs are tabulated below.

Risk Presumptions for Terrestrial Animals ( Birds and Wild Mammals)

Risk Presumption RQ LOC Acute High Risk EEC1/ LC50 or LD50/ sq ft2 or LD50/ day3 0.5 Acute Restricted Use EEC/ LC50 or LD50/ sq ft or LD50/ day ( or LD50 < 50 mg/ kg) 0.2 Acute Endangered Species EEC/ LC50 or LD50/ sq ft or LD50/ day 0.1 Chronic Risk EEC/ NOEC 1 1 abbreviation for Estimated Environmental Concentration ( ppm) on avian/ mammalian food items 2 mg / ft LD50 * wt. of bird 2 3 mg of toxicant consumed / day LD50 * wt. of bird Risk Presumptions for Aquatic Animals Risk Presumption RQ LOC Acute High Risk EEC1/ LC50 or EC50 0.5 Acute Restricted Use EEC/ LC50 or EC50 0.1 Acute Endangered Species EEC/ LC50 or EC50 0.05 Chronic Risk EEC/ MATC or NOEC 1 1 EEC = ( ppm or ppb) in water 15 Risk Presumptions for Plants Risk Presumption RQ LOC Terrestrial and Semi Aquatic Plants Acute High Risk EEC1/ EC25 1 Acute Endangered Species EEC/ EC05 or NOEC 1 Aquatic Plants Acute High Risk EEC2/ EC50 1 Acute Endangered Species EEC/ EC05 or NOEC 1 1 EEC = lbs ai/ A 2 EEC = ( ppb/ ppm) in water A. Exposure and Risk to Nontarget Terrestrial Animals Terrestrial organisms may be exposed to HCB through direct contact with treated lumber and contact with HCB in soil that has leached from treated lumber. Exposure and toxicity data for direct contact with treated lumber were not readily available for terrestrial organisms. As a result, the terrestrial assessment was based on exposure to HCB in soil.

Risks to mammalian and avian species were evaluated using a simple terrestrial food web model. To estimate exposure for receptor species, it was assumed that HCB leaching occurs from treated wood into the surrounding soil. The leached HCB in soil was then assumed to be taken up by terrestrial vegetation. Based on these conditions, the model assumed that birds and mammals would be exposed to HCB through ingestion of vegetation and incidental ingestion of soil.

The estimated environmental concentration ( EEC) for soil used in the terrestrial risk assessment was determined by modeling described in U. S. EPA's " Environmental Fate Modeling of Hexachlorobenzene in Technical Grade Pentachlorophenol" ( EPA, 2002). Using the state of Ohio to represent a typical regional environmental setting, U. S. EPA calculated the amount of HCB released from an estimated 1,500,000 utility poles within the state. Based on the calculated release rates of HCB via volatilization, wood erosion, and leaching, the fate and transport of HCB releases was modeled using the Level III model. The HCB released was assumed to partition into the air, soil, water, and sediments. According to the results, terrestrial organisms would be exposed to 2.58 mg/ kg of HCB in the soil. 16 Table 5. Summary of HCB Concentrations in Air, Water, Sediments, and Soils, as Calculated in U. S. EPA ( 2002). Concentration Air ( g/ m3) 1.39E 12 Water ( mg/ L) 1.76E 11 Sediment ( µ g/ g) 8.13E 08 Soils ( mg/ kg) 2.58E 08 The methods used to estimate the potential dose of HCB to avian and mammalian species through the ingestion of vegetation and soil are provided in Appendix 9. The results of the exposure assessment for avian and mammalian species have been provided in Tables A1 and A2 ( in Appendix 9).

Generally, the hierarchy of sources for selecting which toxicity value to use when more than one was identified is: 1) core studies from the pesticide database, 2) supplemental studies from the pesticide database, 3) ECOTOX studies, and 4) studies from the open literature. However, since the original studies were not available for review, when more than one toxicity reference value was available, the toxicity data that had the most complete study information ( e. g., reported study duration or toxicity endpoint defined) and had the lowest toxicity value was selected in order to be conservative.

i. Birds As shown in Table A1, the risks to avian species from exposure to HCB in soil as a result of HCB leaching from treated lumber does not appear to be significant. The RQ of 4.3E 13 for acute exposure to HCB was much less than all LOCs. The chronic RQ for HCB, 2.5E 10, was also much less than the LOC.

ii. Mammals Based on the results in Table A2, exposure to concentrations of HCB in soil would not result in adverse effects to mammalian species. The RQ for acute exposure to HCB, 1.3E 12, was below all the defined LOCs. Similarly, the chronic RQ for HCB estimated using the predicted soil concentration, 9.9E 10, was much less than any of the LOCs.

iii. Conclusions of the Risk Assessment for Birds and Mammals:

The results indicate that risks to birds and mammals exposed to concentrations of HCB in soil is low. It should be noted, however, that the risk assessment was only based on exposure to HCB components in soil. A quantitative assessment of the risks to birds and mammals from direct contact with HCB treated lumber was not conducted due to the lack of exposure and toxicity data available. As a result, the potential risks from direct contact with HCB treated wood were not evaluated. Additional uncertainties associated with the assessment are discussed in the Uncertainty section of the report. 17 iv. Insects Exposure to HCB in treated lumber and the surrounding soil may result in adverse effects to insects such as honey bees. The potential risks to insects could not be quantitatively evaluated, however, since toxicity and exposure data for insects were not readily available.

B. Exposure and Risk to Nontarget Freshwater and Marine/ Estuarine Aquatic Organisms Nontarget freshwater and marine/ estuarine aquatic organisms could potentially be exposed to HCB leached from treated wood into the aquatic environment. RQs were calculated for acute and chronic effects on aquatic organisms using EECs for surface water. The HCB EEC in water, 1.76E 11 mg/ L was determined by modeling the fate and transport of HCB released from utility poles in the environment and detailed in U. S. EPA ( 2002).

