Page 1 of 65 ECOLOGICAL RISK ASSESSMENT Dioxins And Furans: Hazard and Risk to Wildlife and Aquatic Organisms 3/ 4/ 05 Page 2 of 65 I. Executive Summary Chlorinated dibenzo dioxins ( CDDs) and chlorinated dibenzo furans ( CDFs) occur in pentachlorophenol as contaminants. These compounds are inherently toxic, as well as environmentally persistent, and their presence may increase the ecological risk associated with the use of pentachlorophenol. There are many congeners of CDDs/ CDFs, ranging from monochlorinated to octachlorinated. The most toxic for each compound seems to be the 2, 3, 7, 8 tetrachlorinated congener, referred to as TCDD or TCDF for dioxin or furan, respectively.

Data on the ecological effects of CDDs and CDFs are relatively limited. There are no Guideline studies required for either compound, since they are not registered as pesticides. Most research efforts have been focused primarily on 2,3,7,8 chlorinated CDD/ CDFs, especially 2,3,7,8 TCDD. The majority of the data presented in this chapter are based on compilations of data in U. S. EPA ( 1990), U. S. EPA ( 1993), and the U. S. EPA ACQUIRE database.

A. Toxicity to Aquatic Organisms:

In aquatic toxicity studies, the most toxic congener of polychlorinated dibenzo p dioxins and dibenzofurans is 2,3,7,8 TCDD, probably due to its tendency to be taken up more readily than other congeners. Fish show greater sensitivity to dioxins than do aquatic invertebrates, but dioxins are very highly toxic to all aquatic organisms. The available 96 hour LC 50 values for fish range from 1.83 ng/ L ( ppt) in rainbow trout ( Bol et al, 1989) to 5.60 ng/ L in coho salmon ( Miller et al, 1973). The only available 48 hour LC 50 value for aquatic invertebrates was > 1,030 ng/ L in daphnids ( Adams et al, 1986). Longer term LC 50 values indicate greater toxicity [ 0.0011 ng/ L in rainbow trout ( Mehrle et al , 1988) to 1.7 ng/ L in fathead minnows ( Adams et al, 1986) for 28 day exposures], due to TCDD's tendency to cause delayed mortality. TCDD induced mortality typically occurs between 30 to 80 days after the initial exposure, even from short term exposure durations. Early life stage testing in fish resulted in a NOEC value of 0.0011 ng/ L ( Mehrle et al, 1988). Based on the available data, the most sensitive fish life stage is during yolk sac resorption in the developing embryo.

Data on the toxicity of furans to aquatic organisms indicate that they are generally less toxic than dioxins, but 2,3,7,8 TCDF is still very highly toxic to fish [ 13 day LC50 value of 16 ng/ L in medaka ( Wisk & Cooper, 1990a)].

Many sublethal effects have also been reported for polychlorinated dioxins and furans, including: reduced growth; weight loss; abnormal hatching; cellular alterations; numerous types of lesions in a broad spectra of organs; and behavioral effects on swimming, reduced feeding, and loss of response to external stimuli.

B. Toxicity to Terrestrial Organisms: Page 3 of 65 Avian toxicity data are more limited for the CDDs/ CDFs. The available LD 50 values for 2,3,7,8 TCDD range from 15

g/ kg in Northern bobwhite to 810

g/ kg in Ringed turtle dove ( Hudson et al, 1984), all of which indicate very high acute toxicity. An LC 50 of 167 ppt for Northern bobwhite in an 8 day dietary exposure study also demonstrates the very high toxicity of TCDD to birds. Studies examining effects of injecting TCDD into eggs have shown that very low levels ( NOEC of 0.1

g/ kg) cause mortality in chicks up to 28 days post hatch ( Nosek et al, 1992c).

An LC 50 of 0.264

g/ kg bw/ day in female mink ( Hochstein et al, 1988) indicates that TCDD has very high acute toxicity to mammals. Studies of the reproductive effects of TCDD in monkeys have shown it to cause spontaneous abortions and other toxic effects at very low levels ( Kociba and Schwetz, 1982).

CDDs and CDFs are very highly toxic to birds, mammals and aquatic organisms. CDDs ( and possibly furans) are capable of producing lasting toxic effects; even a relatively short exposure to TCDD ( as little as 6 hours) can result in mortality occurring as much as 80 days later. TCDD is a known endocrine disruptor, and it is likely that other dioxin congeners and furans produce similar effects. While there are no EPA Guideline studies available for these compounds, the literature data are sufficient to conclude that these compounds have the potential to adversely affect wildlife and aquatic organisms.

C. Ecological Risk There is a great deal of uncertainty in these quantitative estimates, however, due to a lack of standard toxicity endpoints and exposure information. There are also uncertainties and limitations inherent to the exposure models used, as described in the Environmental Modeling chapter of this document. Based on the information that is available, it is unlikely that the CDDs/ CDFs resulting from the use of pentachorophenol on utility poles are posing an immediate risk to wildlife or aquatic organisms; however, it is possible that the buildup of these persistent and bioaccumulative compounds over time may eventually reach levels that pose a risk to these organisms. Additionally, pentachlorophenol is only one of many sources of CDDs/ CDFs in the environment; it is not currently possible to quantify the portion of the aggregate environmental risk from CDDs/ CDFs that is attributable to the use of pentachlorophenol on utility poles.

II. Background/ Introduction CDD/ CDFs are dispersed throughout the global ecosystem. They are chemically stable and can bioaccumulate in animal tissue. Data on the ecological effects of CDD/ CDFs are limited. Most research efforts have been focused primarily on 2,3,7,8 chlorinated CDD/ CDFs, especially 2,3,7,8 TCDD. This chapter presents a compilation and evaluation of the available ecological effects data for CDD/ CDFs on aquatic life ( e. g., fish, arthropods, etc.), wild mammals ( e. g., mink, monkeys), and birds.

The majority of the data presented in this chapter are based on compilations of data in U. S. EPA ( 1990), U. S. EPA ( 1993), and the U. S. EPA ACQUIRE database. In some instances, data from the same study are presented in multiple tables. A brief discussion of some selected studies is also Page 4 of 65 presented. The reader is referred to U. S. EPA ( 1990) and U. S. EPA ( 1993) for a more detailed presentation of the studies used therein.

III. Toxicity of CDDs/ CDFs to Aquatic Organisms While there are no EPA Guideline studies available for CDDs/ CDFs, literature data provide some studies with comparable endpoints. These values are presented in Table 1, below.

Table 1: Selected Aquatic Toxicity Values for CDDs/ CDFs Compound Species Test Type/ Duration Endpoint Toxicity classification Citation 2,3,7,8 TCDD Coho salmon 96 hour acute LC50= 5.60 ppt very highly toxic Miller et al, 1973 Rainbow trout sac fry 96 hour acute LC50= 1.83 ppt very highly toxic Bol et al, 1989 Rainbow trout swimup fry 28 day early life stage LC50= 0.046 ppt NOEC= 0.0011 ppt very highly toxic Mehrle et al, 1988 fathead minnows 28 day LC50= 1.7 ppt very highly toxic Adams et al, 1986 Daphnia magna 48 hour acute EC50> 1020 ppt very highly toxic Adams et al, 1986 Annelid worm 55 days NOEC = 200 ppt Miller et al, 1973 Snail 12 days NOEC = 200 ppt Miller et al, 1973 Dibenzofuran Sheepshead minnow 96 hour acute LC50= 1800 ppb moderately toxic Heitmuller et al, 1981 fathead minow 96 hour acute LC50= 1140 ppb moderately toxic Brooke, 1991 Daphnia magna 48 hour acute LC50= 1340 ppb moderately toxic Maas, 1990 Mysid 96 hour acute LC50= 1310 ppb moderately toxic EPA, 1978 Diatom, Skeletonema costatum 96 hour acute EC50= 1500 ppb EPA, 1978 2,3,7,8 TDCF Medaka 13 day LC50= 16 ppb very highly toxic Wisk & Cooper, 1990a Based on the available data, the most toxic CDD congener is 2,3,7,8 TCDD and the most toxic CDF congener is 2,3,7,8 TCDF. Many sublethal effects have also been reported for CDDs and Page 5 of 65 CDFs, including: reduced growth; weight loss; abnormal hatching; cellular alternations; numerous types of lesions in a broad spectra of organs; and behavioral effects on swimming, reduced feeding, and loss of response to external stimuli.

A. Aquatic Toxicity of CDDs The available studies on CDDs and CDFs published between 1984 and 1989 were collected and reviewed in a document entitled " Background Document to the Integrated Risk Assessment for Dioxins and Furans from Chlorine Bleaching in Pulp and Paper Mills" ( U. S. EPA, 1990). This 1990 EPA report evaluated the aquatic toxicity literature published subsequent to the comprehensive EPA reviews in the ambient water quality criteria document ( U. S. EPA, 1984a) and the health and environmental effects profile ( U. S. EPA, 1984b). Thus, the information contained in U. S. EPA ( 1990) does not include a detailed review of earlier aquatic toxicity data, but does contain a minimal amount of data from studies prior to this timeframe. Test concentrations and effects data reported in U. S. EPA ( 1990) are summarized for CDDs in Appendix 1.

The document, " Interim Report on Data and Methods for Assessment of 2,3,7,8 Tetrachlorodibenzo p dioxins Risks to Aquatic Life and Associated Wildlife" ( U. S. EPA, 1993) provides a compilation of data on the toxic effects of 2,3,7,8 TCDD to aquatic life. A summary of this compilation is provided in Appendix 2. This report was prepared as part of the Agency's effort to conduct a scientific reassessment of the risk of TCDD.

U. S. EPA ( 1993) conducted a literature search to obtain aquatic toxicity and accumulation data for both fresh and saltwater organisms. Laboratory tests and aquatic model ecosystem studies were included. Most of the studies involved short term exposure periods where the organisms were exposed to TCDD. Appendix 2 contains the available data for freshwater plants, freshwater fish, freshwater invertebrates, amphibians, and saltwater fish.

EPA maintains a database ( ACQUIRE) where ecological data from laboratory studies on various chemicals are stored. Summary data for aquatic toxicity from the EPA ACQUIRE database are presented in Appendix 3. The data presented are only those data where a toxic endpoint was provided.

The most toxic CDD to aquatic life is 2,3,7,8 TCDD. CDDs appear to be more toxic to fish than to invertebrates; however, laboratory data indicate that TCDD is very highly toxic to all aquatic organisms.

The mode of action of 2,3,7,8 TCDD is unknown, but it appears to be related to neural toxicity and may also affect the immune system, like many other organochlorine chemicals. The pattern of TCDD induced mortality typically occurs between 30 to 80 days after the initial exposure, even for a duration of exposure as short as 6 hours. Mortality in fish appears to be a function of both exposure duration and the test concentration. The available bioconcentration tests were not of sufficient duration to achieve a steady state bioconcentration factor ( BCF). All BCF values have been estimated using the uptake rate ( k 1) and depuration rate ( k 2) values. Page 6 of 65 Aquatic studies with 2,3,7,8 TCDD have generally been conducted either as continuous long term exposures or short term exposures ( 6 to 96 hours) with long term observation periods. The long observation periods are necessary to monitor the delayed mortality ( e. g., 30 to 80 days postexposure which is characteristic of 2,3,7,8 TCDD. The occurrence of delayed effects has some serious implications for both hazard evaluations and risk assessments. For example, the reported results of typical acute toxicity tests are probably not an accurate indication of toxicity because of the short observation periods. In risk scenarios, sporadic acute releases may cause serious delayed effects. Of most concern is that there are insufficient data to assess the effects from the many complex combinations of exposure concentrations and the various durations of the exposure that occur in nature.

The delayed mortality demonstrated in dioxin studies indicates that the most appropriate method of reporting the effects of dioxin is not a simple 96 hour LC 50 value or a maximum acceptable toxicant concentration ( MATC) range. The toxicity of 2,3,7,8 TCDD is a function of the duration of the exposure, as well as the dose level. At any given test concentration, continuous exposures appear to produce toxic effects in a shorter time than do short term exposures. Dioxin effects from various routes of exposure, including water, sediment, diet, and intraperitoneal injection are summarized in Table 1. Based on these data, exposure via water appears to be the most toxic, single route of exposure. However, it is likely that the degree of an effect could be increased by a combination of exposure routes, such as would be found in a natural ecosystem. This is especially likely since the low water solubility and high octanol/ water partition coefficient of 2,3,7,8 TCDD suggest most of that which would be found in the aquatic ecosystem would be sorbed to or bioaccumulated in dietary sources.

The duration of an exposure is important when evaluating the inherent chronic toxicity of 2,3,7,8 TCDD; however, toxic effects were reported even for the shortest 2,3,7,8 TCDD exposure period in water ( 6 hours), following which rainbow trout mortalities occurred between 78 and 139 days after the exposure ( Branson et al., 1985). Adams et al. ( 1986) estimated that 90 percent of steady state uptake of 2,3,7,8 TCDD in rainbow trout would take longer than 48 days. Consequently, an adequate chronic study would require than an exposure period be extended until the maximum level of 2,3,7,8 TCDD has been reached, plus sufficient additional time for effects to be produced. Review of the studies in Table 1 indicates that the studies with water exposures are not of sufficient duration for a steady state concentration of 2,3,7,8 TCDD to have been achieved in fish tissues.

