[ Chapter 7 Part A discusses Cancer Effects]
EPA reassessment index page: http://www.epa.gov/ncea/pdfs/dioxin/part1and2.htm
Chloracne is one of the best known of the medical consequences of exposure to 2,3,7,8-TCDD-contaminated substances. In general, it has been observed in most incidents where substantial exposure has occurred, particularly among TCP production workers (Goldman, 1972; May, 1973; Bleiberg et al., 1964; Bond et al., 1987; Suskind and Hertzberg, 1984; Moses et al., 1984; Zober et al., 1990) and Seveso residents (Reggiani, 1978; Caramaschi et. al., 1981; Ideo et al., 1985; Mocarelli et al., 1986; Assennato et al., 1989). As previously stated, chloracne appears within several weeks to months from the time of exposure, often resolving after discontinuation of exposure (Moses et al., 1984; Suskind and Hertzberg, 1984), although for some it may remain for extended periods after exposure ended (Moses et al., 1984).
The amount of exposure necessary for development of chloracne has not been resolved, but studies suggest that high exposure (both high acute and long-term exposure) to 2,3,7,8-TCDD increases the likelihood of chloracne, as evidenced by chloracne in TCP production workers and Seveso residents who have documented high serum 2,3,7,8-TCDD levels (Beck et al., 1989; Fingerhut et al., 1991a; Mocarelli et al., 1991; Neuberger et al., 1991) or in individuals who have a work history with long duration of exposure to 2,3,7,8-TCDD-contaminated chemicals (Bond et al., 1989). The absence of substantial chloracne in U.S. Army Vietnam veterans whose mean serum 2,3,7,8-TCDD levels were at background (4 pg/g) (Centers for Disease Control Vietnam Experience Study, 1988d) and U.S. Air Force Ranch Hands whose serum 2,3,7,8-TCDD levels fell intermediate to those of workers and Army Vietnam veterans (Roegner et al., 1991; Burton et al., 1997) suggests that there is a higher incidence of the disorder among those with higher serum 2,3,7,8-TCDD levels.
In earlier studies, chloracne was considered to be a "hallmark of dioxin intoxication" (Suskind, 1985). However, only in two studies were risk estimates calculated for chloracne. Both were studies of different cohorts of TCP production workers (Suskind and Hertzberg, 1984; Bond et al., 1989); one group was employed in a West Virginia plant, the other in a plant in Michigan. Of the 203 West Virginia workers, 52.7% (p<0.001) were found to have clinical evidence of chloracne, and 86.3% reported a history of chloracne (p<0.001) (Suskind and Hertzberg, 1984). None of the unexposed workers had clinical evidence or reported a history of chloracne. Among the Michigan workers, the relative risk for cases of chloracne was highest for individuals with the longest duration of exposure ($60 months; RR = 3.5, 95% CI = 2.3-5.1), those with the highest cumulative dose of TCDD (based on duration of assignment across and within 2,3,7,8-TCDD-contaminated areas in the plant) (RR = 8.0, 95% CI = 4.2-15.3), and those with the highest intensity of 2,3,7,8-TCDD exposure (RR = 71.5, 95% CI=32.1-159.2) (Bond et al., 1989).
Chloracne is associated with exposure to other polyhalogenated chemicals, including dibenzofurans, PCBs, naphthalenes, and others (Taylor, 1979). The likelihood of exposure to other polyhalogenated chemicals in the populations studied.
is probably low, particularly among the Seveso children, whose exposure was to TCP reactant effluents that were primarily contaminated with 2,3,7,8-TCDD. The issue is more relevant in chemical workers, who by virtue of their occupation, have the potential for exposure to other chemicals. Yet, much of the documented chloracne appeared shortly after TCP reactor releases (Ashe and Suskind, 1950; Goldman, 1972; May, 1973) or during TCP or 2,4,5-T production (Bond et al., 1989), suggesting that 2,3,7,8-TCDD was the chloranegenic agent.