Using this EEC, RQ values were calculated and the results are presented in the following tables. Both acute and chronic RQs were estimated using the predicted surface water EEC. As previously noted, when more than one toxicity reference value was available, generally, the lowest value was selected in order to be conservative.

i. Freshwater Fish Acute and chronic RQs for freshwater fish are tabulated below. The lowest toxicity ( LC50) values were used for HCB in order to be conservative.

Table 6. HCB Acute and Chronic Risk Quotients for Freshwater Fish Acute Exposures Chronic Exposures LC50 ( mg/ L) EEC ( mg/ L) Acute RQ ( EEC/ LC50) NOEC ( mg/ L) EEC ( mg/ L) Chronic RQ ( EEC/ NOEC) HCB 6.4 ( Mayer and Ellerseick, 1986) 1.76E 11 2.8E 12 ( a) 3.68 ( Spehar, 1986) 1.76E 11 4.8E 12 ( a)

a = none of the LOCs have been exceeded.

Results indicate that none of the RQs exceed any LOC.

ii. Freshwater Invertebrates Acute and chronic RQs for aquatic invertebrates have been tabulated below. The lowest toxicity ( LC50) value was used in order to be conservative. 18 Table 7. HCB Acute and Chronic Risk Quotients for Freshwater Invertebrates Acute Exposures Chronic Exposures EC50 ( mg/ L) EEC ( mg/ L) Acute RQ ( EEC/ EC50) NOEC ( mg/ L) EEC ( mg/ L) Chronic RQ ( EEC/ NOEC) HCB 0.005 ( Laseter et al., 1976) 1.76E 11 3.5E 09 ( a) 0.007 ( Spehar, 1986) 1.76E 11 2.5E 09 ( a)

a = none of the LOCs have been exceeded.

The RQs based on the EEC values from the leachate study indicate that the none of the LOCs are exceeded for acute or chronic exposures.

iii. Estuarine and Marine Fish Acute risk quotients ( RQs) for estuarine and marine fish are tabulated below. The lowest toxicity ( LC50) value was used in order to be conservative.

Table 8. HCB Acute Risk Quotient for Estuarine/ Marine Fish Acute Exposures LC50 ( mg/ L) EEC ( mg/ L) Acute RQ ( EEC/ LC50)

HCB 0.14 ( Parrish et al., 1976) 1.76E 11 1.3E 10 ( a) a = none of the LOCs have been exceeded.

The calculation indicates that the acute RQ did not exceed any LOCs.

No chronic toxicity data for marine/ estuarine fish or invertebrate species were available. Therefore, risk quotients could not be calculated for these organisms. Acute fish toxicity data indicate that marine/ estuarine fish species are more sensitive to HCB than freshwater species; therefore, the chronic risk posed to these species by HCB may be greater than the chronic risks posed to freshwater species. However, the chronic RQs developed for freshwater species are well below the LOCs, and even estimating a two order of magnitude difference ( based on the freshwater fish acute RQ compared to the marine/ esutarine fish acute RQ), the chronic RQ for marine/ estuarine species would likely remain below the LOC.

iv. Conclusions of the Risk Assessment for Aquatic Animals:

Results of the risk assessment indicate that the potential risk to aquatic animals exposed to HCB leachate in surface water would be low. Both the acute and chronic RQs were significantly below their respective LOQs for each group of aquatic animals evaluated. However, as discussed in the hazard section of this assessment, HCB has a strong tendency to bioaccumulate, and could therefore potentially cause adverse effects to aquatic organisms over time, and could possibly cause adverse effects to higher trophic level organisms which consume HCB contaminated aquatic organisms. 19 C. Exposure and Risk to Plants i. Terrestrial Plants The use of HCB as a wood preservative may result in the leaching of HCB into the surrounding soil. However, no terrestrial plant studies were identified and, thus, the risks to terrestrial plants could not be quantitatively evaluated.

ii. Aquatic Plants Acute risk quotients ( RQs) for aquatic plants are tabulated below. The lowest toxicity value was used in order to be conservative.

Table 9 HCB Acute Risk Quotient for Aquatic Plants Acute Exposures EC50 ( mg/ L) EEC ( mg/ L) Acute RQ ( EEC/ EC50)

HCB 0.003 ( Wong et al., 1984) 1.76E 11 8.8E 09 ( a) a = none of the LOCs have been exceeded.

The RQ based on the EEC value for HCB leached from utility poles is below all LOCs.

VI. Endangered Species The results of the risk assessment indicate that threatened and endangered species would not be expected to be adversely affected directly by exposure HCB. However, as discussed above, the strong tendency of HCB to bioaccumulate could lead to secondary adverse effects to higher trophic level organisms, or direct effects to organisms exposed to HCB over longer periods of time.

The Agency has developed a program ( the Endangered Species Protection Program) to identify pesticides whose use may result in adverse impacts to endangered and threatened species, and to implement mitigation measures that will eliminate the adverse impacts. At present, the program is being implemented on an interim basis as described in a Federal Register notice ( 54 FR 27984 28008, July 3, 1989), and is providing information to pesticide users to help them protect these species on a voluntary basis. As currently planned, the final program will call for label modifications referring to required limitations on pesticide uses, typically as depicted in countyspecific bulletins or by other site specific mechanisms as specified by state partners. A final program, which may be altered from the interim program, will be described in a future Federal Register notice. The Agency is not imposing label modifications at this time through the RED. Rather, any requirements for product use modifications will occur in the future under the Endangered Species Protection Program. 20 VII. Uncertainties in the Risk Assessment All of the toxicity values used in the assessment were from the ECOTOX database or the open literature rather than submitted studies that meet the pesticide guideline requirements, as HCB is not a registered pesticide. Testing conditions in these studies were likely quite variable, which causes greater variability across test results.

Another source of uncertainty in the toxicity data used in the assessment is that most of the toxicity studies were conducted in a laboratory environment, and therefore may not accurately reflect field conditions experienced by wild organisms. Toxicity studies also frequently use different terrestrial or aquatic species than those expected to occur in the natural environment. The comparative sensitivity between wild organisms and laboratory organisms is not well known, and, therefore, some wild species may be substantially more sensitive to HCB than is indicated by the toxicity endpoints used in this assessment.