The longest exposure to 2,3,7,8 TCDD via water was reported to be 28 days, with an additional 28 days post treatment for observation for rainbow trout fry ( Mehrle et al., 1988). In that study, a NOEC was not determined. At the lowest test concentration ( 0.038 ng/ l), 45 percent mortality occurred before the test ended.

The 28 day trout study ( Mehrle et al., 1988) did not investigate deposition of 2,3,7,8 TCDD in the eggs by the female. Some evidence is available which suggests that for trout the most sensitive life stage for chlorinated chemicals occurs when the toxicant stored in the yolk is sorbed into the developing embryo at the end of the yolk sac stage. Similarly, Helder ( 1980) reported that following a 96 hour exposure to 2,3,7,8 TCDD, the highest mortality in freshly fertilized pike eggs occurred during resorption of the yolk. Mortality reached almost 100 percent at 2,3,7,8 Page 7 of 65 TCDD exposure concentration of 10 ng/ l. The NOEC was reported as less than the lowest test concentration ( 0.1 ng/ l).

Sublethal effects of CDDs have been studied by several researchers, but it appears that the studies were not conducted for sufficient duration to determine the inherent toxicity of CDDs. The effects of 2,3,7,8 TCDD on the development of medaka embryos have been studied by Wisk and Cooper ( 1986). It was found that 20 to 40

g/ l caused tube hearts, hemostasis, and liver necrosis. In another study ( Helder and Seinen, 1986), 2,3,7,8 TCDD and 2,3,7,8 TCDF produced lethal edematous conditions in rainbow trout. Similar effects of 2,3,7,8 TCDD were reported by Helder ( 1980) during early life stages of the pike, in which the sublethal effects included: reduced size at hatch, tail first hatching, alteration of blood vessel walls, pericardial edema, enlarged nuclei and degradation in hepatocytes, and severe generalized edemas preceding death.

Sublethal effects on rainbow trout have been studied by Spitsberg et al. ( 1988a) following single intraperitoneal injections of either 1, 5, 25, or 125

g 2,3,7,8 TCDD/ kg. At both 25 and 125

g/ kg, mortality occurred before body weight loss could be detected. At 5

g/ kg, effects included reductions in activity, feeding, and growth. Moderate to marked leukopenia and thrombocytopenia were also found. The 2,3,7,8 TCDD treated trout also suffered higher stress from handling than controls. Numerous types of lesions were reported in lymphomyeliod tissues of the thymus, spleen, and kidneys, as well as epithelial lesions of the stomach mucosa, liver, pancreas, gill, and skin. The types of lesions included the following: thymic involution; splenic lymphoid depletion; decreased kidney hematopoiesis; multifocal necrosis, atrophy, and hyperplasia of the stomach mucosa; mild to severe hepatocyte vacuolation and ballooning degeneration of the liver and hyperplasia of the bile duct; vacuolation of pancreatic exocrine cells; mild fusion of gill lamellae; and necrosis of fin margins. Histologic changes in epithelial organs resembled the lesions found in early stage studies of rainbow trout exposed to water borne 2,3,7,8 TCDD.

Sublethal effects in yellow perch following single intraperitoneal injections of either 1, 5, 25, or 125

g 2,3,7,8 TCDD/ kg have also been reported ( Spitsberg et al., 1988b). Many effects were, in general, similar to those reported in rainbow trout. Yellow perch did not, however, show lethality resulting from handling as seen in rainbow trout. Lesions in the thymus, spleen, kidney, stomach, and skin are similar to the effects observed in rainbow trout. Cardiac lesions not seen in rainbow trout were found, including: necrosis of myocytes subjacent to the epicardial surface of the ventricle; fibrinous pericarditis; and hypertrophy and hyperplasis of the pericardial mesothelium. Many of the histologic lesions, including fibrinous pericarditis, have also been found in mammals, chickens, or turkeys treated with 2,3,7,8 TCDD.

Spitsberg et al. ( 1986) measured immune responses in rainbow trout at 14 and 30 days after single intraperitoneal injections of either 0.1, 1.0, or 10

g/ kg of 2,3,7,8 TCDD. Trout injected with 2,3,7,8 TCDD at either 0.1 or 1.0

g/ kg remained clinically normal. Trout treated with 10

g/ kg of 2,3,7,8 TCDD became hypophagic and exhibited fin necrosis, ascites, and suppression of hematopoiesis. Concanavalin A induced blastogenesis of thymic and splenic lymphocytes were not significantly changed, however, suppression of the pokeweed mitogen induced response of splenic lymphocytes occurred. No statistically significant alterations occurred in humoral immune responses, and phagocytic activity of peritoneal macrophages was not decreased. In rainbow Page 8 of 65 trout, immunosupression was evident only at 2,3,7,8 TCDD doses approaching 20

g/ kg, which is the 80 day, single dose, parenteral LD 50 value.

Spitsberg et al. ( 1988c) found no statistically significant effects on rainbow trout mortality or mean time until death following a challenge with infectious haematopoietic necrosis virus ( IHNV). Rainbow trout were injected intraperitoneally with single doses of either 0.01, 0.1, 1.0, or 10

g/ kg 2,3,7,8 TCDD. In virus free trout, TCDD induced effects first appeared 4 to 5 weeks after injection of 10

g/ kg. The effects included fin necrosis, as well as reduction in activity and food consumption. No deaths occurred in the virus free rainbow trout during the 5 weeks after treatment.

Behavioral effects induced by 2,3,7,8 TCDD in rainbow trout have also been reported by Mehrle et al. ( 1988). Rainbow trout were severely stressed at all test concentrations ( i. e., 0.038 to 0.789 ng/ l of 2,3,7,8 TCDD). Behavioral impairments increased over time and with increasing test concentration. The behavioral changes included: lethargic swimming, abnormal head up swimming posture, feeding inhibition, and lack of response to external stimuli. All of these behaviors would increase the likelihood of predation in a natural ecosystem.

B. Aquatic Toxicity of CDFs Data compiled in U. S. EPA ( 1990) on the toxic effects of CDFs are summarized in Appendix 4, and a summary of aquatic toxicity data for CDFs from the ACQUIRE database is presented in Appendix 5.. The longest exposure of 2,3,7,8 TCDF via water was 28 days, with an additional 28 days post treatment for observation of the rainbow trout fry ( Mehrle et al., 1988). The NOEC for 2,3,7,8 TCDF was reported to be 0.41 ng/ l. It appears that 2,3,7,8 TCDF levels in the trout may have reached equilibrium at the higher concentration ( 3.93 ng/ l), but it is not evident that equilibrium was achieved in 28 days at the lower concentration ( 0.41 ng/ l). Consequently, it is uncertain whether the only study available concerning 2,3,7,8 TCDF was of sufficient duration to produce maximum toxic effects. The behavioral effects induced by 2,3,7,8 TCDF were similar to those observed in 2,3,7,8 TCDD exposures; however, the observed responses were of lesser magnitude ( Mehrle et al., 1988).

C. Conclusions Concerning Aquatic Toxicity of CDDs and CDFs CDDs appear to be more toxic to fish than to aquatic invertebrates; however, the available toxicity values demonstrate that CDDs are very highly toxic to all the aquatic organisms that were tested. Many aquatic tests have been conducted with 2,3,7,8 TCDD, but most of these studies were either bioavailability tests or short term exposure studies with long post exposure observation periods. The available test data on 2,3,7,8 TCDD indicate that none of the studies are adequate to define an acceptable, no observed effect concentration ( NOEC). Only one study exposed a fish to 2,3,7,8 TCDD in water for a reasonable duration ( 28 days). Even then, the exposure duration in the rainbow trout study was only about half of the 48 days estimated to be required to achieve 90 percent of the steady state BCF level in fish. The NOEC was less than 0.038 ng/ l, a concentration which produced 45 percent mortality in rainbow trout fry.

A definitive NOEC has not been reported for 2,3,7,8 TCDD. Even the lowest test concentration ( 0.038 ng/ l) produced 45 percent mortality in rainbow trout exposed to 2,3,7,8 TCDF for 28 Page 9 of 65 days. The reported NOEC value for 2,3,7,8 TCDF is 0.41 ng/ l, but this is also uncertain because of the limited duration of observation. The toxicity data available on 2,3,7,8 TCDD, and possibly 2,3,7,8 TCDF, do not adequately define the inherent toxicity of these substances for two reasons: ( 1) the exposure periods are of insufficient duration for a steady state equilibrium to be reached; and ( 2) the studies do not address toxic effects on developing embryos resulting from deposition of either 2,3,7,8 TCDD or 2,3,7,8 TCDF in the eggs by the female.

Data on the toxicity of furans to aquatic organisms indicate that they are generally less toxic than dioxins, but 2,3,7,8 TCDF is still very highly toxic to fish [ 13 day LC50 value of 16 ng/ L in medaka ( Wisk & Cooper, 1990a)].

The most sensitive stage in the life of a rainbow trout for chlorinated chemicals appears to occur during resorption of the yolk sac in developing embryos. During this stage, the lipophilic substances ( e. g., CDDs and CDFs) deposited in yolk by the female become more concentrated in the yolk on a per unit weight basis and they are resorbed. This occurs as the fry metabolizes and incorporates nutrients contained in the yolk. As the toxic substance is released during metabolism of the yolk, the amount of substance entering the blood stream reaches a higher concentration than would otherwise normally be found. It is at this stage of development of organisms that the greatest sensitivity has been observed for substances such as DDE in both birds and fish.

IV. Toxicity Of CDDs and CDFs to Birds and Terrestrial Mammals The adverse effects to individual wildlife species from 2,3,7,8 TCDD have been documented in laboratory studies. Using the results of these studies to estimate effects on wild populations has limitations because the route and medium of administration and the duration of exposure to 2,3,7,8 TCDD for laboratory animals usually will differ from that of wild animals. Using these studies to assess effects on wild species assumes that the wild species are as sensitive or more sensitive to 2,3,7,8 TCDD than the laboratory species. The methodologies for predicting the effects of chemicals on terrestrial wildlife populations and ecosystems, however, are still in development.

A. Toxicity Assessment for Birds and Bird Eggs 1. Avian Toxicity Data Used in This Risk Assessment While there are no EPA Guideline studies submitted for CDDs/ CDFs, the literature data provide TCDD toxicity values for some comparable endpoints. These are provided in Table 2, below.

Table 2: Selected Avian Toxicity Values for 2,3,7,8 TCDD Species Test type/ duration Endpoint Toxicity Classification Citation Northern bobwhite Acute oral LD50= 15

g/ kg very highly toxic Hudson et al, 1984 Northern bobwhite 8 day dietary LC50= 167 ppt very highly toxic Kenaga and Norris, 1983 Page 10 of 65 Ringneck pheasant 10 week reproduction NOAEL= 0.014

g/ kg day ( fertility, embryo mortality) Nosek et al., 1992b and 1993 A full compilation of the toxicity data for 2,3,7,8 TCDD in birds and bird eggs from U. S. EPA ( 1990) is provided in Appendix 6. The data from U. S. EPA ( 1993) for birds are presented in Appendix 7.

2. Summaries of Additional Avian Toxicity Data 2,3,7,8 TCDD was administered at 100 ng/ kg body weight/ day in a corn oil/ acetone vehicle to 3 day old white leghorn chickens ( Schwetz et al., 1973). This dose was administered for 21 days and produced no adverse effects. It was assumed that the 2,3,7,8 TCDD was 100 percent absorbed from the corn oil/ acetone vehicle. However, it is possible that the absorption of 2,3,7,8 TCDD from laboratory feed or food sources for wild animals would not be the same as the assumed 100 percent absorption of 2,3,7,8 TCDD from a corn oil/ acetone vehicle.

Bird eggs can contain 2,3,7,8 TCDD transferred from the mother's body burden of 2,3,7,8 TCDD. Eggs are an important model to consider in determining a toxicity endpoint because of the sensitivity of eggs to 2,3,7,8 TCDD. Sullivan et al. ( 1987) concluded that the LOAEL for chicken embryos is 65 pg/ g in the egg ( 65 ppt), based on a study that found a 2 fold increase in cardiovascular malformations in chicken embryos at an estimated egg concentration of 65 pg/ g. Although effects were found at lower concentrations of 2,3,7,8 TCDD, the study concluded that the evidence for effects at these lower levels was inconclusive; thus, the 65 ppt value can be used in risk assessment context for comparison with predicted egg concentrations for wild species.

B. Toxicity Assessment for Terrestrial Mammals 1. Mammalian Data Used in This Risk Assessment While there are no EPA Guideline studies submitted for CDDs/ CDFs, the literature data provide TCDD toxicity values for some comparable endpoints. These are provided in Table 3, below.