Animal studies have been effective in describing the relationship between 2,3,7,8-TCDD and chloracne, particularly in rhesus monkeys (McNulty, 1977; Allen et al., 1977; McConnell et al., 1978). Subsequent to exposure to 2,3,7,8-TCDD, monkeys developed chloracne and swelling of the meibomian gland, a modified sebaceous gland. The histologic changes in the meibomian gland are physiologically similar to those observed in human chloracne (Dunagin, 1984). In summary, the evidence provided by the various studies convincingly states what is already presumed, that chloracne is a common sequela of high levels of exposure to 2,3,7,8- TCDD. More information is needed to determine the level and frequency of 2,3,7,8-TCDD exposure needed to cause chloracne and whether personal susceptibility plays a role in the etiology. Finally, it is important to recall that the absence of chloracne does not imply lack of exposure (Mocarelli et al., 1991).
There appears to be a consistent pattern of increased GGT levels among individuals exposed to 2,3,7,8-TCDD-contaminated chemicals. Elevated levels of serum GGT have been observed within a year after exposure in Seveso children (Caramaschi et al., 1981; Mocarelli et al., 1986) and 10 or more years after cessation of exposure among TCP and 2,4,5-T production workers (May, 1982; Martin, 1984; Moses et al., 1984; Calvert et al., 1992) and among Ranch Hands (Roegner et al., 1991; Grubbs et al., 1995). All of these groups had a high likelihood of substantial exposure to 2,3,7,8-TCDD. In addition, for those studies that evaluated dose-response relationships with 2,3,7,8-TCDD levels, the effect was observed only at the highest levels or categories of 2,3,7,8- TCDD. In contrast, although background levels of serum 2,3,7,8-TCDD suggested minimal exposure to Army Vietnam veterans, GGT was increased, at borderline significance, among Vietnam veterans compared to non-Vietnam veterans (Centers for Disease Control Vietnam Experience Study, 1988a). In addition, despite the increases observed in some occupational cohorts, other studies of TCP production workers from West Virginia or Missouri residents measured but did not report elevations in GGT levels (Suskind and Hertzberg, 1984; Webb et al., 1989).
In clinical practice, GGT is often measured because it is elevated in almost all hepatobiliary diseases and is used as a marker for alcoholic intake (Guzelian, 1985). In individuals with hepatobiliary disease, elevations in GGT are usually accompanied by increases in other hepatic enzymes, e.g., AST and ALT, and metabolites, e.g., uro- and coproporphyrins. Significant increases in hepatic enzymes other than GGT and metabolic products were not observed in individuals whose GGT levels were elevated 10 or more years after exposure ended, suggesting that the effect may be GGT-specific. However, in the Seveso male children and those with chloracne (both sexes), ALT was significantly increased concomitantly with GGT within 1 year of the reactor release (Mocarelli et al., 1986; Caramaschi et al., 1981). In a longitudinal analysis, both enzymes returned to normal levels within 5 years after exposure (Mocarelli et al., 1986; Assennato et al., 1989). In the NIOSH cohort, elevated GGT levels occurred only among workers with both high 2,3,7,8-TCDD concentrations and lifetime alcohol consumption. Yet, no other enzymes were found to be outside of normal ranges. Likewise, GGT was the single enzyme found to be significantly elevated in Ranch Hands compared to nonexposed referents (Roegner et al., 1991; Grubbs et al., 1995). These data suggest that in the absence of increases in other hepatic enzymes, elevations in GGT are associated with exposure to 2,3,7,8-TCDD, particularly among individuals who were exposed to high 2,3,7,8-TCDD levels. The fact that investigators observed a decline in enzyme levels in Seveso children but a continued elevation in TCP workers may reflect (1) differences in how exposure occurred (i.e., acute but high doses in Seveso versus continuous or frequent long-term, medium to high doses in TCP workers), (2) differences in the metabolism of the maturing versus mature system, (3) the fact that children grow rapidly, thus "diluting" a peak exposure within that group, or (4) some combination of these.
The animal data with respect to 2,3,7,8-TCDD-related effects on GGT are sparse. Statistically significant changes in hepatic enzyme levels, particularly AST, ALT, and ALK, have been observed after exposure to 2,3,7,8-TCDD in rats and hamsters (Gasiewicz et al., 1980; Kociba et al., 1978; Olson et al., 1980). Only one study evaluated GGT levels (Kociba et al., 1978). Moderate but statistically nonsignificant increases were noted in rats fed 0.10 µg/kg 2,3,7,8-TCDD daily for 2 years, and no increases were observed in control animals.