There are several uncertainties associated with the exposure assessment for HCB in terrestrial and aquatic environments. The terrestrial risk assessment assumed that birds and mammals were only exposed to HCB in soil, since toxicity and exposure data were not readily available for direct contact with pentachlorophenol treated wood. Likewise, the aquatic risk assessment assumed that aquatic organism exposure to HCB would occur only from leaching from pentachlorophenoltreated wood. As discussed in the separate Environmental Fate Modeling of HCB chapter, volatilization, wood erosion and leaching all contribute HCB to the environment. The combined amount of HCB released from pentachlorophenol treated wood, rather than the amount released from one specific pathway, would provide a more accurate estimate of the levels to which terrestrial and aquatic organisms are exposed. The model also included many assumptions regarding the indicator species, such as diet, food ingestion rate, and home range size, that may not accurately reflect actual conditions in the environment. 21 VIII. REFERENCES:

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G. Zaroogian. 1981. Interlaboratory Comparison   Acute Toxicity Tests Using the 48 Hour Oyster Embryo Larval Assay. U. S. EPA, Narragansett, RI: 17. 30 Appendix 1. Toxicity Data for Hexachlorobenzene ( EPA, 1985   Health Assessment for Chlorinated Benzenes)

Species Exposure Duration/ Type Endpoint  Concentration Reference Fish Acute/ Subchronic Toxicity Data Largemouth black bass ( Micropterus salmoides) 10 days ( static)

15 days ( static) NOAEL 10

g/ L NOAEL 26

g/ L Laska et al., 1978 Laska et al., 1978 Sheepshead minnow ( C. variegatus) 96 hrs. ( constant flow) NOAEL 0.13 mg/ L Parrish et al., 1974 ( conc were nominal actual conc were probably lower) Pinfish ( Lagodon rhomboides) 96 hrs. ( constant flow) NOAEL 1.0 mg/ L Parrish et al., 1974 ( conc were nominal actual conc were probably lower) Rainbow trout ( S. gairdneri) 48 hrs. LC50 > 0.03 mg/ La Calamari et al., 1983 Brachydanio rerio 48 hrs. LC50 > 0.03 mg/ La Calamari et al., 1983 Aquatic Crustaceans   Acute Toxicity Data Crayfish ( Procambarus clarkii) Duration unspecified ( static or flow through system) 0.02 mg/ L   NOAEL Laska et al., 1978 ( saturated aqueous solution) Shrimp ( Crangon septemspinosa) 96 hr. ( static) 0.0072 ( mg/ L)   no mortality McLeese and Metcalfe, 1980 Water flea ( Daphnia magna) 24 hr. IC50 > 0.03 mg/ La ( immobilization concentration for 50 percent of animals) Calamari et al., 1983 Aquatic Plants   Acute Toxicity Data Freshwater algae ( Selenastrum capricornatum) 96 hr. ( static) EC50 > 0.03 mg/ L Calamari et al., 1983 Mixed diatoms and green algae ( Thalassiosira pseudonana, Dumaliella tertiolecta) 72 hrs. ( static) 0.1 mg/ L NOEC   no growth inhibition Biggs et al., 1979 Green algae ( Chlorella pyrenoidosa) 76 hrs. 10.0 mg/ L   Growth reduction ( growth reduced to 87.5 percent of control cultures) measured by dry mass Parasher et al., 1978 a Although the endpoint values numbers were reported in some sources as less than ("<") the specific test concentration, according to the text of the actual Calamari et al. ( 1983) study, it appears that the endpoints were greater than the actual concentration ( e. g., > 0.03 mg/ L) 31 Appendix 2. Toxicity Values for Aquatic Animals and Plants ( EPA, 1988)

Species Test Method / Durationa LC50 or EC50 (

g/ L) b Reference Acute/ Subchronic Toxicity Hydra ( adult) static, measured/ ( duration not reported) > 65.9 Sabourin et al. 1986 Leech ( adult) ( Nephelopsis obscura) static, measured/ ( duration not reported) > 73.2 Sabourin et al. 1986 Snail ( adult) ( Aplexa hypnorum) static, measured/ ( duration not reported) > 77.6 Sabourin et al. 1986 Cladoceran ( Ceriodaphnia dubia) Static, measured/ ( 7 days) EC50 > 7.0 Spehar 1986 Cladoceran < 24 hr. ( Daphnia magna) 28.3 hr > 6 ( EC10) ( immobilization) Sugatt et al., 1984 Cladoceran ~ 5 days ( Moina macrocopa) 3 hr > 6 ( LC50) Yoshioka et al., 1986 Amphipod ( adult) ( Gammarus pseudolimnaeus) Static, measured/ ( duration not reported) > 77.6 Sabourin et al. 1986 Crayfish ( adult) ( Procambarus clarki) Flow through, measured/ ( 10 days) LC50 > 27.3c Laseter et al., 1976 Laska et al., 1978 Crayfish ( juvenile) ( Procambarus sp.) Static, measured/ ( duration not reported) LC50 > 5.2d Laseter et al., 1976 Laska et al., 1978 Stonefly ( nymph) ( Pteronarcys sp.) Static, measured/ ( duration not reported) > 69.7 Sabourin et al. 1986 Midge ( 3rd 4th instar), ( Tanytarsys dissimilis) Static, measured/ ( duration not reported) > 58.1 Call et al. 1983 Rainbow Trout ( juvenile) ( Salmo gairdneri) Flow through, measured/ ( duration not reported) > 3.76 Spehar, 1986 Rainbow trout ( Salmo gairdneri) 24 hr LC50 > 30 Calamari et al., 1983 Fathead minnow ( Pimephales promelas) Flow through, measured/ ( duration not LC50 > 6e Veith et al., 1983a, b Ahmad at al., 1984 Carlson and Kosian, 1987 Species Test Method / Durationa LC50 or EC50 (

g/ L) b Reference 32 reported) Fathead minnow ( 0.7 g) ( Pimephales promelas) Static, unmeasured ( duration not reported) LC50 > 10,000 Johnson and Finley 1980, Mayer and Ellersieck 1986 Appendix 2. Toxicity Values for Aquatic Animals and Plants ( EPA, 1988) ( Cont'd)