Table 3: Selected Mammalian Toxicity Values for 2,3,7,8 TCDD Species Test type/ duration Endpoint Citation Mouse acute oral LD50= 114 284

g/ kg Kociba and Schwetz, 1982 Rat reproduction   exp osed days 6 15 of gestation NOAEL ( litter size and pup weight)= 0.125

g/ kg day Khera and Ruddick, 1973 A summary of toxic effects of 2,3,7,8 TCDD to wild mammals from U. S. EPA ( 1990) is presented in Appendix 8. Summary data from U. S. EPA ( 1993) are presented in Appendix 9. Page 11 of 65 2. Summaries of Some Additional Mammalian Data In a previous study, Murray et al. ( 1979) dosed rats at 100, 10, and 1 ng/ kg/ day through the diet, and studied the effects on subsequent generations. At the 10 ng/ kg/ day level, Murray et al. ( 1979) found decreased fertility in the f 1 and f 2 generations. A NOAEL level of 1 ng/ kg/ day was reported.

For larger mammals, the expected dose for wild species is compared to the lowest dose observed to produce adverse reproductive effects in rhesus monkeys. Schwetz et al. ( 1973) reported that rhesus monkeys were given 1.7 ng/ kg body weight of 2,3,7,8 TCDD in the diet. Of the seven pregnancies which occurred, four resulted in chemical induced abortions.

In both laboratory studies, it was assumed that absorption from a laboratory diet is similar to the absorption from a wild diet, and that these doses are directly comparable to the daily dose to wild species from the ingestion of prey items.

V. Ecological Hazard Conclusions In aquatic toxicity studies, the most toxic congener of polychlorinated dibenzo p dioxins and dibenzofurans appears to be 2,3,7,8 TCDD, probably due to its tendency to be taken up more readily than other congeners. Dioxins appear to be more toxic to fish than to aquatic invertebrates, but are considered to be very highly toxic to all aquatic organisms tested.

The mode of action of 2,3,7,8 TCDD is unknown, but it appears to be related to neural toxicity and may also affect the immune system, like many other organochlorine chemicals. TCDDinduced mortality typically occurs between 30 to 80 days after the initial exposure, even from exposure durations as short as 6 hours. Mortality in fish appears to be a function of both exposure duration and the test concentration.

Many sublethal effects have been reported for polychlorinated dioxins and furans. 2,3,7,8 TCDD elicits a broad range of toxic effects which include: reduced growth; weight loss; abnormal hatching; cellular alterations; numerous types of lesions in a broad spectra of organs; and behavioral effects on swimming, reduced feeding, and loss of response to external stimuli.

Continuous exposures to fish appear to produce toxic effects more rapidly than short term exposure to the same test concentration. Based on the available data, the most sensitive fish life stage is during yolk sac resorption in the developing embryo.

The 2,3,7,8 TCDD toxicological endpoints for fresh water fishes, marine and estuarine fishes, and invertebrates are provided in Appendices 10, 11, and 12, respectively.

The lowest NOEL for 2,3,7,8 TCDD exposure to birds is 100 ng/ kg/ day from an oral dose to white leghorn chickens over a 21 day exposure period. For wild mammals, a NOAEL of 1 ng/ kg/ day in the rat has been reported in a multi generation effects study.

The acute and chronic 2,3,7,8 TCDD toxicological endpoints for birds and mammals are summarized in Appendices 13, 14, 15, and 16, respectively. Page 12 of 65 VI. Environmental Risk Assessment Risk characterization integrates the results of the exposure and ecotoxicity data to evaluate the likelihood of adverse ecological effects. The means of this integration is called the quotient method. Risk quotients ( RQs) are calculated by dividing exposure estimates by acute and chronic ecotoxicity values.

RQ = EXPOSURE/ TOXICITY RQs are then compared to OPP's Levels of Concern ( LOCs). These LOCs are used by OPP to analyze 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 may be mitigated through restricted use classification, ( 3) acute endangered species  endangered species may be adversely affected, and ( 4) chronic risk  the potential for chronic risk is high regulatory action may be warranted. Currently, EFED does not perform assessments for chronic risk to plants, acute or chronic risks to nontarget insects, or chronic risk from granular/ bait formulations to birds or mammals.

The ecotoxicity test values ( measurement endpoints) used in the acute and chronic risk quotients are derived from required studies. Examples of ecotoxicity values derived from 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 long term 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 generally is used as the ecotoxicity test value in assessing chronic effects, although 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 assessing chronic effects to fish and aquatic invertebrates. However, the NOEC is used if the measurement end point is production of offspring or survival.

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

Table 4: Risk Presumptions for Terrestrial Animals Risk Presumption RQ LOC Birds Acute High Risk EEC1/ LC50 or LD50/ sqft2 or LD50/ day3 0.5 Acute Restricted Use EEC/ LC50 or LD50/ sqft or LD50/ day ( or LD50 < 50 mg/ kg) 0.2 Acute Endangered Species EEC/ LC50 or LD50/ sqft or LD50/ day 0.1 Chronic Risk EEC/ NOEC 1 Table 4: Risk Presumptions for Terrestrial Animals Risk Presumption RQ LOC Page 13 of 65 Wild Mammals Acute High Risk EEC/ LC50 or LD50/ sqft or LD50/ day 0.5 Acute Restricted Use EEC/ LC50 or LD50/ sqft or LD50/ day ( or LD50 < 50 mg/ kg) 0.2 Acute Endangered Species EEC/ LC50 or LD50/ sqft or LD50/ day 0.1 Chronic Risk EEC/ NOEC 1 1 abbreviation for Estimated Environmental Concentration ( ppm) on avian/ mammalian food items 2 mg/ ft2 3 mg of toxicant consumed/ day ( LD50 * wt. of bird)

Table 5: 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 Table 6: 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 1. Exposure and Risk to Nontarget Terrestrial Animals Page 14 of 65 Terrestrial organisms may be exposed to CDDs/ CDFs through direct contact with treated lumber and contact with CDDs/ CDFs 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 CDDs/ CDFs 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 CDDs/ CDFs leaching occurs from treated wood into the surrounding soil. The leached compounds in soil were then assumed to be taken up by terrestrial vegetation. Based on these conditions, the model assumed that birds and mammals would be exposed to CDDs/ CDFs 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 the " Environmental Fate Modeling RED Chapter for Dioxin in Technical Grade Pentachlorophenol". Using the state of Ohio to represent a typical regional environmental setting, U. S. EPA calculated the amount of CDDs/ CDFs released from an estimated 1,500,000 utility poles within the state. Based on the calculated release rates of CDDs/ CDFs via volatilization, wood erosion, and leaching, the fate and transport of CDDs/ CDFs releases were modeled using the Level III model. The amounts released were assumed to partition into the air, soil, water, and sediments. Amounts of individual CDD and CDF congeners in each compartment are reported in the Environmental Fate Modeling chapter of this document. Since the toxicity of dioxins and dioxin like compounds is known to be additive ( Tillit, 1999), the combined concentration of CDDs/ CDFs as Toxic Equivalents ( TEQ) was used as the EEC in this risk assessment, as well as the concentration of TCDD ( the most toxic congener) alone. According to the modeling results, terrestrial organisms would be exposed to 2.32e 02 mg/ kg dioxin TEQ in soil, 2.38E 04 mg/ kg of which is due to TCDD.. Risk Quotients ( RQ ) were calculated by comparing the toxicity endpoints for TCDD to the TCDD concentration, as well as the TEQ concentration.

The methods used to estimate the potential dose of CDD/ CDF to avian and mammalian species through the ingestion of vegetation and soil are provided in Appendix 17. A summary of the results is provided in the table below:

Table 7: Summary of Risk Quotients for Birds and Mammals Species Compound Soil Concentration ( mg/ kg) Total Dose ( EEC) ( mg/ kgday Endpoint RQ Bobwhite CDD/ CDF TEQ 2.32E 02 9.49E 04 TCDD LD50= 1.5E 05 mg/ kg ( Hudson et al. 1984) 63.32a TCDD NOAEL = 1.49E 05 mg/ kg day ( Schwetz et al., 1973) 67.78b TCDD 2.38E 04 9.70E 06 TCDD LD50 = 1.5E 05 mg/ kg ( Hudson et al. 1984) 0.65c Page 15 of 65 TCDD NOAEL = 1.49E 05 mg/ kg day ( Schwetz et al., 1973) 0.69 Meadow vole CDD/ CDF TEQ 2.32E 02 3.57E 03 TCDD LD50 0.09 mg/ kg ( Kociba and Schwetz, 1982, adjusted for mouse/ vole body weight) 0.04 TCDD NOAEL = 9.94E 04 ( Khera and Ruddick, 1973, adjusted for rat/ vole body weight) 3.59b TCDD 2.38E 04 3.66E 05 TCDD LD50 0.09 mg/ kg ( Kociba and Schwetz, 1982, adjusted for mouse/ vole body weight) 4.07E 04 TCDD NOAEL = 9.94E 04 ( Khera and Ruddick, 1973, adjusted for rat/ vole body weight) 0.04 a Exceeds the Acute High Risk LOC b Exceeds the Chronic Risk LOC c Exceeds the Restricted Use LOC i. Birds As shown in Table 7, when the TEQ concentrations of CDDs/ CDFs are used to determine the EEC, the resulting RQs exceed the LOC for acute high risk and chronic risk. When the single TCDD congener soil concentration is used to determine the EEC, the risk is lessened, but still exceeds the Restricted Used LOC for acute risk; however, using only the TCDD soil concentration to address exposure may underestimates the true exposure and risk to terrestrial mammals from CDDs/ CDFs from pentachlorophenol treated structures, as discussed below.

ii. Mammals Based on the results of the risk assessment, the TEQ soil concentration would result in an RQ that exceeds the chronic LOC for mammals. When only the TCDD soil concentration is considered, the risk is lessened, but this underestimates the true exposure and risk to terrestrial mammals from CDDs/ CDFs from pentachlorophenol treated structures, as discussed below.

iii. Risk Characterization for Birds and Mammals Page 16 of 65 Using only the TCDD soil concentration to estimate exposure to CDDs/ CDFs does not address the true exposure of these compounds. As described the Environmental Fate Modeling chapter, numerous CDD/ CDF congeners will be present in soil surrounding pentachlorophenol treated structures, and, while TCDD is the most toxic congener to birds and mammals, it is present in lower concentrations than many other congeners. Bird and mammal toxicity data for CDDs/ CDFs focuses primarily on TCDD, since it is recognized as the most toxic congener, but it is known that the toxicity of CDDs/ CDFs is additive ( Tillit, 1999). Focusing solely on TCDD to address exposure and risk would therefore likely underestimate the risks posed to terrestrial wildlife from CDDs/ CDFs. Even comparing the TEQ EEC to TCDD endpoints likely underestimates the risk, but methods to extrapolate TCDD toxicity to be representative of a combined dose of CDDs/ CDFs were not available at the time of this assessment. This underestimation may be offset, at least in part, by the conservative parameters of the model used. The model assumes that the entire diet of the bird or mammal consists of soil and vegetation from the immediate vicinity of a pentachlorophenol treated structure, and thus represents a " worst case" scenario. In reality, the foraging area of the animal would likely include areas that are not impacted by CDDs/ CDFs released from a treated structure, which would result in a lower actual exposure to CDDs/ CDFs. Data and models to quantify this were not available at the time of this assessment.

The risk assessment for birds and mammals was based solely on exposure to CDDs/ CDFs in soil. A quantitative assessment of the risks to birds and mammals from direct contact with the treated wood was not conducted due to the lack of exposure and toxicity data. As a result, the potential risks from direct contact with pentachlorophenol treated wood were not evaluated.

Considering the uncertainties discussed above, there are potential acute and chronic risks to birds and chronic risks to mammals from CDDs/ CDFs from pentachlorophenol treated wood, especially considering the tendency of CDDs/ CDFs to persist and bioaccumulate. Sensitive animals, such as endangered and threatened species, may be particularly at risk. There are also CDDs/ CDFs present in the environment from other sources, and those contributed by the use of pentachlorophenol are adding to the total environmental load and subsequent risk of these compounds B. Exposure and Risk to Fish and Aquatic Invertebrates Fish and aquatic invertebrates may potentially be exposed to CDDs/ CDFs leached from utility poles into surrounding soil, which then enters aquatic habitats via runoff. Most (> 99 %) CDD/ CDF will partition into the sediment, but a small amount will remain in the water column. Peak concentrations of TCDD/ TCDF ( from PRZM/ EXAMS modeling) are presented in Table 8, below. Detailed descriptions of the models used to predict these EECs can be found in the Environmental Fate Modeling chapter of this document.

Table 8: Peak concentrations ( 36 year) of TCDD, TCDF and TEQ in various aquatic media, as predicted by PRZM 3 and EXAMS models ( excerpted from Environmental Modeling chapter of this document) Page 17 of 65 2,3,7,8 TCDD TEQ Water column, total ( ng/ L) 3.68E 8 2.69E 6 Water column, dissolved ( ng/ L) 2.77E 9 6.45E 8 Water column sediment ( ng/ L) 1.10E 6 8.53E 5 Sediment total ( ng/ L) 1.13E 6 1.23E 4 Sediment, pore water diss. ( ng/ L) 2.83E 9 1.74E 7 Sediment, benthic ( ng/ g) 1.13E 6 1.23E 4 Sediment, benthic organism ( ng/ g) 1.81E 6 1.66E 4 Fish ( ng/ g lipid) 1.02E 6 1.73E 5 The peak total water column concentration of 2,3,7,8 TCDD was compared to the aquatic toxicity endpoints provided in Table 1. Since the toxicity of the various congeners of dioxins and furans is known to be additive ( Tillitt, 1999), the total concentration ( as TEQ) was also compared to the 2,3,7,8 TCDD endpoint; however, this may still underestimate the risk posed by CDDs/ CDFs since the toxicity endpoint is for only the most sensitive congener, and not a combination of congeners.