Among humans, increased levels of GGT may suggest activity such as cholestases, liver regeneration, or drug or xenobiotic metabolism. In human adults, most of 2,3,7,8-TCDD is stored in the adipose tissue and has a half-life of approximately 7 years (Pirkle et al., 1989). Continued GGT activity in adults with serum 2,3,7,8-TCDD levels many times over background levels may reflect continuous, low-level metabolism of 2,3,7,8-TCDD. In summary, GGT is the only hepatic enzyme examined that was found in a number of studies to be chronically elevated in adults exposed to high levels of 2,3,7,8-TCDD. The consistency of the findings in a number of studies suggests that the finding may reflect a true effect of exposure but for which the clinical significance is unclear. Long-term pathologic consequences of elevated GGT have not been illustrated by excess mortality from liver disorders or cancer, or in excess morbidity in the available cross-sectional studies. It must be recognized that the absence of an effect in a cross-sectional study, for example, liver enzymes, does not obviate the possibility that the enzyme levels may have increased concurrent to the exposure but declined after cessation. The apparently transient elevations in ALT levels among the Seveso children suggest that hepatic enzyme levels other than GGT may react in this manner to 2,3,7,8-TCDD exposure.
Levels of reproductive hormones have been measured with respect to exposure to 2,3,7,8-TCDD in three cross-sectional medical studies. Testosterone, LH, and FSH were measured in TCP and 2,4,5-T production workers (Egeland et al., 1994), in Army Vietnam veterans (Centers for Disease Control Vietnam Experience Study, 1988d), and in Ranch Hands (Thomas et al., 1990; Grubbs et al., 1995). The risk of abnormally low testosterone was two to four times higher in exposed workers with serum 2,3,7,8-TCDD levels above 20 pg/g than in unexposed referents (Egeland et al., 1994). In both the 1987 and 1992 examinations, mean testosterone concentrations were slightly, but not significantly higher in Ranch Hands (Roegner et al., 1991; Grubbs et al., 1995). FSH and LH concentrations were no different between the exposed and comparison groups. No significant associations were found between Vietnam experience and altered reproductive hormone levels (Centers for Disease Control Vietnam Experience Study, 1988d). Only the NIOSH study found an association between serum 2,3,7,8-TCDD level and increases in serum LH.
The NIOSH study excluded from analysis participants who had conditions that might have influenced gonadotropin and/or testosterone levels: history of prostate cancer, thyroid or other hormone usage, or liver cirrhosis. Similarly, in the Ranch Hand study, individuals with orchiectomies or who were taking testosterone medication were excluded from the analysis of testosterone; no participants were excluded from the analyses of FSH. The CDC study of Vietnam veterans did not describe the exclusions.
In rats, 2,3,7,8-TCDD has been shown to decrease testosterone levels (Moore et al., 1985; Moore and Peterson, 1988; Mebus et al., 1987) through a decrease in testosterone synthesis (Kleeman et al., 1990) or by decreasing the production of pregnenolone from cholesterol (Ruangwies et al., 1991). In addition, 2,3,7,8-TCDD has been shown in rats to reduce the responsiveness of the pituitary to testosterone (Bookstaff et al., 1990a) and of the Leydig cells to LH stimulation (Moore et al., 1991). The findings of the NIOSH and Ranch Hand studies are plausible given the pharmacological and toxicologicalal properties of 2,3,7,8-TCDD. A mechanism responsible for the effects may involve the ability of 2,3,7,8-TCDD to influence hormone receptors. The aryl hydrocarbon (Ah) receptor to which 2,3,7,8-TCDD binds can cross-talk with steroid hormone receptors in both structure and mode of action. Studies suggest that 2,3,7,8-TCDD modulates hormone receptors, including estrogens (Romkes and Safe, 1988; Romkes et al., 1987), prolactin, and its own Ah receptor (Poland and Glover, 1980; Morrow et al., 1986). However, the effect of 2,3,7,8-TCDD on testosterone receptors has not been evaluated. In summary, the results from both the NIOSH and Ranch Hand studies are limited by the cross-sectional nature of the data and the type of clinical assessments conducted. However, the available data provide evidence that alterations in human male reproductive hormone levels are associated with serum 2,3,7,8-TCDD.