Species Test Method / Durationa LC50 or EC50 (

g/ L) b Reference 33 Channel catfish ( 0.8g) ( Ictalurus punctatus) Static, unmeasured ( 24 hrs.) LC50  13,500 Johnson and Finley 1980, Mayer and Ellersieck 1986 Channel catfish ( 1.3g) ( Ictalurus punctatus) Static, unmeasured ( duration not reported) LC50 > 100,000 Mayer and Ellersieck 1986 Channel catfish ( sac fry) ( Ictalurus punctatus) Static, unmeasured ( 24 hrs.) LC50  7,000 Mayer and Ellersieck 1986 Bluegill ( 1.5g) ( Lepomis macrochirus) Flow through, measured/ ( duration not reported) > 78.4 Call et. al, 1983 Bluegill ( 1g) ( Lepomis macrochirus) Static, unmeasured ( duration not reported) LC50  12,000 Johnson and Finley 1980, Mayer and Ellersieck 1986 Bluegill ( 1.6g) ( Lepomis macrochirus) Flow through, unmeasured ( duration not reported) LC50 > 1,000 Mayer and Ellersieck 1986 Largemouth bass ( Micropterus salmoides) Flow through, measured/ ( 10 days) LC50 > 25.8e Laseter et al., 1976 Laska et al., 1978 Largemouth bass ( 0.5g) ( Micropterus salmoides) Static, unmeasured   96 hrs. LC50  12,000 Johnson and Finley 1980, Mayer and Ellersieck 1986 Largemouth bass 10 days 3.5

g/ L ( liver and kidney damage) Laseter et al., 1976 Coho salmon Static  96 hrs. LC50  50,000 Johnson and Finley, 1980 Guppy ( 2 3 mo.) ( Poecilia reticulata) 14 days > 320 ( LC50) Konemann 1979, 1981 Green alga ( Ankistrodesmus falcatus) 4 hr > 3 ( EC50) Wong et al., 1984 Protozoan ( Colpidium compylum) 43 hr 10,000 ( No effect) Dive et. al., 1980 Protozoan ( Tetrahymena pyriformis) 24 hr > 6 ( EC50) Yoshioka et al., 1985 Cladoceran ( Daphnia magna) 24 hr > 30 ( EC50) Calamari et al., 1983 Alga, ( Scenedesmus subspicatus) 4 day duration > 10 ( EC10) Geyer et al., 1985 Alga, ( Scenedesmus subspicatus) 4 day duration > 10 ( EC50) Geyer et al., 1985 Appendix 2. Toxicity Values for Aquatic Animals and Plants ( EPA, 1988) ( Cont'd)

Species Test Method / Durationa LC50 or EC50 (

g/ L) b Reference 34 Alga, ( Selenastrum capricornatum) 4 day duration > 30 ( EC50) Calamari, 1983 CHRONIC TOXICITY Cladoceran, ( Ceriodaphnia dubia) Life cycle/ partial life cycle   7 days > 7f ( NOEC) Spehar, 1986 Rainbow trout ( Salmo gairdneri) Early life stage   90 days > 3.68f ( NOEC) Spehar, 1986 Fathead minnow ( Pimephales promelas) Early life stage   32 days > 4.8f ( NOEC) Ahmad et al., 1984: Carlson and Kosian, 1987 Cladoceran ( Daphnia magna) 14 days 16 ( 50 percent reduction in fertility)   EC50 Calamari et al., 1983 a For several of the acute and subchronic toxicity data, the test duration was not reported. b For several of the acute and subchronic toxicity data, the distinction was not made between the LC50/ EC50 endpoint. c The exposure was continued for a period of ten days without any mortalities attributable to hexachlorobenzene. d The exposure was continued for a period of 8 days without any mortalities attributable to hexachlorobenzene. e The LC50 was above saturation, which was estimated to be 6

g/ L based on Metcalf et al., 1973. f Highest tested concentration; no tested concentration caused an unacceptable effect. 35 Appendix 3. Aquatic Toxicity Data from ECOTOX Database ( EPA, 2001)

Species Duration/ Type Endpoint/ Concentration Reference Acute/ Subchronic Toxicity Channel catfish ( Ictalurus punctatus) 24 hrs. exp type   static LC50   11.3, 16.2 mg/ L Mayer and Ellersieck, 1986 Channel catfish ( Ictalurus punctatus) 24 hrs. exp type   static LC50 > 100 mg/ L Mayer and Ellersieck, 1986 Channel catfish ( Ictalurus punctatus) 24 hrs. exp type   static LC50   7.0, 6.4, 7.7 mg/ L Mayer and Ellersieck, 1986 Channel catfish ( Ictalurus punctatus) 96 hrs. exp type  static LC50   14, 11,16 mg/ L Johnson W. and M. Finley, 1980 Brine shrimp ( Artemia salina) 24 hrs. exp type   static LC50 > 0.0116 mmol/ m3 Abernathy, S., Bobra, A., Shiu W., Wells P., and D MacKay, 1986 Toad ( Bufo bufo japonicus) 24 hrs. exp type   static LC50 > 4.2 mg/ L Nishiuchi, Y. 1980 American/ virginia oyster ( Crassostrea virginica) 48 hrs. exp type   not reported EC50   6.5 mg/ L ( morphology) Zaroogian, G., 1981 Diatom ( Cyclotella meneghiniana) 48 hrs. exp type   static EC50  2

g/ L ( genetic effect) Figueroa, I. And M. Simmons, 1991 Zebra danio ( Danio rerio) 48 hrs. exp type   static LC50 > 30

g/ L Calamari, D., Galassi, S., Setti, F., and Vighi, M., 1983 Water flea ( Daphnia magna) 48 hrs. exp type   static EC50 > 0.0166 mmol/ m3 ( intoxication) Abernathy, S., Bobra, A., Shiu W., Wells P., and D MacKay, 1986 Water flea ( Daphnia magna) 14 days Exp Type  Renewal EC50   16