Table 9: TCDD and TEQ Risk Quotients for Aquatic Organisms Organism class Species Endpoint type and value RQ TCDD RQ TEQ Freshwater fish Rainbow trout, Oncorhynchus mykiss 96 hour acute = 1.83 ng/ L ( Bol et al., 1989) 2.01 E 08 1.40E 06 Freshwater fish Rainbow trout, Oncorhynchus mykiss 28 day NOEC = 1.1E 3 ng/ L ( Mehrle et al., 1988) 3.34 E 05 2.44E 03 Freshwater invertebrate Water flea, Daphnia magna 48 hour acute EC50 = 1020 ng/ L ( Adams et al., 1986) 3.61 E 11 2.64E 09 Freshwater invertebrate Snail ( exact species not reported) 12 day reproductive NOEC = 200 ng/ L ( Miller et al., 1973) 1.84 E 10 1.34E 08 All of these are well below any Level of Concern for aquatic organisms.

CDDs/ CDFs which are bound to sediment can be taken up by benthic organisms, which in turn are consumed by fish and higher order aquatic invertebrates. The bioaccumulative tendencies of CDDs/ CDFs could result in their concentrations magnifying through the food chain, eventually reaching toxic levels. The likelihood of this occurring with CDD/ CDF residues resulting from the use of pentachlorophenol is low, since the initial levels of CDDs/ CDFs entering soil, water and sediment are well below reported toxic levels. However, there are CDDs/ CDFs present in the Page 18 of 65 environment from other sources, and those contributed by the use of pentachlorophenol are adding to the total environmental load of these compounds.

C. Exposure and Risk to Plants There are no terrestrial or aquatic plant endpoints available to use in a risk assessment. A single dibenzofuran EC50 value of 1500 ppb for the diatom Skeletonema costatum was found in the ACQUIRE database. Since the endpoints for other aquatic organisms indicate that they are more sensitive than aquatic plants to dibenzofuran, it is assumed that the risk assessment for fish and aquatic invertebrates is sufficient to protect aquatic plant species as well.

D. Ecological Risk Conclusions Based on the quantitative estimates provided above, the acute and short term chronic risk to fish, aquatic organisms and terrestrial wildlife from pentachlorophenol related CDDs/ CDFs appears to be low. There is a great deal of uncertainty in these quantitative estimates, however, due to a lack of standard toxicity endpoints and exposure information. There are also uncertainties and limitations inherent to the exposure models used, as described in the Environmental Modeling chapter of this document. Based on the information that is available, it is unlikely that the CDDs/ CDFs resulting from the use of pentachorophenol on utility poles are posing an immediate risk to wildlife or aquatic organisms; however, it is possible that the buildup of these persistent and bioaccumulative compounds over time may eventually reach levels that pose a risk to these organisms. Additionally, pentachlorophenol is only one of many sources of CDDs/ CDFs in the environment; it is not currently possible to quantify the portion of the aggregate environmental risk from CDDs/ CDFs that is attributable to the use of pentachlorophenol on utility poles. Page 19 of 65 VII. REFERENCES Adams, W. J.; Blain, K. M. ( 1986) A water solubility determination of 2,3,7,8 TCDD. Chemosphere. 15: 1397 1400.

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Chemical Form Test Species Duration ( days) Concentration ( ng/ l) Toxic Effects Reference Exposure Observation Monochloro No Data Dichloro No Data Trichloro No Data Tetrachloro

1,2,3,7 TCDD Rainbow Trout ( fry  0.2 g) 5 ( water) + 24 Days 134 54 No deaths No deaths Muir et al. ( 1985)

Fathead Minnow ( 0.2 g) 5 ( water) + 24 Days 28 23 No deaths No deaths Muir et al. ( 1985)

Rainbow Trout ( 0.5  1.0g) 30 ( diet) + 30 Days 110 ng/ g No significant ( P= 0.05)

adverse effects Muir et al. ( 1988)

Fathead Minnow ( 1.0 g) 30 ( diet) + 30 Days 110 ng/ g No significant ( P= 0.05)

adverse effects Muir et al. ( 1988)

1,3,6,8 TCDD Rainbow Trout ( 0.1  0.5 g) 5 ( water) + 48 Days + 48 Days + 48 Days 211 74 4 No deaths No deaths No deaths Muir et al. ( 1986)

Fathead Minnow ( 1.5  2.5 g) 5 ( water) + 29 Days + 48 Days 41 10 No deaths No deaths Muir et al. ( 1986)

2,3,7,8 TCDD Fathead Minnow 20 ( water) + 20 Days 1.0 3 out of 30 dead Adams et al. ( 1986)

Fathead Minnow 1,2,3,4 ( water + 60 Days 0.0 0.12 10% dead < 5% dead Adams et al. ( 1986) Appendix 1. Compilation of Aquatic Toxicity Data for Chlorinated Dibenzo p Dioxins ( CDDs) ( U. S. EPA, 1990) ( continued)

Chemical Form Test Species Duration ( days) Concentration ( ng/ l) Toxic Effects Reference Exposure Observation Page 30 of 65 Fathead Minnow 28 ( water) + 20 Days 1.7 6.7 63 53% dead ( 28 days)

100% dead ( 22 days)

100% dead ( 12 days)

28 d LC50 1.7 ng/ L Adams et al. ( 1986)

2,3,7,8 TCDD ( continued) Coho Salmon 1 ( water) 0.056 Deaths after several weeks Miller et al. ( 1973)

Rainbow Trout ( 35 g) 0.25 ( water) + 139 Days 107 4 Deaths, one each on Days 78, 136, 137, 139 and decreased growth Branson et al. ( 1985)

Rainbow Trout ( yolk sac fry) 4 ( water) + 16 Days 54.6 33.1 20.1 7.4 4.5 2.7 1.6 100% dead ( 13 days)

100% dead ( 15 days)

100% dead ( 17 days)

95% dead ( 20 days)

63% dead ( 20 days)

20% dead ( 20 days)

4% dead ( 20 days) Helder and Seinen ( 1985 Rainbow Trout ( fry  0.38 g) 28 ( water) + 28 Days 0.789 0.382 0.176 0.079 0.038 0.0011 ( control) 85% dead ( 28 days)

73% dead ( 28 days)

50% dead ( 28 days)

95% dead ( 56 days)

18% dead ( 28 days)

83% dead ( 56 days)

6% dead ( 28 days)

45% dead ( 56 days)

5% dead ( 28 days)

7% dead ( 56 days) Mehrle et al. ( 1988)

Daphnia magna 2 ( water) + 7 Days 0.2  1,030 No mortality 48 hr EC50 > 1,030 ng/ L Adams et al. ( 1986) Appendix 1. Compilation of Aquatic Toxicity Data for Chlorinated Dibenzo p Dioxins ( CDDs) ( U. S. EPA, 1990) ( continued)

Chemical Form Test Species Duration ( days) Concentration ( ng/ l) Toxic Effects Reference Exposure Observation Page 31 of 65 Rainbow Trout 91 ( diet) + 91 Days 494 No overt effects Kleeman et al. ( 1986a)

Yellow Perch 91 ( diet) + 91 Days 494 No overt effects Kleeman et al. ( 1986b)

2,3,7,8 TCDD ( continued) Rainbow Trout 1 dose intraperitoneal injection ( i. p.) + 14 Days 10 ng/ kg 1.0 0.1 0.01 Hypophagic, etc.

Normal Normal Normal Spitsbergen ( 1986)

Rainbow Trout ( LC50  10

g/ kg) 1 dose ( i. p.) + 80 Days 125 25 5 1 95% dead ( 80 days)

90% dead ( 80 days)

20% dead ( 80 days)

0% dead ( 80 days) Spitsbergen et al. ( 1988a)

Rainbow Trout ( LD50  3

g/ kg) 1 dose ( i. p.) + 80 Days 125 25 5 1 95% dead ( 22 days)

100% dead ( 40 days)

80% dead ( 75 days)

8% dead ( 80 days) Kleeman et al. ( 1986a)

Carp ( LD50  3

g/ kg) 1 dose ( i. p.) + 80 Days 125 25 5 1 100% dead ( 45 days)

100% dead ( 55 days)

90% dead ( 75 days)

5% dead ( 80 days) Kleeman et al. ( 1988)

Bullhead ( LD50  5

g/ kg) 1 dose ( i. p.) + 80 Days 125 25 5 1 100% dead ( 22 days)

100% dead ( 22 days)

50% dead ( 65 days)

3% dead ( 80 days) Kleeman et al. ( 1988) Appendix 1. Compilation of Aquatic Toxicity Data for Chlorinated Dibenzo p Dioxins ( CDDs) ( U. S. EPA, 1990) ( continued)

Chemical Form Test Species Duration ( days) Concentration ( ng/ l) Toxic Effects Reference Exposure Observation Page 32 of 65 Largemouth Bass ( LD50  11

g/ kg) 1 dose ( i. p.) + 80 Days 125 25 5 1 100% dead ( 22 days)

95% dead ( 80 days)

5% dead ( 80 days)

5% dead ( 80 days) Kleeman et al. ( 1988)

Bluegill ( LD50  16

g/ kg) 1 dose ( i. p.) + 80 Days 125 25 5 1 100% dead ( 23 days)

80% dead ( 76 days)

2% dead ( 80 days)

5% dead ( 80 days) Kleeman et al. ( 1988)

2,3,7,8 TCDD ( continued) Yellow Perch ( LD50  10

g/ kg) 1 dose ( i. p.) + 80 Days 125 25 5 1 100% dead ( 35 days)

90% dead ( 40 days)

20% dead ( 60 days)

5% dead ( 80 days) Kleeman et al. ( 1986b)

Yellow Perch ( LC50  3

g/ kg) 1 dose ( i. p.) + 80 Days 125 25 5 1 < 95% dead ( 28 days)

< 95% dead ( 28 days)

80% dead ( 80 days)

0% dead ( 80 days) Spitsbergen et al. ( 1988b)

Pentachloro

1,2,3,4,7 PCDD Rainbow Trout ( fry  0.2 g) 5 ( water) + 24 Days 16 No deaths Muir et al. ( 1985)

Fathead Minnow ( 0.2 g) 5 ( water) + 24 Days + 24 Days 19 11 No deaths No deaths Muir et al. ( 1985)

Rainbow Trout ( 0.5  1.0 g) 30 ( diet) + 30 Days 105 ng/ g No significant ( P= 0.05)

adverse effects Muir et al. ( 1988)

Fathead Minnow ( 1.0 g) 30 ( diet) + 30 Days 105 ng/ g No significant ( P= 0.05)

adverse effects Muir et al. ( 1988) Appendix 1. Compilation of Aquatic Toxicity Data for Chlorinated Dibenzo p Dioxins ( CDDs) ( U. S. EPA, 1990) ( continued)

Chemical Form Test Species Duration ( days) Concentration ( ng/ l) Toxic Effects Reference Exposure Observation Page 33 of 65 Hexachloro

1,2,3,4,7,8 HxCDD Rainbow Trout ( fry  0.2g) 5 ( water) + 24 Days 47 26% dead ( 12 days) and weight loss Muir et al. ( 1985)

Rainbow Trout ( fry  0.2g) 5 ( water) + 48 Days 10 No effects Muir et al. ( 1985)

Fathead Minnow ( 0.2 g) 5 ( water) + 24 Days + 48 Days 18 7 No deaths No deaths Muir et la. ( 1985)

1,2,3,4,7,8 HxCDD ( continued) Rainbow Trout ( 0.5  1.0 g) 30 ( diet) + 30 Days 109 ng/ g Significant ( P= 0.05)

growth rate effects Muir et al. ( 1988)

Fathead Minnow ( 1.0 g) 30 ( diet) + 30 Days 109 ng/ g No significant ( P= 0.05)

adverse effects Muir et al. ( 1988)

1,2,3,7,8,9 HxCDD Rainbow Trout ( yolk sac fry) 4 ( water) + 10 Days Less than 10 5 times as toxic as 2,3,7,8 TCDD Helder and Seinen ( 1986)

Heptachloro 1,2,3,4,6,7,8 HpCDD Rainbow Trout ( fry  0.2 g) 5 ( water) + 24 Days + 48 Days 55 11 No deaths No deaths Muir et al. ( 1985)

Fathead Minnow ( 0.2 g) 5 ( water) + 24 Days + 48 Days 39 8 No deaths No deaths Muir et al. ( 1985)

Rainbow Trout ( 0.5  1.0 g) 30 ( diet) + 30 Days 109 ng/ g No significant ( P= 0.05)

adverse effects Muir et al. ( 1988)

Fathead Minnow ( 1.0 g) 30 ( diet) + 30 Days) 109 ng/ g No significant ( P= 0.05)

adverse effects Muir et al. ( 1988)

Octachloro Appendix 1. Compilation of Aquatic Toxicity Data for Chlorinated Dibenzo p Dioxins ( CDDs) ( U. S. EPA, 1990) ( continued)

Chemical Form Test Species Duration ( days) Concentration ( ng/ l) Toxic Effects Reference Exposure Observation Page 34 of 65 Rainbow Trout ( 0.1  0.5 g) 5 ( water) + 18 Days + 32 Days 415 20 No deaths No deaths Muir et al. ( 1986)

Fathead Minnow ( 1.5  2.5 g) 5 ( water) + 48 Days 9 No deaths Muir et al. ( 1986)

Source: U. S. EPA, 1990. Page 35 of 65 Appendix 2. Summary of the Toxic Effects of 2,3,7,8 TCDD to Aquatic Life ( U. S. EPA, 1993)

Test Species Test Method Water Conc.