The following section describes outcomes that may be related to 2,3,7,8-TCDD exposure. Further research would assist in the final evaluation of the effects of dioxin for the following outcomes.
Animal studies indicate that 2,3,7,8-TCDD is associated with generally increased serum cholesterol and serum triglyceride levels. The effect of exposure to 2,3,7,8-TCDD-contaminated chemicals on lipids is not consistent in the available epidemiology studies. Elevations in total cholesterol and triglyceride levels were reported after high 2,3,7,8- TCDD exposure in TCP workers (Pazderova-Vejlupkova et al., 1981; Martin, 1984) and.
laboratory workers (Oliver, 1975). Despite their very high exposure to 2,3,7,8-TCDD-contaminated chemicals, neither adults nor children from Seveso had lipid levels above the referent level. Risk factors such as dietary fat intake, familial hypercholesterolemia, alcohol consumption, and exercise, which also affect cholesterol and other lipid levels, may be factors that were not considered in these studies. Ranch Hands and the NIOSH cohort continue to have marginally elevated lipid concentrations despite the extended length of time between exposure and testing (Grubbs et al., 1995; Calvert et al. 1995). These most recent data suggest that high exposure to 2,3,7,8-TCDD contaminated substances are not related to significantly increased lipid concentrations, specifically total cholesterol and triglycerides. Nevertheless, slight but chronic elevations in serum lipids may put an individual at increased risk for disorders such as atherosclerosis and other conditions affecting the vascular system.
Diabetes mellitus is a heterogeneous disorder that is a consequence of alterations in the number or function of pancreatic beta cells responsible for insulin secretion and carbohydrate metabolism. Depending on its type, diabetes has been attributed to endogenous factors such as genetic predisposition, to autoimmune processes, and to exogenous factors such as viral infections (Yoon et al., 1987) and chemical exposures, notably a rat poison (Miller et al., 1978) and some medications (Wilson and LeDoux, 1989), environmental toxins (Diabetes Epidemiology Research International, 1987), age, obesity, reduced physical exercise, diet, socioeconomic status, increased insulin resistance by the beta cells, and possibly parity (Pareschi and Tomasi, 1987). The long-term risk of developing diabetes or other alterations in glucose metabolism after exposure to 2,3,7,8-TCDD is not well addressed by the available toxicological data. The results of the animal studies suggest that glucose levels are altered, generally decreased, by short-term, high-dose exposure to 2,3,7,8-TCDD, and that the response may be species-specific. Studies of rats and rhesus monkeys showed consistent decreases in serum glucose levels after daily doses administered over 30 days (Zinkl et al., 1973) or after a single dose of 2,3,7,8-TCDD (McConnell et al., 1978a; Gasiewicz et al., 1980; Schiller et al., 1986; Ebner et al., 1988) (Table 7-54). In one study, glucose levels continued to drop up to 3 weeks postexposure (Gorski et al., 1990). Dose-related decreases were also noted in CD rats fed 0.1, 1.0, or 10.0 µg/kg daily for 30 days (Zinkl et al., 1973). In contrast, lower daily doses of approximately 0.004 µg/kg/day administered to guinea pigs over a 90-day period produced no significant changes in either glucose or insulin levels (DeCaprio et al., 1986). Wasting syndrome, observed in many species after high dose exposure to 2,3,7,8-TCDD, is hypothesized to be related to various changes in the glucose transport mechanism and is mediated through the Ah receptor. Differences among the results of human and most animal studies may, hypothetically, be due to a number of factors, such as the species studied, the length of the exposure and the short observation periods of the toxicology studies, the rate of insulin metabolism after 2,3,7,8-TCDD exposure, possible differential effects of 2,3,7,8-TCDD on the various types of islet cells, and the high, usually single, 2,3,7,8-TCDD dose administered to the animals. Additionally in humans, with some exceptions, onset of diabetes occurs later in life; unfortunately, these characteristics were not evaluated in the chronic toxicity studies. Long-term feeding studies to evaluate the relationship among glucose levels, the development of diabetes, and 2,3,7,8-TCDD dose would be helpful in assessing the effect of exposure on the physiologic integrity of the islet cells. In addition, such studies may identify other factors that affect either directly or indirectly the function of the islet cells and the effect of 2,3,7,8-TCDD on the glucose transport mechanism. More discussion of glucose transport system effects and their potential relationship to diabetes is found in Part II: Chapter 