g/ L ( reproduction) Calamari, D., Galassi, S., Setti, F., and Vighi, M., 1983 Water flea ( Daphnia magna) 21 days Exp Type  Renewal NOEC > 37.9 (

g/ L) ( behavioral effects) Caspers, N., Hartmann, P., Kanne, R., and Knoop, G., 1993. Water flea ( Daphnia magna) 21 days Exp Type  Renewal NOEC > 6.7 (

g/ L) ( reproduction effects) Caspers, N., Hartmann, P., Kanne, R., and Knoop, G., 1993. Green algae ( Haematococcus pluvialis) Duration and exposure type not reported EC10 > 40

g/ L ( population effect) Knie, J., Halke A., Juhnke, I., and W. Schiller, 1983 Channel catfish ( Ictalurus punctatus) 24 h exp type   static LC50   13.5, 11.3, 16.2 mg/ L Mayer and Ellersieck, 1986 Appendix 3. Aquatic Toxicity Data from ECOTOX Database ( EPA, 2001) ( Cont'd)

Species Duration/ Type Endpoint/ Concentration Reference 36 Channel catfish ( Ictalurus punctatus) 24 hrs. exp type   static LC50 > 100 mg/ L Mayer and Ellersieck, 1986 Channel catfish ( Ictalurus punctatus) 24 hrs. exp type   static LC50   7.0, 6.4, 7.7 mg/ L Mayer and Ellersieck, 1986 Channel catfish ( Ictalurus punctatus) 96 hrs. exp type   static LC50   14, 11,16 mg/ L Johnson W. and M. Finley, 1980 37 Appendix 4. Additional Aquatic Toxicity Data  Miscellaneous Sources Species Exposure Duration/ Type Endpoint/ Concentration Source Amphipod ( Hyalella azteca) 14 days ( spiked sediment) NOEC   42 mg/ kg Barber et al., 1997 Midge ( Chironomus tentans) 14 days ( spiked sediment) NOEC   42 mg/ kg Barber et al., 1997 Estuarine amphipod ( Leptocheirus plumulosus) 10 days ( spiked sediment) NOEC   60 mg/ kg Fuchsman et al., 1998 Freshwater amphipod ( Hyalella azteca) 10 days ( spiked sediment) NOEC   60 mg/ kg Fuchsman et al., 1998 Midge ( Chironomus tentans) 10 days ( spiked sediment) NOEC   60 mg/ kg Fuchsman et al., 1998 tubificid sludgeworms ( oligochaeta) 72 hrs. LC50/ EC50 > 1000 mg/ kg Meller et al., 1998 Flounder ( Platichthys flesus) 96 hr LC50   0.7 uM ( equivalent to 0.2 mg/ L) Furay and Smith, 1995 Sole ( Solea solea) 96 hr LC50   0.5 uM ( equivalent to 0.14 mg/ L) Furay and Smith, 1995 Water flea ( Daphnia magna) 48 hr ( flow through) LC50  > 5.0

g/ L Nebeker et al., 1989 ( no significant mortality observed at highest conc. tested   saturation snail ( Lymnaea palustris) 10 12 weeks 0.5

g/ L, 5

g/ L ( significantly reduced growth rate) Baturo et al., 1995 Crab ( Portunus pelagicus) 6 weeks 5

g/ L ( reduction in carapace growth rate) Mortimer and Connell, 1995 Freshwater amphipod ( Hyalella azteca) 30 days ( flow through) ( higher HCB water conc. of 2 separate trials) NOEC   4.5

g/ L Mean adult tissue value of 172

g/ g Nebeker et al., 1989 ( no significant mortality or effects to reproduction and growth observed at near water saturation Fathead minnow ( Pimephales promelas) 28 days ( flow through) NOEC  3.8

g/ L Mean tissue value of 47

g/ g Nebeker et al., 1989 ( no mortality or difference in growth rate observed).

Worm ( Lumbriculus variegatus) 49 days ( flow through) NOEC   4.7

g/ L Mean tissue value of 223

g/ g Nebeker et al., 1989 ( no significant effects on survival, growth or asexual reproduction were observed 38 Appendix 5. Terrestrial Toxicity Data ( Invertebrates, Birds, Mammals)   Ecotox Database, EPA 2001 Species Test Duration/ Typea Endpoint/ Concentration Reference Terrestrial Invertebrates Earthworm ( Eisenia fetida) 48 hrs ( dosed one time during the study) LC50 > 1000

g/ cm2 Neuhauser, 1985 Earthworm ( Eisenia fetida) 28 days LC50 > 1000 mg/ kg Ballhorn, 1984 Birds Mallard Duck ( Anas platyrhynchos) 8 days LC50 > 5000 ppm Hill et al., 1975 Ring necked pheasant ( Phasianus colchicus) 8 days ( exposed to HCB for a duration of 5 days) LC50   617 ppm Hill et al., 1975 Japanese quail ( Coturnix japonica) 5 days ( dose admin 1x per day for 5 days) LOEL   500 mg/ kg ( Biochemistry; microsomal protein response and enzymes; 7 ethoxyresorufin O deethylase and glutathione S transferase level) Carpenter, 1985a Japanese quail ( Coturnix japonica) 5 days ( dose admin 1x per day for 5 days) NOEL   500 mg/ kg ( Growth: whole body weight, and morphology: organ weight related to body weight Carpenter, 1985a Japanese quail ( Coturnix japonica) 5 days ( dose admin 1x per day for 5 days) NOEL   500 mg/ kg ( Enzymes, epoxide hydrase, NADPH cytochrome C reductase, and cytochrome P 450 levels) Carpenter, 1985a Japanese quail ( Coturnix japonica) 15 days ( dose admin 1x per day for test duration) LOEL   100 mg/ kg ( biochemistry, protoporphyrin level) Carpenter, 1985b Japanese quail ( Coturnix japonica) 17 days ( dose admin 1x per day for test duration) LOEL   100 mg/ kg ( Enzymes, ALA synthetase level) Carpenter, 1985b Japanese quail ( Coturnix japonica) 17 days ( dose admin 1x per day for test duration) LOEL   100 mg/ kg ( Morphology, organ weight related to body weight) Carpenter, 1985b Japanese quail ( Coturnix japonica) 90 days LOEL   20 ppm ( Egg volume) Vos, 1971 Japanese quail ( Coturnix japonica) 90 days NOEL   5 ppm ( Egg volume) Vos, 1971 Japanese quail ( Coturnix japonica) 90 days LOEL   5 ppm ( Morphology, organ weight related to body weight) Vos, 1971 Japanese quail ( Coturnix japonica) 90 days NOEL   1 ppm ( Morphology, organ weight related to body weight) Vos, 1971 Appendix 5. Terrestrial Toxicity Data ( Invertebrates, Birds, Mammals)   Ecotox Database, EPA 2001 ( Cont'd)