( ng/ L) a Organism Conc.

( pg/ g) b Duration Effect Reference Exposure Observation AQUATIC LIFE Freshwater Species Algae, Oedogonium cardiacum Model ecosystem 1,330 2,295,000 33 d No toxic effect Isensee and Jones, 1975;

Isensee, 1978 Vascular plant Duckweed,

Lemna minor Model ecosystem 1,330 33 d No toxic effect Isensee and Jones, 1975;

Isensee, 1978 7.13 30,700 33 d No toxic effect Isensee and Jones, 1975 Annelid Worm, Paranais sp. Water ( static) 200c 55 d No decrease in reproductive success Miller et al., 1973 Mollusc Snail ( adult),

Physa sp. Model ecosystem 1,330 502,000 33 d No toxic effect Isensee and Jones, 1975;

Isensee, 1978 Water ( static) 200c 36 d 12 d No decrease in reproductive success Miler et al., 1973 Arthropod Mosquito ( larvae),

Aedes aegypti Water ( static) 200c 17 d 23 d No effect on pupation Miller et al., 1973 Cladoceran ( adult),

Daphnia magna Model ecosystem 1,330 1,570,000 33 d No toxic effect Isensee and Jones, 1975 Cladoceran ( 1 21 d),

Daphnia magna Water ( renewal) 1,030 48 h 7 d No toxic effect Adams et al., 1986 Fish Coho salmon, Oncorhynchus kisutch Juvenille ( 3.5 g) Water ( static) 0.56 96 h 114 d No toxic effect Miller et al., 1979 Appendix 2. Summary of the Toxic Effects of 2,3,7,8 TCDD to Aquatic Life ( continued)

Test Species Test Method Water Conc.

( ng/ L) a Organism Conc.

( pg/ g) b Duration Effect Reference Exposure Observation Page 36 of 65 5.60 96 h 56 d 50% mortality Miller et al., 1973, 1979 Rainbow trout,

Oncorhynchus mykiss Eggs Water ( renewal) 0.10c 96 h 160 d Delayed development,

reduced growth of fry Helder 1981; 1982a, b 1c 96 h 160 d Reduced growth,

mortality in sac fry Helder, 1981; 1982a, b 10c 96 h 40 d 100% mortality in sac fry Helder, 1982a, b Egg injection 230 ( in eggs) Single injection Fertilized egg to swim up LR50 ( sac fry of McConaughy strain) Walker and Peterson,

1991 240 ( in eggs) Single injection Fertilized egg to swim up LR50 ( sac fry of Erwin strain) Walker and Peterson,

1991 374 ( in eggs) Single injection Fertilized egg to swim up LR50 ( sac fry of Arlee strain) Walker and Peterson,

1991 488 ( in eggs) Single injection Fertilized egg to swim up LR50 ( sac fry of Eagle Lake strain) Walker and Peterson,

1991 421 ( in eggs) Single injection > 48 h to post swim up LR50 ( sac fry of Fish Lake strain) Walker et al., 1992 Water ( renewal) 279 ( in eggs) 48 h > 48 h to post swim up Significant mortality in sac fry Walker et al., 1992 439 ( in eggs) 48 h > 48 h to post swim up LR50d ( sac fry) Walker et al., 1992 Sac fry Water ( renewal) 1c 96 h 160 d Reduced growth,

mortality Helder 1981; 1982a, b 10c 96 h 10 d 100% mortality Helder 1982a, b 12.2c 96 h 16 d 100% mortality Helder and Seinen, 1985 1.83c 96 h 21 d LC50 Bol et al., 1989 Appendix 2. Summary of the Toxic Effects of 2,3,7,8 TCDD to Aquatic Life ( continued)

Test Species Test Method Water Conc.

( ng/ L) a Organism Conc.

( pg/ g) b Duration Effect Reference Exposure Observation Page 37 of 65 Swim up fry ( 0.38 g) Water ( flow thru) 0.176 3,220 28 d 28 d 95% mortality Mehrle et al., 1988 0.0011 21g 28 d 28 d NOAELe Mehrle et al., 1988 0.038 765g 28 d 28 d LOAELf ( 45% mortality) Mehrl et al., 1988 0.046 28 d 28 d LC50 Mehrle et al., 1988 Juvenile ( 0.85 g) Water ( renewal) 10 96 h 68 d Reduced growth,

mortality Helder 1981; 1982a, b 100 96 h 23 d 100% mortality Helder 1981; 1982a, b Fingerling ( 25 45 g) i. p. injection 1,000c Single injection 25 d Significant hematological changes Spitsbergen et al., 1988a 5,000c Single injection 20 d 20% mortality Spitsbergen et al., 1988a 5,000c Single injection 11 12 wk 20% mortality, increased liver weight van der Weiden et al.,

1990 10,000c Single injection 80 d LD50 Spitsbergen et al.,

1988a; Kleeman et al.,

1988 Fingerling ( 35 g) Water ( static) 107 650 2,580 6 h 42 139 d Mortality, fin rot,

increased liver weight Branson et al., 1985 Fingerling ( 7.8 cm) Diet ( 3.29 ng/ g) 314 71 d No effect on survival and growth Hawkes and Norris,

1977 Diet ( 1,700 ng/ g) 276,000 71 d 100% mortality Hawkes and Norris,

1977 Fingerling ( 3 7 g) Diet ( 0.494 ng/ g) 250 13 wk 13 wk No toxic effect Kleeman et al., 1986a Fingerling ( 8 g)

Yearling ( 100 150 g) i. p. injection 10,000c Single injection 2 4 wk post exposure Fin necrosis, no effect on immune suppression Spitsbergen et al., 1986;

1988c Appendix 2. Summary of the Toxic Effects of 2,3,7,8 TCDD to Aquatic Life ( continued)

Test Species Test Method Water Conc.

( ng/ L) a Organism Conc.

( pg/ g) b Duration Effect Reference Exposure Observation Page 38 of 65 Juvenile ( 46 g) i. p. injection 300 3,060 Single injection 6 12 wk Fin hemorrhage, spleen histopathology, EROD induction, P4501A1 induction van der Welden et al.,

1992 790 Single injection 3 wk ED50 for EROD induction van der Welden et al.,

1992 Immature Adult ( 300

400 g) i. p. injection 640c Single injection 72 h ED50 for AHH induction Janz and Metcalfe, 1991 Lake trout, Salvelinus namaycush Eggs Water ( renewal) 34 ( in eggs) 48 h > 48 h to post swim up NOAEL Walker et al., 1991 40 ( in eggs) 48 h > 48 h to post swim up 23% mortality in sac fry Spitsbergen et al., 1991 55 ( in eggs) 48 h > 48 h to post swim up LOAEL ( sac fry mortality) Walker et al., 1991 65 ( in eggs) 48 h > 38 h to post swim up LR50 ( sac fry) Walker et al., 1991 Egg injection 47 ( in eggs) Single injection Fertilized egg to swim up fry LR50 ( sac fry) Walker et al., 1992 Adult Dieth 59 ( in eggs) 90 d Eggs to swimup fry LR50 ( sac fry) Walker, 1991; Walker et al., 1993 104 ( in eggs) 90 d Eggs to swimup fry 100% mortality to sac fry Walker, 1991; Walker et al., 1993 Northern pike, Esox lucius Eggs Water ( renewal) 0.1c 96 h 72 d Delayed hatch, reduced growth of fry Helder, 1980; 1982a, b 1.0 96 h 53% mortality to fry Helder, 1980; 1982a, b 10.0 96 h 99% mortality to fry Helder, 1980; 1982a, b Appendix 2. Summary of the Toxic Effects of 2,3,7,8 TCDD to Aquatic Life ( continued)

Test Species Test Method Water Conc.

( ng/ L) a Organism Conc.

( pg/ g) b Duration Effect Reference Exposure Observation Page 39 of 65 Carp, Cyprinus carpio Juvenile ( 20 g) i. p. injection 3,000c Single injection 80 d LD50 Kleeman et al., 1988 Adult Water ( flow thru) 0.060 2,200 71 d 61 d Mortality and pathology Cook et al., 1991 Zebrafish, Brachydanio revio Adult Diet ( 1.7 ng/ g) Single dose 22 d No effect Wannemacher et al.,

1992 Diet (

8.3 ng/ g) Single dose 1 2 Spawnings Reduced eggs per spawn,

100% larval mortality Wannemacher et al.,

1992 Fathead minnow, Pimephales promelas Juvenile ( 1.0 g) Water ( flow thru) 0.049

0.067 71 d 61 d Mortality and pathology Cook et al., 1991 1.7 28 d LC50 Adams et al., 1986 Juvenile ( 1.0 2.0 g) Water ( static) 7.1 24 h 60 d 40% mortality Adams et al., 1986 Bullhead, Ictalurus melas Juvenile ( 6 g) i. p. injection 5,000c Single injection 80 d LD50 Kleeman et al., 1988 Channel catfish, Ictalurus punctatus Fingerling Model ecosystem 2.4 4.2 15 20 d 100% mortality Yockim et al., 1978 Appendix 2. Summary of the Toxic Effects of 2,3,7,8 TCDD to Aquatic Life ( continued)

Test Species Test Method Water Conc.

( ng/ L) a Organism Conc.

( pg/ g) b Duration Effect Reference Exposure Observation Page 40 of 65 Japanese medaka, Oryzias latipes Eggs S, M 3.5 6.0 Fertilized egg to 3 d post hatch EC50 ( embryos with lesions) Wisk and Copper,

1990a, b 9.0 13.0 Fertilized egg to 3 d post hatch LC50 Wisk and Cooper,

1990a, b 14.0 15.0 Fertilized egg to 3 d post hatch EC50 ( embryos with severe lesions) Wisk and Cooper,

1990a, b 14.0 17.0 Fertilized egg to 3 d post hatch EC50 ( prevent hatch) Wisk and Cooper,

1990a, b Water ( static) 240 ( in embryos) Fertilized egg to 3 d post hatch ER50i ( embryos with lesions) Wisk and Cooper, 1990b Mosquito fish,

Gambusia affinis Model ecosystem 2.4 4.2 15 d 100% mortality Yockim et al., 1978 Guppy, Poecilia reticulata Juvenile ( 8 12) Water ( static) 0.1c 24 h 42 d Significant increase in fin necrosis Miller et al., 1979 Juvenile ( 9 40 mm) Water ( static) 100c 120 h 37 d 100% mortality Norris and Miller, 1974 Bluegill, Lepomis macrochirus Juvenile ( 30 g) i. p. injection 16,000c Single injection 80 d LD50 Kleeman et al., 1988 Largemouth bass, Micropterus salmoides Juvenile ( 7 g) i. p. injection 11,000c Single injection 80 d LD50 Kleeman et al., 1988 Appendix 2. Summary of the Toxic Effects of 2,3,7,8 TCDD to Aquatic Life ( continued)

Test Species Test Method Water Conc.

( ng/ L) a Organism Conc.

( pg/ g) b Duration Effect Reference Exposure Observation Page 41 of 65 Yellow perch, Perca flavescens Juvenile ( 3 6 g) Diet ( 0.494 ng/ g) 143 13 wk 13 wk No toxic effect Kleeman et al., 1986b Juvenile i. p. injection 3,000c Single injection 80 d LD50 Spitsbergen et al.,

1988b; Kleeman et al.,

1988 Amphibian Bullfrog, Rana catesbeiana Tadpole i. p. injection 1,000,000c Single injection 50 d No effect on metamorphosis Beatty et al., 1976 Adult i. p. injection 500,000c Single injection 35 d No toxic effect Beatty et al., 1976 Saltwater Species Rays Little Skate, Raja erinacea 500 1,100 g i. p. injection 1,000c Single injection 10 d Increased enzyme activity Bend et al., 1974 4,500c Two injections 7 12 d 10 fold increase in enzyme activity Pohl et al., 1974 Fish, Mummichog, Fundulus heteroclitus Eggs Water ( static) 200c Fertilized egg to hatch 20% mortality and 50%

lesions in embryos Cooper, 1989; Prince and Cooper, 1989 Winter flounder, Pleuronectes americanus 250 g Oral dose 4,500c Two doses 8 d Increased enzyme activity Pohl et al., 1974 a Measured TCDD concentration in water.

b Measured TCDD concentration in organism ( wet weight).

c Unmeasured TCDD concentration in water or organism ( wet weight).

d LR50 ( corrected for control mortality) term defined in this report as the measured residue concentration in eggs that caused 50% mortality to sac fry.

e NOAEL = No observed adverse effect level.

f LOAEL= Lowest observed adverse effect level. Appendix 2. Summary of the Toxic Effects of 2,3,7,8 TCDD to Aquatic Life ( continued)

Page 42 of 65 g NOAEL and LOAEL values ( based on mean measured wet weight organism concentrations) were calculated for this report.

h Diet consisted of 22 ng/ g pelletized feed followed by fathead minnows injected with 500 pg/ fish.

i ER50  Term defined in this report as the measured residue concentration in eggs that caused 50% effect. Page 43 of 65 Appendix 3. Summary of Aquatic Toxicity Data for CDDs from the ACQUIRE Database Test Loc CAS #/ Chemical Latin name, Common name Endpoint Effect Water Type Dur ( days)/

Exp Typ Conc ( ug/ L) Ref.