3. Acute Subchronic, and Chronic Toxicity, Section 3.5. Also, a study by Matsumura and his colleagues is underway to determine whether there is a biological basis for the association between type II diabetes mellitus and dioxin levels in Ranch Hand veterans (personal communication, J. Michalek, 2000). The results of the Ranch Hand morbidity study are in direct contrast with the results of the NIOSH mortality study, which included workers who were more highly exposed, with greater frequency and severity of exposure. The prevalence of diabetes in the exposed, and referent groups was similar and there was no significant positive trend between prevalence of diabetes and serum glucose levels. What was notable is that 60% of the workers who had serum 2,3,7,8-TCDD concentrations >1,500 pg/g lipid at the time of the study met the case definition for diabetes, and that the adjusted geometric mean glucose concentration measured at the time of the study was statistically significantly higher than referent levels only in the half-life adjusted TCDD category 1,860-30,000 pg/g lipid. However, the analysis of the mortality data by exposure score, found decreasing prevalence of mortality from diabetes with increasing exposure score. The exposure score incorporated duration and intensity of exposure, among other factors. This puzzling picture suggests that in the NIOSH cohort, diabetes may not be well correlated with serum 2,3,7,8-TCDD concentration or duration of exposure. The studies of Seveso populations and the NIOSH and IARC occupational cohort found very modest and generally statistically nonsignificant increases in mortality from diabetes. The results of each of these studies should be evaluated within the context of the limitations of the studies. Although the Seveso study is limited by a short follow-up period (15 years), all of the studies have small numbers of deaths from diabetes. It is well known that diabetes has a complex etiology ranging from genetic susceptibility, viral infections, and chemical insult to physical states such as obesity or pregnancy. The various mortality studies that examined the relationship between rates of death from diabetes and 2,3,7,8-TCDD exposure evaluated no data on important confounders such as the body mass index of cases, family history of diabetes, or the relationship between year of onset of the diabetes and exposure to 2,3,7,8-TCDD or chlorophenols. In addition, although the studies have reasonably good estimates of exposure to 2,3,7,8-TCDD, misclassification of exposure for many subjects is possible. The analyses by Longnecker and Michalek (2000) provide only weak evidence to support a causal relationship between very low levels of serum 2,3,7,8-TCDD and increased risk of diabetes or changes in serum glucose or insulin. The risk measures are generally not strong or statistically significant, nor do they increase monotonically with increasing dose. The current epidemiologic and toxicological data to date do not support a strong relationship between exposure to 2,3,7,8-TCDD and diabetes or alterations in glucose metabolism. However, there is some evidence to suggest that, particularly at high doses, 2,3,7,8-TCDD may perturb glucose metabolism in some species, a fact which needs further exploration.
Given that postnatal developmental effects of dioxin have been studied only in one human population (with the exception of some of the thyroid measures), these studies are being place in the "potential" category. Additional studies in other groups are recommended, as well as followup of these findings over time to evaluate whether these changes are temporary, with no long-term health effects, or an early indication of chronic effects. All the effects in this section were part of one or both of the Dutch developmental studies. The exposures assessed here are different from the more typical "dioxin" study: the first series of studies (in Amsterdam) examined dioxins and furans, while the second (in Rotterdam and Grongenin) examined dioxins, furans, and PCBs. Thus, any effects observed could be from one agent or some mixture. Even though the studies may have picked out certain exposures as statistically significant, this does not mean that other factors not selected are not associated. For example, in the Rotterdam/Grongenin studies, only PCBs were evaluated prenatally and at birth, but these values were significantly correlated with dioxins, furans, and PCBs collected about 2 weeks after delivery. Many of the findings in Rotterdam/Grongenin were associated with in utero PCB exposures measured in maternal blood (IUPAC 118, 138, 153, 180). However, of these, only 118 is considered dioxin-like. As the technology improves to measure dioxin levels in smaller samples, direct measurements will help clarify the issues related to surrogate measures (e.g. PCBs)..