Species Test Duration/ Typea Endpoint/ Concentration Reference 39 Japanese quail ( Coturnix japonica) 90 days LOEL   5 ppm ( Morphology, organ weight) Vos, 1971 Japanese quail ( Coturnix japonica) 90 days NOEL   1 ppm ( Morphology, organ weight) Vos, 1971 Japanese quail ( Coturnix japonica) 4 weeks LOEL   120 ppm ( Biochemistry, hemoglobin level and packed cell volume) Vos, 1971 Japanese quail ( Coturnix japonica) 4 weeks NOEL   20 ppm ( Biochemistry, hemoglobin level and packed cell volume) Vos, 1971 Japanese quail ( Coturnix japonica) 4 weeks LOEL   20 ppm ( Enzymes, glutamicoxaloacetic transaminase) Vos, 1971 Mammals Deer mouse ( Peromyscus maniculatus) 3 days ( dosed one time per study period) LD50   710 mg/ kg Schafer and Bowles, 1985 a All tests are oral unless otherwise indicated. 40 Appendix 6. Terrestrial Toxicity Data  EPA 1985, Health Assessment for Chlorobenzenes ( Mammals Only)

Species Test Duration/ Type Endpoint/ Concentration Reference Intermediate/ Subchronic Exposure Rat 43 days ( dosed every other day) 100 mg/ kg   effect: suggested covalent bonding of HCB metabolites to cytosolic proteins. Koss et al., 1980a Rat 98 days 100 mg/ kg diet   effect: porphyria ( increased liver lobe porphyrins), decreased activity of uroporphyrinogen decarboxylase Smith et al., 1980 Rat 103 days ( dosed every other day) 14 mg/ kg   effect: porphyria in treated females, poss. related to estrogen levels Rizzardini and Smith, 1982 Rat ( females) 6 weeks exposed ( dosed every other day) then held for additional 18 months 100 mg/ kg   effect: porphyria ( liver uroporphyrin levels peaked 7 months postexposure hadn't returned to normal by 18 months) Koss et al., 1983 Dog ( female) 21 days 50 or 150 mg/ kg/ day   liver and hepatocyte enlargement, electroencephalogram dysrhythmias Sundlof et al., 1981 Monkey ( nursing) 60 days 7.5   186 ppm milk   2 of 3 infants died as a result of exposure Bailey et al., 1980 Mink From gestation until 17 weeks of age 1 or 5 mg/ kg diet   dose related increase in offspring mortality, induction of hepatic enzymes in exposed offspring. Rush et al., 1983 Hamster 90 days 200 mg/ kg diet   hepatic lesions, bile duct hyperplasias and hepatomas Lambrecht et al., 1982 Appendix 6. Terrestrial Toxicity Data  EPA 1985, Health Assessment for Chlorobenzenes ( Mammals Only) ( Cont'd)

Species Test Duration/ Type Endpoint/ Concentration Reference 41 Chronic Exposure Rat ( females) 75 90 weeks 6 8 mg/ kg/ day   effect: decline in body weight, porphyria, enlarged liver and liver tumors Smith and Cabral, 1980 Rat Up to 2 years 4 5 mg/ kg/ day   effect: porphyria, increased heptomas, heptocarcinomas, bile duct adenomas, renal adenomas and renal carcinomas Lambrecht et al., 1983, a, b Rat Gestation through lifetime ( 130 weeks) 0.32 mg/ kg diet  NOEC Arnold et al., 1985 Rat Gestation through lifetime ( 130 weeks)   chronic 1.6 mg/ kg diet   NOAEL for liver effects ( 0.08 mg/ kg/ day) Arnold et al., 1985 Rat Gestation through lifetime ( 130 weeks)   chronic 8.0 mg/ kg diet   LOAEL  increase in liver pathologies ( 0.29 mg/ kg/ day) Arnold et al., 1985 Rat Gestation through lifetime ( 130 weeks)   chronic 40 mg/ kg diet   increased mortality as pups, increase in liver and kidney pathologies, and parathyroid tumors in males Arnold et al., 1985 Rat Fo to F1a and F1b generations 21 day LD 50 values for pups were 100 and 140 mg/ kg for the F1a and F1b generations, respectively Kitchin et al., 1982 42 Appendix 7. Terrestrial Toxicity Data ( Mammals)  ATSDR, 2000   Toxicity Profile for Hexachlorobenzene Species Test Duration/ Type Endpoint/ Concentration Reference Acute Exposure Rat NRa LD50   3,500 mg/ kg Savitskii ( 1964, 1965) Mouse NR LD50   4,000 mg/ kg Savitskii ( 1964, 1965) Rabbit NR LD50   2,600 mg/ kg Savitskii ( 1964, 1965) Cat NR LD50   1,700 mg/ kg Savitskii ( 1964, 1965) Intermediate/ Subchronic Exposure Rat 80 days 100 mg/ kg/ day   lethality ( 60 of 90 individuals died) Cuomo et al., 1991 Rat 13 weeks 19 mg/ kg/ day   lethality ( 4 of 9 individuals died) Den Besten et al. ( 1993)

Rat 80 days LOAEL   100 mg/ kg/ day ( increased hepatic porphyrins) Cantoni et al. ( 1990)

Rat 8 weeks ( 1x per day) LOAEL   1000 mg/ kg/ day ( increased liver weight and and hepatic porphyrins) Kleiman de Pisarev et al. ( 1990)