LAB 1746016, 2,3,7,8 Tetrachlorodibenzo[ b, e][ 1,4] dioxin Ameiurus melas, Black bullhead LD50 MOR FW 80.00/ I F 5 ug/ kg Kleeman et al., 1988 Cyprinus carpio, Common, mirror, colored, carp ED50 ENZ FW 7.00/ I F 0.048 ug/ kg Van der Weiden et al.,

1994a ED50 ENZ FW 14.00/ I F 0.069 mmol/ kg ED50 ENZ FW 14.00/ I F 0.084 mmol/ kg ED50 ENZ FW 14.00/ I F 0.166 mmol/ kg ED50 ENZ FW 14.00/ I F 0.059 mmol/ kg LD50 MOR FW 80.00/ I F 3 ug/ kg Kleeman et al., 1988 LOEC ENZ FW 7.00/ I F 0.30 ug/ kg Van der Weiden et al.,

1994a LOEC ENZ FW 7.00/ I F 0.03 ug/ kg LOEC ENZ FW 14.00/ I F 0.03 ug/ kg LOEC ENZ FW 14.00/ I F 0.03 ug/ kg LOEC ENZ FW 14.00/ I F 0.03 ug/ kg LOEC ENZ FW 14.00/ I F 0.03 ug/ kg NOEC ENZ FW 7.00/ I F 0.01 ug/ kg Fundulus heteroclitus, Mummichog ED50 PHY SW 27.00/ S F 2025 pg/ g Prince& Cooper, 1995 Lepomis macrochirus, Bluegill LD50 MOR FW 80.00/ I F 16 ug/ kg Kleeman et al., 1988 Micropterus salmoides, Largemouth bass LD50 MOR FW 80.00/ I F 11 ug/ kg LAB 1746016, 2,3,7,8 Tetrachlorodibenzo[ b, e][ 1,4] dioxin Oncorhynchus mykiss, Rainbow trout, donaldson trout ED50 BIO FW 7.00/ I F 0.17 ug/ kg Newsted & Geiger, 1993 ED50 BIO FW 7.00/ I F 0.79 ug/ kg ED50 CYT FW 7.00/ I F 0.5  5.0 ug/ kg ED50 ENZ FW 10.00/ I F 1.00 ug/ kg ED50 ENZ FW 20.00/ I F 0.85 ug/ kg ED50 ENZ FW 21.00/ I A 0.79 ug/ kg Van der Weiden et al.,

1992 ED50 ENZ FW 5.00/ I F 1.05 ug/ kg Newsted et al., 1995 ED50 ENZ FW 7.00/ I F 0.73 ug/ kg ED50 ENZ FW 7.00/ I F 0.66 ug/ kg ED50 ENZ FW 3.00/ R F 3 ug/ kg Harris et al., 1994 LD50 MOR FW 30.00/ I F 242 pg/ g Zabel & Peterson, 1996 LD50 MOR FW 30.00/ I F 409 pg/ g LD50 MOR FW ~ 30.00/ I A > 250  < 300 pg/ g LD50 MOR FW ~ 30.00/ I A > 200  < 300 pg/ g LD50 MOR FW ~ 30.00/ I A > 300  < 350 pg/ g LD50 MOR FW ~ 30.00/ I A > 400  < 500 pg/ g LD50 MOR FW ~ 30.00/ I A > 300  < 400 pg/ g LD50 MOR FW ~ 30.00/ I F 0.171 ng/ g LD50 MOR FW ~ 30.00/ I F 0.261 ng/ g Test Loc CAS #/ Chemical Latin name, Common name Endpoint Effect Water Type Dur ( days)/

Exp Typ Conc ( ug/ L) Ref.

Page 44 of 65 LD50 MOR FW ~ 30.00/ I F 0.341 ng/ g LD50 ~ MOR FW 2.00/ S F 439 pg/ g Walker et al., 1992 LD50 MOR FW 7.00/ I F 0.240 Walker & Peterson., 1991 LD50 MOR FW 7.00/ I F 0.374 LD50 MOR FW 7.00/ I F 0.488 LD50 MOR FW 7.00/ I F 0.230 LD50 MOR FW 80.00/ I F 10 ug/ kg Spitsbergen, et al., 1988a LD50 MOR FW 80.00/ I F 10 ug/ kg Kleeman et al., 1988 LOEC BIO FW 7.00/ I F 0.01 ug/ kg Newsted and Giesy, 1993 LOEC ENZ FW 10.00/ I F 0.5 ug/ kg Newsted et al., 1995 LOEC ENZ FW 2.00  16.00/ D A 0.072 ug/ kg Parrot et al., 1995 LOEC ENZ FW 20.00/ I F 0.1 ug/ kg Newsted et al., 1995 LOEC ENZ FW 5.00/ I F 0.5 ug/ kg LOEC ENZ FW 7.00/ I F 0.1 ug/ kg Newsted and Giesy, 1993 NOEC ENZ FW 7.00/ I F 0.01 ug/ kg NOEC GRO FW 56.00/ F F < 0.000038 Mehrle et al., 1988 NOEC MOR FW 56.00/ F F < 0.000038 LAB 1746016, 2,3,7,8 Tetrachlorodibenzo[ b, e][ 1,4] dioxin Oryzias latipes, Medaka, high eyes EC50 ABN FW > 13.00/ S F 0.006 Wisk & Cooper, 1990a EC50 ABN FW > 13.00/ S F 0.015 EC50 ABN FW 3.00/ S F 3500 EC50 ABN FW 3.00/ S F 14000 EC50 DVP FW 5.00  6.00/ S F 0.012 Harris et al., 1994 EC50 HAT FW 3.00/ S F 14000 Wisk & Cooper, 1990b LC50 MOR FW > 13.00/ S F 0.013 Wisk & Cooper, 1990a LC50 MOR FW 17.00/ S F 0.020 Harris et al., 1994 LC50 MOR FW 0.0  17.00/ S F 0.0057 Metcalfe et al., 1997 LC50 MOR FW 6.00/ S F 9000 Wisk & Cooper, 1990b Perca flavescens, Yellow perch LD50 MOR FW 80.00/ I F 3 ug/ kg Spitsbergen, et al., 1988b LD50 MOR FW 80.00/ I F 3 ug/ kg Kleeman et al., 1988b Platichthys flesus, Starry, european flounder ED50 ENZ SW 10.00/ D F 1.62 ug/ kg Besselink et al., 1997 Poecilia reticulata, Guppy LT50 ~ MOR FW 11.60/ S F 1.0 Norris & Miller, 1974 LT50 ~ MOR FW 18.20/ S F 10.0 LT50 ~ MOR FW 21.70/ S F 0.1 Salvelinus namaycush, Lake trout, siscowet LC50 ~ MOR FW 2.00/ R F 85 pg/ g Zabel et al., 1995c LD50 ~ MOR FW 2.00/ R F 0.065 Walker et al,., 1991 LD50 ~ MOR FW 2.00/ S F 65 pg/ g Walker et al., 1992 LD50 MOR FW F 65 pg/ g Walker, 1992 Test Loc CAS #/ Chemical Latin name, Common name Endpoint Effect Water Type Dur ( days)/

Exp Typ Conc ( ug/ L) Ref.

Page 45 of 65 LD50 MOR FW F 47 pg/ g LOEC ~ GRO FW 2.00/ R F 0.055 Walker et al,., 1991 LOEC ~ GRO FW 2.00/ R F 0.226 LOEC ~ MOR FW 2.00/ R F 0.055 LOEC MOR FW F 40 pg/ g Walker, 1992 NOEC ~ MOR FW 2.00/ R F 0.034 Walker et al,., 1991 NOEC MOR FW F 34 pg/ g Walker, 1992 LAB 3268879, Octachlorodibenzo[ b, e][ 1,4] dioxin Brachydanio rerio, Zebra danio, zebrafish ET50 HAT FW 3.48/ F A 0.033 Berends et al., 1997 ET50 HAT FW 3.54/ F A 0.031 LAB 34465468, Hexachlorodibenzo[ b, e][ 1,4] dioxin Cyprinus carpio, Common, mirror, colored, carp ED50 ENZ FW 14.00/ I F 0.500 mmol/ kg Van der Weiden et al.,

1994b ED50 ENZ FW 14.00/ I F 0.536 mmol/ kg LOEC ENZ FW 14.00/ I F 0.31 ug/ kg LOEC ENZ FW 14.00/ I F 1.53 ug/ kg Oncorhynchus mykiss, Rainbow trout, donaldson trout LD50 MOR FW ~ 30.00/ I F 8.58 ng/ g Zabel et al, 1995b LAB 34816530, 1,2,7,8, Tetrachlorodibenzo[ b, e][ 1,4] dioxin Oryzias latipes, Medaka, high eyes EC50 ABN FW > 13.00/ S F 19.000 Wisk & Cooper, 1990a EC50 ABN FW > 13.00/ S F > 50.000 LC50 MOR FW > 13.00/ S F > 50.000 LAB 35822469, 1,2,3,4,6,7,8 Heptachlorodibenzo[ b, e][ 1,4] dioxin Oncorhynchus mykiss, Rainbow trout, donaldson trout LD50 MOR FW ~ 30.00/ I F 110 ng/ g Zabel et al, 1995b LOEC ENZ FW 2.00  16.00/ D A 0.74 ug/ kg Parrot et al., 1995 LAB 39001020, 1,2,3,4,5,6,7,8 Octachlorodibenzofuran Brachydanio rerio, Zebra danio, zebrafish ET50 HAT FW 3.43/ F A 0.032 Berends et al., 1997 ET50 HAT FW 3.54/ F A 0.036 LAB 40321764, 1,2,3,7,8 Pentachlorodibenzo[ b, e][ 1,4] dioxin Cyprinus carpio, Common, mirror, colored, carp ED50 ENZ FW 14.00/ I F 0.025 mmol/ kg Van der Weiden et al.,

1994b ED50 ENZ FW 14.00/ I F 0.025 mmol/ kg LOEC ENZ FW 14.00/ I F 0.002 ug/ kg LOEC ENZ FW 14.00/ I F 0.01 ug/ kg Oncorhynchus mykiss, Rainbow trout, donaldson trout LD50 MOR FW ~ 30.00/ I A > 450  < 500 pg/ g Zabel, et al., 1995a LD50 MOR FW 7.00/ I F 0.566 Walker & Peterson., 1991 LOEC ENZ FW 2.00  16.00/ D A 0.039 ug/ kg Parrot et al., 1995 Oryzias latipes, Medaka, high eyes EC50 ABD FW > 13.00/ S F 0.013 Wisk & Cooper, 1990a EC50 ABD FW > 13.00/ S F 0.032 LC50 MOR FW > 13.00/ S F 0.027 LAB 57653857, 1,2,3,4,7,8 Hexachlorodibenzo[ b, e][ 1,4] dioxin Oncorhynchus mykiss, Rainbow trout, donaldson trout LD50 MOR FW 7.00/ I F 1.427 Walker & Peterson., 1991 LOEC ENZ FW 2.00  16.00/ D A 0.25 ug/ kg Parrot et al., 1995 Oryzias latipes, Medaka, high eyes EC50 ABN FW > 13.00/ S F 0.200 Wisk & Cooper, 1990a EC50 ABN FW > 13.00/ S F 1.100 LC50 MOR FW > 13.00/ S F 2.9 Page 46 of 65 Endpoint: LC50 Median lethal concentration. Used only when death is observed endpoint. TLms and TL50s ~ LC50 EC50 Median effective concentration. Used when an effect other than death is the observed endpoint.

NOEC No observed effect concentration ED50 Effective dose LD50 Lethal dose LOEC Lowest observed effect concentration LT50 Median lethal concentration. Used only when death is observed endpoint. TLms and TL50s ~ LC50 Effect: MOR Mortality. Effect expressed as percent death or percent survival.

ENZ Enzyme effect. Change in enzyme activity.

BIO Change in physiochemical process, including glycogen uptake, cholesterol levels, and lipid analysis.

GRO Growth. Measurable change in length and/ or weight.

ABN Physical malformation during development stages.

DVP Change in the ability to grow to a more mature life stage and in time between separate life stages.

HAT Hatchability. Change in percent hatch or time to hatch.

Water Type:

SW Saltwater FW Freshwater Exposure Type:

S Static NR Not reported F Flow through I Injection R Renewal D Diet or oral exposure or both simultaneously Source: ACQUIRE Database ( retrieved October 1999). Page 47 of 65 Appendix 4. Aquatic Toxicity Data for Chlorinated Dibenzofurans ( CDFs) ( U. S. EPA, 1990)

Chemical Form Test Species Duration ( days) Concentration ( ng/ l) Toxic Effects Reference Exposure Observation Monochloro No Data Dichloro No Data Trichloro No Data Tetrachloro

1,3,7,8 TCDF Rainbow Trout ( yolk sac fry) 4 ( water) + 10 Days 5.7 x 10 4 times as toxic as 2,3,7,8 TCDD Helder and Seinen ( 1986)

2,3,7,8 TCDF Rainbow Trout ( yolk sac fry) 4 ( water) + 10 Days 9.5 x 10 5 times as toxic as 2,3,7,8 TCDD 2,3,7,8 TCDF Rainbow Trout ( fry  0.38 g) 28 ( water) + 28 Days 8.78 3.93 1.79 0.90 0.41 Controls 28% dead ( 28 days)

46% dead ( 56 days)

18% dead ( 28 days)

22% dead ( 56 days)

3% dead ( 28 days)

3% dead ( 56 days)

6% dead ( 28 days)

6% dead ( 56 days)

2% dead ( 28 days)

2% dead ( 56 days)

0% dead ( 56 days) Mehrle et al. ( 1988)

Rainbow Trout ( same effects as same conc. as TCDD) Lethal edema Helder ( 1980)

Rainbow Trout ( yolk sac fry) 4 + 10 Days 8.9 x 10 2 times as toxic as 2,3,7,8 TCDD Helder and Seinen ( 1986)

Source: U. S. EPA, 1990. Page 48 of 65 Appendix 5. Summary of Aquatic Toxicity Data for CDFs from the ACQUIRE Database Test Loc CAS #/ Chemical Latin name, Common name Endpoint Effect Water Type Dur ( days)/

Exp Typ Conc ( ug/ L) Ref.

LAB 132649, Dibenzofuran Cyprinodon variegatus, Sheepshead minnow LC50 MOR SW 1.00/ S F > 3200 Heitmuller et al., 1981 LC50 MOR SW 2.00/ S F > 3200 LC50 MOR SW 3.00/ S F 3100 LC50 MOR SW 4.00/ S F 1800 NOEC MOR SW 4.00/ S F 1000 Daphnia magna, Water flea LC50 MOR FW 1.00/ S F 7500 LeBlanc, 1980 LC50 MOR FW 2.00/ S F 1700 LC50 MOR FW 2.00/ S F 1340 Maas, 1990 LC50 MOR FW 2.00/ S F 12000 Mysidopsis bahia, Opossum shrimp LC50 MOR SW 4.00/ NR F 1310 EPA, 1978 Pimephales promelas, Fathead minnow LC50 MOR FW 4.00/ F F 1780 Geiger et al., 1988 LC50 MOR FW 4.00/ F F 1850 LC50 MOR FW 4.00/ F F 1050 Brooke, 1991 LC50 MOR FW 4.00/ S F 3620 LC50 MOR FW 4.00/ S F 1140 LC50 MOR FW 4.00/ S F 3020 Poecilia reticulata, Guppy LC50 MOR FW 4.00/ S F 1800 Maas, 1990 LC50 MOR FW 4.00/ S F 18000 Skeletonema costatum, Diatom EC50 PSE SW 4.00/ NR F1500 EPA, 1978 LAB 39001020, 1,2,3,4,5,6,7,8 Octachlorodibenzofuran Brachydanio rerio, Zebra danio, zebrafish ET50 HAT FW 3.43/ F A 0.032 Berends et al., 1997 ET50 HAT FW 3.54/ F A 0.036 LAB 51207319, 2,3,7,8 Tetrachlorodibenzofuran Cyprinus carpio, Common, mirror, colored, carp ED50 ENZ FW 14.00/ I F 17.75 mmol/ kg Van der Weiden et al.,

1994b ED50 ENZ FW 14.00/ I F 18.63 mmol/ kg LOEC ENZ FW 14.00/ I F 10.51 ug/ kg Oncorhynchus mykiss, Rainbow trout, donaldson trout ED50 ENZ a a F 0.57 umol/ kg Metcalfe & Niimi, 1993 LD50 MOR FW 7.00/ I F 8.086 Walker & Peterson., 1991 LOEC ENZ FW 2.00  16.00/ D A 0.12 ug/ kg Parrot et al., 1995 NOEC GRO FW 56.00/ F F 0.00041 Mehrle et al., 1988 NOEC MOR FW 56.00/ F F 0.00179 Oryzias latipes, Medaka, high eyes EC50 ABN FW > 13.00/ S F 0.007 Wisk & Cooper, 1990a EC50 ABN FW > 13.00/ S F 0.019 LC50 MOR FW > 13.00/ S F 0.016 Appendix 5. Summary of Aquatic Toxicity Data for CDFs from the ACQUIRE Database ( continued)

Page 49 of 65 a Not reported.

Endpoint: LC50 Median lethal concentration. Used only when death is observed endpoint. TLms and TL50s ~ LC50 EC50 Median effective concentration. Used when an effect other than death is the observed endpoint.

NOEC No observed effect concentration ED50 Effective dose LD50 Lethal dose LOEC Lowest observed effect concentration Effect: MOR Mortality. Effect expressed as percent death or percent survival.

ENZ Enzyme effect. Change in enzyme activity.

BIO Change in physiochemical process, including glycogen uptake, cholesterol levels, and lipid analysis.

GRO Growth. Measurable change in length and/ or weight.

ABN Physical malformation during development stages.

Water Type:

SW Saltwater FW Freshwater Exposure Type:

S Static NR Not reported F Flow through I Injection R Renewal D Diet or oral exposure or both simultaneously Source: ACQUIRE Database ( retrieved October 1999). Page 50 of 65 Appendix 6. Summary of Studies on the Toxicity of 2,3,7,8 TCDD to Birds ( U. S. EPA, 1990)

Species Exposure Level Symptoms/ Effects Reference Chickens 1  g/ kg/ day oral dose Positive chick edema ( lesions) Schwetz et al. ( 1973)

Eastern Bluebird 10,000 ppt LOAEL for eggs Thiel et al. ( 1988)

Chickens 10 10 M concentration of TCDD 60% reduction in tymphoid cell numbers Nikolaidis et al. ( 1988)

White Leghorn Chickens 30 65 ppt LOAEL for eggs Sullivan et al. ( 1987)

Northern Bobwhite 15  g/ kg body weight single oral dose LD50 Eisler ( 1986)

Ringed Turtle Dove 810  g/ kg body weight single oral dose LD50 Eisler ( 1986)

Mallards 108  g/ kg body weight single oral dose LD50 Eisler ( 1986)

Bobwhite 167 ppt oral exposure in a feeding test for 5 days with total observation period of 8 days total LC50 Kenaga and Norris ( 1983)

3 ppt 18 wk/ 18 wk test 0.3 ppt 18 wk/ 18 wk test No effect on reproduction Mallard 278 ppt oral exposure in a feeding test for 5 days with total observation period of 8 days LC10 Kenaga and Norris ( 1983)

Turkey > 259 ppt 11 day/ 11 day test " EC0" ( no effect level) Kenaga and Norris ( 1983)

Chicken > 500 ng/ kg/ day 21 day/ 21 day test > 200 ng/ kg/ day 21 day/ 21 day test > 100 ng/ kg/ day 21 day/ 21 day test LC90 Reduced feeding and growth " LC0" ( no mortality)

" LC0" ( no mortality) Kenaga and Norris ( 1983)

Source: U. S. EPA, 1990. Page 51 of 65 Appendix 7. Summary of the Toxic Effects of 2,3,7,8 TCDD to Birds ( U. S. EPA, 1993)

Test Species Test Method Water Conc.

( ng/ L) a Organism Conc.

( pg/ g) b Duration Effect Reference Exposure Observation Bobwhite quail,

Colinus virginianus Oral dose 15,000c Single dose 37 d LD50 Hudson et al., 1984 Mallard, Anus platyrynchos Oral dose > 108,000c Single dose 37 d LD50 Hudson et al., 1984 Ringed turtle dove,

Stretopelia risoria Oral dose > 810,000c Single dose 37 d LD50 Hudson et al., 1984 Ring necked pheasant, Phasianus colchicus Eggs Egg injection,

yolk 2,100c Single dose 28 d post hatch LD50 Nosek et al., 1992c 10,000c Single dose 28 d post hatch LOAELd mortality Nosek et al., 1992c 1,000c Single dose 28 d post hatch NOAELe Nosek et al., 1992c Egg injection,

albumin 1,400c Single dose 18 d post hatch LD50 Nosek et al., 1992c 1,000c Single dose 28 d post hatch LOAEL mortality Nosek et al., 1992c 100c Single dose 28 d post hatch NOAEL Nosek et al., 1992c Adult hen i. p. injection 100,000c Single dose 77 d 100% mortality after 42

d Nosek et al., 1992a 25,000c Single dose 77 d 80% mortality Nosek et al., 1992a 6,250c Single dose 77 d 0% mortality Nosek et al., 1992a 1,000c Weekly injections for 10 wk 7 wk post dose LOAEL hen mortality hen weight egg production embryo mortality Nosek et al., 1992a 100c Weekly injections for 10 wk 7 wk post dose NOAEL Nosek et al., 1992a a Measured TCDD concentration in water.

b Measured TCDD concentration in organism ( wet weight).

c Unmeasured TCDD concentration in water or organism ( wet weight).

d LOAEL= Lowest observed adverse effect level.

e NOAEL = No observed adverse effect level. Page 52 of 65 Appendix 8. Summary of the Toxic Effects of 2,3,7,8 TCDD to Wildlife Mammals ( U. S. EPA, 1990)

Test Species Test Method Water Conc.

( ng/ L) a Organism Conc.

( pg/ g) b Duration Effect Reference Exposure Observation Mink, Mustela vison Newborn i. p. injection 1,000c Daily for 12 d 133 d 100% mortality after 14

d Aulerich et al., 1988 1,000c Daily for 12 d 133 d 62% mortality after 133

d Aulerich et al., 1988 Adult Oral dose 4,200c Single dose 28 d LD50 Hochstein et al., 1988 a Measured TCDD concentration in water.

b Measured TCDD concentration in organism ( wet weight).

c Unmeasured TCDD concentration in water or organism ( wet weight).

Appendix 9. Summary of Studies on the Toxicity of 2,3,7,8 TCDD to Wildlife Mammals ( U. S. EPA, 1993)

Species Exposure Level Symptoms/ Effects Reference Rhesus Monkeys .0017  g/ kg body weight daily ( 7 29 months) Abortion and weight loss Eisler ( 1986)

Monkeys .011  g/ kg/ day ( 9.3 months) Generalized toxicity: anemia, hair loss,

death. 1 of 2 pregnancies aborted. Kociba and Schwetz ( 1982)

.0017  g/ kg/ day ( 7 20 months) Slight loss of weight and hair. 4 of 7 pregnancies aborted.

.24  g/ kg 3 times weekly for 3 weeks of early gestation Severe toxicity ( death). 2/ 2 pregnancies aborted. Kociba and Schwetz ( 1982)

.048  g/ kg 3 times weekly for 3 weeks of early gestation Slight toxicity. 3/ 4 pregnancies aborted. Species Exposure Level Symptoms/ Effects Reference Page 53 of 65 .0095  g/ kg 3 times weekly for 3 weeks of early gestation No toxicity. 1/ 4 pregnancies aborted.

Control 3/ 11 pregnancies aborted. Page 54 of 65 Appendix 10. Summary of 2,3,7,8 TCDD Toxicological Endpoints for Fresh Water Fishes Test Species Growth Stages Exposure Toxic Effect Toxic. Endpoints References Route Duration Observed LC50 ( ng/ L in water) ( ppt)

Fathead Minnow  water 28 days 28+ 20 days mor. 1.7 Adams et al. ( 1986)

Japanese Medaka eggs  fertilized egg to 3 d post hatch  mor. 9.0 13.0 Wisk and Copper 1990 a, b Rainbow Trout Swim up fry ( 0.38 g) water ( flow through) 28 days 28 days mor. 0.046 Mehrl et al. ( 1988)

Sac frey water ( renewal) 96 hours 21 days mor. 1.83 Bolet L., ( 1989)

EC50 ( ng/ L in water ) ( ppt)

Japanese Medaka eggs  fertilized egg to 3 d post hatch  embryos with lesions 3.5 6.0 Wisk and Copper 1990 a, b LOEC ( ng/ L in water) ( ppt)

Rainbow Trout Swim up fry ( 0.38 g) water ( flow through) 28 days 28 days mor. 0.038 Mehrle et al. ( 1988)

NOEC ( ng/ L in water) ( ppt)

Rainbow Trout  water ( flow through) 56 days  gro/

mor < 0.038 Mehrle et al. ( 1988)

Swim up fry ( 0.38 g) water ( flow through) 28 days 28 days mor. 0.0011 Mehrle et al. ( 1988)

LD50 ( µ g/ kg for dose and ng/ g for conc. in fish)

Blue Gill  i. p. 1 dose 80 days mor. 16 µ g/ kg Kleeman et al. ( 1988)

Bullhead  i. p. 1 dose 80 days mor. 5 µ g/ kg Kleeman et al. ( 1988)

Carp  i. p. 1 dose 80 days mor. 3 µ g/ kg Kleeman et al. ( 1988)

Large Mouth Bass  i. p. 1 dose 80 days mor. 11 µ g/ kg Kleeman et al. ( 1988)

Rainbow Trout  i. p. 1 dose 80 days mor. 10 µ g/ kg Spitsbergen et al. ( 1988a)

 i. p. 1 dose 80 days mor. 3 µ g/ kg Kleeman et al. ( 1988)

Yellow Perch  i. p. 1 dose 80 days mor. 10 µ g/ kg Kleeman et al. ( 1988b Yellow Perch  i. p. 1 dose 80 days mor. 3 µ g/ kg Spitsbergen et al. ( 1988b) Appendix 10. Summary of 2,3,7,8 TCDD Toxicological Endpoints for Fresh Water Fishes ( continued)

Test Species Growth Stages Exposure Toxic Effect Toxic. Endpoints References Route Duration Observed Page 55 of 65 Rainbow Trout  injec. 1 dose 30 days mor. 0.171 ng/ g ( in fish) Zabel & Peterson ( 1996)

 injec. 1 dose 7 days mor. 0.230 ng/ g ( in fish) Walter & Peterson ( 1991)

Finger ling i. p. 1 dose 80 days mor. 10 ng/ g ( in fish) Spitsbergen et al. 1988a Kleeman et al. ( 1988)

Eggs water ( renewal) 48 hours > 48 hrs to post swimup mor.