Rat 4 weeks ( 1x per day) LOAEL   1000 mg/ kg/ day ( increased liver weight) Kleiman de Pisarev et al. ( 1995)

Rat 30 days ( dosed one time per day) LOAEL   500 mg/ kg/ day ( endocrine system, decreased serum T4 levels) Kleiman de Pisarev et al. 1989 Rat 6 weeks ( 1 time/ day, 5 days per week) LOAEL   50 mg/ kg ( increased hepatic and urinary porphyrins Krishnan et al. ( 1992)

Rat 56 days LOAEL   5 mg/ kg/ day ( increased highly carboxylated porphyrins in the liver) Kennedy and Wigtield ( 1990)

Rat 5 weeks, ( dosed 1x per day, 5 days per week) LOAEL   1 mg/ kg/ day ( increased liver weight) Andrews et al ( 1988)

Mouse 7 weeks LOAEL   26 mg/ kg/ day ( increased hepatic porphyrins) Vincent et al. ( 1989)

Chronic Exposure Rat 2 generations NOAEL   2 mg/ kg/ day ( hematological, renal and body weight effects): Arnold et al. ( 1985) Appendix 7. Terrestrial Toxicity Data ( Mammals)  ATSDR, 2000   Toxicity Profile for Hexachlorobenzene ( Cont'd)

Species Test Duration/ Type Endpoint/ Concentration Reference 43 Rat 2 generations NOAEL   0.08 mg/ kg/ day ( hepatic effects) Arnold et al. ( 1985)

Rat 2 generations LOAEL   0.4 mg/ kg/ day ( increased liver weight) Arnold et al. ( 1985)

Rat 2 generations NOAEL   2 mg/ kg/ day ( reproductive effects) Arnold et al. ( 1985)

Rat 2 generations LOAEL   0.016 mg/ kg/ day ( developmental effects, lymphocytosis and fibrosis of the liver) Arnold et al. ( 1985)

Rat 90 weeks LOAEL   10 mg/ kg/ day ( hepatocyte hypertrophy, bile duct hyperplasia) Smith et al. ( 1985)

Mouse 18 months LOAEL   13 mg/ kg/ day ( hepatocyte hypertrophy) Smith et al.( 1989)

a NR = Not Reported 44 Appendix 8. Terrestrial Toxicity Data   Miscellaneous Sources Species Test Duration/ Type Endpoint/ Concentration Reference Birds Japanese quail 5 day LC50   568 ppm Hill, E. and M. Camardese, ( 1986) Bobwhite quail NRa LD50  575 mg/ kg U. S. National Library of Medicine. ( 1995) Mallard duck NR LD50  1450 mg/ kg U. S. National Library of Medicine. ( 1995) Coturnix quail Acute ( 5 days) NOAEL   22,500

g/ kg BW day Hill and Camardese ( 1986) Japanese quail Acute ( 5 days)   dosed one time each day for study duration LOEL   500 mg/ kg per day ( hepatic toxicity, elevation in level of hepatic porphyrins) Carpenter et al. ( 1985) Carpenter, Williams and Buhler   1985 Herring gulls ( Larus argentatus) Egg injection study LD50   4300

g/ kg ( embryo mortality) Boersma et al., 1986 Herring gulls ( Larus argentatus) Egg injection study 1500

g/ kg   significant reduction in embryonic weight Boersma et al., 1986 Double breasted Cormorant ( Phalacrocorax auritus) NR 95 percent reduction in egg hatchability from a contaminated colony   0.01 mean ppm in 18 eggs, wet weight Weseloh et al., 1983 Mammals Rat Chronic/ sub chronic   30 days LOAEL   1 g/ kg BWday ( 1000 mg/ kg BWday   hepatoxicity, porphyria, possible alteration of oestrus cycles. Alvarez, 2000 Mink Chronic ( 47 weeks) LOAEL   1 ppm ( 0.16 mg/ kg BW day)   reduced birthweight and increased mortality Bleavins et al., 1984 Rat Chronic (> 247 days) NOAEL   1,600

g/ kg BW day ( reproductive effects) Grant et al., 1977 a NR = Not Reported 45 APPENDIX 9: Terrestrial Exposure Assessment The meadow vole ( Microtus pennsylvanicus) was selected as the mammalian receptor species for evaluating potential effects of HCB to mammals. The meadow vole is primarily herbivorous and is widely distributed in the United States ( U. S. EPA, 1993b). The northern bobwhite quail ( Colinus virginianus) was selected as the avian receptor species for evaluating potential effects of the components of HCB to avian species. The northern bobwhite quail feeds mainly on seeds and lowlying vegetation. The bobwhite range includes the eastern and central U. S. as well as portions of the Rocky Mountains and the southwest. The mouse and bobwhite were selected because a relatively large proportion of the diet of both species is comprised of vegetation and there is an extensive amount of toxicity data available for these species, particularly for the bobwhite.

The following discussion presents the methods used to calculate the potential ingestion of chemicals by the mouse and bobwhite via the ingestion of food ( i. e., terrestrial plants) and surface soil. The equations presented below were derived based on equations presented by U. S. EPA ( 1989). The following equation was used to calculate the dose of chemicals that a mouse or bobwhite would be expected to obtain from the ingestion of terrestrial plants:

( 1) Dose FI * C plant diet =

where:

Dose plant = amount of chemical ingested per day via ingestion of plants ( mg/ kg bw d); FI = food ingestion rate ( kg/ kg bw d); and C diet = estimated chemical concentration in diet ( mg/ kg).

Food ingestion rates ( FI) of 0.35 kg/ kg bw d for adult voles ( U. S. EPA, 1993) and 0.093 kg/ kg bw d for adult bobwhites ( U. S. EPA, 1993) were used in the assessment.

The estimated dietary concentration ( C diet) was calculated using the following equation:

( 2) C P* C diet p p =

where:

P p = proportion of diet consisting of vegetation ( unitless); and C p = estimated concentration of contaminant of concern in vegetation ( mg/ kg).