( sac fry) 0.439 ng/ g ( in eggs) Walker et al. ( 1992)

ED50 ( µ g/ kg for dose and ng/ g for conc. in fish)

Carp  injec. 1 dose 7 days enz 0.048 µ g/ kg Van der Weiden et al. 1994a Rainbow Trout  injec. 1 dose 7 days bio 0.17 µ g/ kg Newsted & Geiger, 1993  injec. 1 dose 7 days enz 0.66 µ g/ kg Newsted et al. 1995 Juvenile ( 46 g) i. p. 1 dose 3 weeks EROD Induction 0.79 ng/ g ( in fish) Van der Wekdeb et al. 1992 immature adult ( 350 g) i. p. 1 dose 72 h AHH Induction 0.64 ng/ g ( in fish) Janz and Metcalfe, 1991 LOAEL ( µ g/ kg in fish)

Carp  injec. 1 dose 7 days enz 0.3 µ g/ kg ( in fish) Van der Weiden et al. 1994a  injec. 1 dose 14 days enz 0.03 µ g/ k ( in fish) Van der Weiden et al. 1994a Rainbow Trout  injec. 1 dose 7 days enz 0.01 µ g/ kg ( in fish) Newsted and Giesy, 1993  injec. 1 dose 14 days enz 0.1 µ g/ kg ( in fish) Newsted et al., 1995  diet 2 16 days  enz 0.072 µ g/ kg ( in fish) Parrot et al., 1995 Appendix 10. Summary of 2,3,7,8 TCDD Toxicological Endpoints for Fresh Water Fishes ( continued)

Test Species Growth Stages Exposure Toxic Effect Toxic. Endpoints References Route Duration Observed Page 56 of 65 Lake Trout eggs water ( renewal 48 hours > 48 hours to post swimup mor.

( Sac fry) 0.055 ng/ g ( in egg) Walker et al., 1991 NOAEL ( µ g/ kg in fish)

Carp  injec. 1 dose 7 days enz 0.01 µ g/ kg ( in fish) Van der Weiden et al. 1994a Rainbow Trout  injec. 1 dose 7 days enz 0.01 µ g/ kg ( in fish) Newsted and Giesy, 1993 Lake Trout eggs water renewal 48 hours > 48 hours to post swimup  0.034 ng/ g ( in egg) Walker et al., 1991 Exposure Routes ( Test Methods):

i. p. Intraperitoneal injection injec. injection water water water ( static) static water water ( flow through) flow through water water ( renewal) water renewal diet fish diet Toxic Effects:

mor. mortality enz effects on Enzyme Activity bio biological effects gro growth retardation phy physical deformation Page 57 of 65 Appendix 11. Summary of 2,3,7,8 TCDD Toxicological Endpoints for Marine/ Estuarine Fish Test Species Growth Stages Exposure Toxic Effect Toxic. Endpoints References Route Duratio n Observed ED50 European flounder  diet 10 days  enz 1.62 µ g/ kg Besselink et al. ( 1997)

Mummichog  water ( static) 27 days  phy 2.025 ng/ g Prince and Cooper ( 1995)

Exposure Routes ( Test Methods): water ( static) static water diet fish diet Toxic Effects: enz effects on enzyme activity phy physical deformation Page 58 of 65 Appendix 12. Summary of 2,3,7,8 TCDD Toxicity Data for Fresh Water Invertebrates Test Species Growth Stages Exposure Toxic Effect Test Conc. ( ng/ L) References Route Duration Observed ( Not a NOEC/ LOEC value)

Annelid, Worm  water ( static) 55 days  no effect on repro. success 200 ng/ L ( water) Miller et al. ( 1973)

Mollusc, Snail adult Model Ecosy. 33 days  No toxic effect 1,330 ng/ L ( water) 502 ng/ g ( organism) Isensee and Jones ( 1975)

adult water ( static) 36 days 12 days no effect on repro. success 200 ng/ L ( water) Miller et al. ( 1973)

Arthropod, Mosquito  water ( static) 17 days 23 days no effect 200 ng/ L ( water) Miller et al. ( 1973)

Arthropod, Cladoceran  Model Ecosy. 33 days  no effect 1,330 ng/ L ( water) 1,570 ng/ g ( organism) Isensee and Jones ( 1975)

Arthropod, Cladoceran 1 21 days water ( renewal) 48 hours 7 days no effect 1,030 ng/ L ( water) Miller et al. ( 1979) Page 59 of 65 Appendix 13. Summary of Acute and Short Term 2,3,7,8 TCDDToxicological Endpoints for Birds Exposure LD50 ( ng/ kg) Reference Route Duration Observation Northern bobwhite quail oral single dose 37 days 15 ( 95% confidence limits: 9.2 24.5) Hudson et al. ( 1984)

Bobwhite oral 5 days 7 days LC50 = 0.167 ( in feed) Kenaga and Norris ( 1983)

Ringed turtle dove oral single dose 37 days > 810 Hudson et al. ( 1984)

Mallard oral single dose 37 days > 108 Hudson et al. ( 1984)

Ring necked pheasant i. p. injection single dose 11 weeks 25 ( LD75) Nosek et al. ( 1992)

Appendix 14. Summary of Subchronic and Chronic 2,3,7,8 TCDD Toxicological Endpoints for Birds Test Species Exposure LOAEL ( µ g/ kg day) NOAEL ( µ g/ kg day) Toxic Effects References Route Duration Observed Domestic Chicken oral daily doses for 20 to 21 days 20  21 days 1.0 0.10 mortality Schwetz et al. ( 1973)

Pheasant i. p. injec. 10 weeks  0.14 0.014 fertility, embryo mortality Nosek et al. ( 1992b) and( 1993 Page 60 of 65 Appendix 15. Summary of Acute and Short Term 2,3,7,8 TCDD Toxicological Endpoints for Mammals Growth Stages Exposure LD50 ( µ g/ kg) Reference Route Duration Observation Dog  oral single dose  100  200 Kociba and Schwetz ( 1982)

Guinea pig  oral single dose  0.6  2.1 Schwetz et al. ( 1973)

Hamster  oral single dose  1,160  5,050 Kociba and Schwetz ( 1982)

Mink male adults oral single dose 28 days 4.2 Hochstein et al. ( 1988)

kits injection single dose 12 days < 0.1 Aulerich et al. ( 1988)

Mouse  oral single dose  114  284 Kociba and Schwetz ( 1982)

Rabbit  oral single dose  115 Schwetz et al. ( 1973)

Rat  oral single dose  22  45 Schwetz et al. ( 1973)

Rhesus monkey  oral single dose  70 Kociba and Schwetz ( 1982)

Appendix 16. Summary of Subchronic and Chronic 2,3,7,8 TCDDToxicological Endpoints for Mammals Exposure Duration Toxic Effects LOAEL ( µ g/ kg day) NOAEL ( µ g/ kg day) References Rat 2 years female mortality 0.1 0.01 Kociba et al. ( 1978)

Rat gestation days 6 to 15 litter size pup weight 0.25 0.125 Khera and Ruddick ( 1973)

Rat 3 generations reproductive capacity 0.01 0.001 Murray et al. ( 1979)

Rhesus Monkey 7 months number of births 0.0021  Allen et al. ( 1979)

Rhesus Monkey 7  48 months maternal reproductive 0.00059 0.00012 Bowman et al. ( 1989) Page 61 of 65 APPENDIX 17: Terrestrial Exposure Assessment The meadow vole ( Microtus pennsylvanicus) was selected as the mammalian receptor species for evaluating potential effects of CDD/ CDFs 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 CDD/ CDFs to birds. 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 vole 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, 1993b) and 0.093 kg/ kg bwd for adult bobwhites ( U. S. EPA, 1993b) 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 ( 1993b) 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 CDD/ CDFs. 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. Page 62 of 65 The concentration of a chemical in a terrestrial plant ( C p) as fresh weight was determined using the following equation:

C p = C soil * Kps ( 3)

where:

C soil = concentration ofcontaminant of concern detected in surface soil ( mg/ kg); and Kps = partition coefficient in plant tissue relative to contaminant soil concentration of TCDD ( unitless) ( Chiao et al., 1994).

The predicted EEC for CDD/ CDF total TEQ in soil ( see Environmental Fate Modeling chapter of this document) was used as the C soil in the model for voles and bobwhites.

Employing equations 1, 2, and 3, the estimated dose that voles and bobwhites would receive from ingestion of plants in the vicinity of pentachlorophenol treated 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 CDD/ CDFs 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 = +

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. Page 63 of 65 In the risk assessment, the total dietary intakes are compared to toxicity endpoints ( e. g., LD50, NOAEL) to determine if adverse effects are likely to occur in voles and bobwhites from the ingestion of CDDs/ CDFs in terrestrial plants and surface soil. The toxicity endpoints used were obtained from laboratory toxicity studies conducted using TCDD.. These endpoints were not converted to TEQ. Toxicity reference values from studies based on avian species other than 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, 1993b).

Following adjustment of the mammalian toxicity reference values, the total dose for the bobwhite and vole of CDD/ CDF TEQ was compared with the toxicity reference values for TCDD to determine the risk quotients ( RQs). The RQs for the bobwhite and vole based on exposure to CDD/ CDF TEQ are shown in Tables A1 and A2, respectively. Page 64 of 65 Table A1. Acute and Chronic Risk Quotient Calculations for Birds Compound EEC ( mg/ kg) Bobwhite Soil Ingestion Rate ( kg/ kg d) Bobwhite Soil Dose ( mg/ kg day) Plant/ soil partifion coefficitent 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 CDD/ CDF TEQ 2.32e 02 0.0093 2.16e 04 3.40e 01 7.89e 03 1 7.89e 03 0.093 7.34e 04 9.49e 04 LD50 1.5e 05 mg/ kg ( Hudson et al.,

1984) acute 63.321 NOAEL 1.4e 05 mg/ kg d ( Scwetz et al.,

1973) chronic 67.782 TCDD 2.38e 04 0.0093 2.20e 06 3.40e 01 8.09e 05 1 8.09e 05 0.093 7.50e 06 9.70e 06 LD50 1.5e 05 mg/ kg ( Hudson et al.,

1984) acute 0.653 NOAEL 1.4e 05 mg/ kg d ( Scwetz et al.,

1973) chronic 0.69 Page 65 of 65 1. Exceeds acute High Risk Level of Concern ( LOC)

2. Exceeds Chronic Risk LOC 3. Exceeds acute Restricted Use LOC Table A2. Acute and Chronic Risk Quotient Calculations for Mammals Compound EEC ( mg/ kg) Vole Soil Ingestion Rate ( kg/ kg d) Vole Soil Dose ( mg/ kg day) Plant/ soil partition coefficient 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) Endpoint ( mg/ kg) RQ CDD/ CDF TEQ 2.32e 02 0.035 8.12e 04 3.40e 01 7.89e 03 1 7.89e 03 0.35 2.76e 03 3.57e 03 LD50*

0.09 mg/ kg ( Kociba and Schwetz, 1982) acute 0.04 NOAEL*

9.94e 04 mg. kg day ( Khera and Ruddick, 1973) chronic 3.592 TCDD 2.38e 04 0.035 8.30e 06 3.40e 01 8.09e 05 1 8.09e 05 0.35 2.83e 05 3.66e 05 LD50*

0.114 mg/ kg ( Kociba and Scwetz, 1982) acute 4.07e 04 NOAEL*

9.94e 04 mg. kg day ( Khera and Ruddick, 1973 chronic 0.04 * Value for mouse ( LD50 of 0.114 mg/ kg day) and rat ( NOAEL of 0.125 mg/ kg day) were adjusted for body weight as described in Sample et al.( 1996) and U. S.

EPA ( 1993b)