The proportion of the diet ( P p) consisting of vegetation for both the vole and bobwhite was assumed to be 100%. This value is fairly consistent with the data compiled in U. S. EPA ( 1993) for these species. For both voles and bobwhites it was also assumed that 100% of the plants ingested are from the areas with elevated concentrations of HCB. This assumption is conservative and may lead to an overestimate of potential risks because the species are likely to also forage in areas that may not have elevated soil concentrations. 46 The concentration of a chemical in a terrestrial plant ( C p) as fresh weight was determined using the following equation:

( 3) C C * BCF p soil =

where:

C soil = concentration ofcontaminant of concern detected in surface soil ( mg/ kg); and BCF = bioconcentration factor for the chemical in terrestrial plants ( unitless).

The predicted EEC for HCB in soil ( U. S. EPA, 2002) was used as the C soil in the model for voles and bobwhites.

Bioconcentration Factors ( BCFs) compiled by U. S. EPA ( 1999) were used to calculate the chemical concentrations in plants for the assessment. Employing equations 1, 2, and 3, the estimated dose that voles and bobwhites would receive from ingestion of plants in the vicinity of pentachlorophenoltreated utility poles is provided in Tables A1 and A2.

In addition to the ingestion of chemicals accumulated in vegetation, bobwhites and voles also may be exposed to chemicals through the inadvertent ingestion of surface soil while foraging or grooming. The following equation was used to calculate the dose of chemical that voles and bobwhites would be expected to obtain from the ingestion of surface soil:

( 4) Dose = SI * C soil soil where:

Dose soil = amount of chemical ingested per day from soil ( mg/ kg bw d); SI = soil ingestion rate ( kg/ kg bw d); and C soil = chemical concentration in surface soil ( mg/ kg).

Based on percent dietary soil ingestion values presented by Beyer et al. ( 1994), it was assumed that 10% of the total mass of both the vole and the bobwhites' diet consists of soil. The percent soil ingestion was multiplied by the food ingestion rates ( FI) presented earlier for these species to estimate soil ingestion rates ( 0.035 kg/ kg bw d for voles and 0.0093 kg/ kg bw d for bobwhites). Employing equation 4, the estimated dose voles and bobwhites would receive from the ingestion of soil for HCB are discussed in the report.

The total dietary exposure levels for voles and bobwhites was determined using the following equation:

( 5) Dose Dose Dose total plant soil = + 47 Using equation 5, the estimated total dose voles and bobwhites would be expected to receive from the ingestion of plants and soil is shown in Tables A1 and A2.

In the Risk Characterization section, the total dietary intakes are compared to toxicity reference values ( e. g., LD50, NOAEL) to determine if adverse effects are likely to occur in voles and bobwhites from the ingestion of HCB in terrestrial plants and surface soil. The toxicity reference values for avian species identified for HCB were based on laboratory toxicity data on several avain species including the bobwhite quail and the Japanese quail. Toxicity reference values from studies based on avian species other that the bobwhite quail were not adjusted for the bobwhite's body weight based on recommendations in Sample et al. ( 1996). Toxicity reference values for mammals were adjusted to the vole's body weight using the following equation provided in Sample et al. ( 1996):

( 6) NOAEL NOAEL * ( bw / bw ) w t tw 1/ 4 =

Where:

NOAEL w = NOAEL for mammalian wildlife species ( e. g., meadow vole) NOAEL t = NOAEL for laboratory mammalian species ( e. g., rat, rabbit, etc.) bw w = body weight for mammalian wildlife species ( e. g., meadow vole) bw t = body weight for laboratory mammalian species ( e. g., rat, rabbit, etc.)

The body weight for the meadow vole was assumed to be 0.044kg and body weights for the rat, rabbit, mouse, and mink were assumed to be 0.35kg, 3.8kg, 0.035kg, and 1.0kg, respectively, based on reference values provided in Sample et al. ( 1996) and U. S. EPA's Wildlife Exposure Factors Handbook ( U. S. EPA, 1993).

Following adjustment of the mammalian toxicity reference values, the total dose for the bobwhite and vole of HCB were compared with the toxicity reference values for these contaminants to determine the risk quotients ( RQs). The RQs for the bobwhite and vole based on exposure to HCB are shown in Tables A1 and A2, respectively. 48 Table A1. Acute and Chronic Risk Quotient Calculations for Birds EEC ( mg/ kg) Bobwhite Soil Ingestion Rate ( kg/ kg d) Bobwhite Soil Ingestion ( mg/ kg day) BCF ( a) Vegetation Residue mg/ kg) Proportion of Vegetation in Bobwhite's Diet ( unitless) Estimated Chemical Concentration in the Diet ( mg/ kg) Food Ingestion Rate ( kg/ kg bw d) Dose from Plant Ingestion ( mg/ kg day) Total Dose ( Plant + Soil)

( mg/ kg day) Endpoint RQ HCB 2.58e 08 0.0093 2.40e 10 2.20e 03 5.68e 11 1 5.68e 11 0.093 5.28e 12 2.45e 10 LD50 568 ( mg/ Kg) acute 4.32E 13 NOAEL 1 ( mg/ kg) chronic 2.45E 10 ( a) Values for Bioconcentration Factors were provided in U. S. EPA 1999.

Table A2. Acute and Chronic Risk Quotient Calculations for Mammals EEC ( mg/ kg) Vole Soil Ingestion Rate ( kg/ kg d) Vole Soil Ingestion ( mg/ kg day) BCF ( a) Vegetation Concentrat ion ( mg/ kg) Proportion of Vegetation in Vole's Diet ( unitless) Estimated Chemical Concentration in the Diet ( mg/ kg) Food Ingestion Rate ( kg/ kg bw d) Dose from Plant Ingestion ( mg/ kg day) Total Dose ( Plant + Soil)

( mg/ kg day) LD50 ( mg/ kg) RQ HCB 2.58e 08 0.035 9.03e 10 4.91e 03 1.27e 10 1 1.27e 10 0.35 4.43e 11 9.47e 10 LD50 708 mg/ kg acute 1.34E 12 NOAEL 0.95 mg. kg day Chronic 9.94E 10 ( a) Values for Bioconcentration Factors were provided in U. S. EPA 1999.