Of the various endpoints covered by the series of reports on the Dutch population, the most interesting findings related the neurobehavioral endpoints. Prechtl's Neurologic Optimality Scores (NOS) and the related postural tone cluster scores and reflex cluster scores were collected at 18 months of age (Huisman et al., 1995a) and, at 42 months of age, children were assessed for cognitive abilities using the Kaufman Assessment Battery for Children (K-ABC) and for verbal comprehension using the Reynell Language Developmental Scales (RDLS) (Patandin et al., 1999). The NOS scores were somewhat arbitrarily divided at the median and compared to the individual dioxins, furans, and PCBs, as well as their summary measures (Huisman et al., 1995a). A number of the levels of the above agents in breast milk were associated with the NOS, while the prenatal PCBs were not (Table 7-46). Coplanar PCB TEQ was associated with hypotonia (measured through the posture tone cluster score). This observation of hypotonia and prenatal PCB exposure is consistent with Rogan et al. (1986). An evaluation of motor function was associated only with the prenatal PCB levels. Because of the small volume of materal and cord blood collected, dioxins and furans were not measured during the prenatal period. An interaction observed with paternal smoking suggests that this issue should be examined further by collecting postnatal materal smoking data. Statistically significant deficits in K-ABC were associated with 3PCBmaternal blood for the entire group, and in RDLS only in the formula-fed children (Patandin et al., 1999). Importantly, the current body burdens in the 42-month-old children were not associated with any cognitive deficits. Statistically significant changes were not observed in the breast-fed children, possibly because of the higher SES status, parental education, and parental verbal IQs. Another possibility is the beneficial effect of breast feeding in general. Even though studies of Yu-Cheng children are not directly comparable to the above studies, they also showed neurobehavioral delays: increased psychomotor delays (Yu et al., 1991) and lower scores on IQ tests (Chen et al., 1992; Lai et al., 1993; Guo et al., 1993; Chen and Hsu, 1994). Many of the other outcomes in the Dutch population were "better" in those with exposure: fluency cluster score (Huisman et al., 1995b), mental development index -- MDI (Koopman-Esseboom et al., 1996), and visual recognition memory test at 7 months of age (Koopman-Esseboom et al., 1995b). This may be a result of the inherent benefit of breast feeding (and length of breast feeding) over formula for those measures, or may be due to the way women select to breast fed (e.g., higher SES women, parents with higher educational levels). As noted above, this later notion is supported in a recent report by Patandin et al. (1999).
Transplacental exposures of mice demonstrate neurobehavioral effects of dioxin and dioxin-like compounds. These include effects on postural endpoints, motor function, visual abilities, and learning. Perinatally exposed monkeys showed a deficit in cognitive function. Exposures are presented by dietary levels or dose given, and thus are difficult to compare to the exposure measures used in human studies. More details on these studies are presented in Chapter 5 (Developmental and Reproductive Toxicity). Endpoints varied in the above studies, as did the components and levels of the exposures, overall; in spite of this, the data suggest the relationship between dioxin and dioxin-like compounds and neurobehavioral outcomes. Examination of other human populations and long-term follow-up of these study groups will greatly benefit this database.
Two recent studies, both in The Netherlands but conducted in different groups, have examined thyroid function (Pluim et al., 1993; Koopman-Esseboom et al., 1994c). The two reports did have a finding in common: both observed higher TSH at 3 months of age with higher TEQs. They both had significant findings for T4, but they were in opposite directions. All these findings, plus other changes found in the second report (an increase in T4/TBG and a decrease in free T4) strongly suggest that more work be done in this area. These findings suggest a possible shift in the distribution of thyroid hormones, and point out the need for collection of longitudinal data to assess the potential for long-term effects associated with these changes.
One study looked at blood measures in 35 perinatally exposed children in The Netherlands (Pluim et al., 1994). AST, ALT, and platelets all varied with exposure (Table 7-26). Even though all but three children had values within "normal" ranges, the distributions of has shifted (e.g., an increase in platelets), which could have some currently unknown/unrecognized short- or long-term effect on health.
The following section describes endpoints for which the animal data have demonstrated exposure-related effects, but the human data are inconclusive and require further study.
In general, the results of the cohort mortality studies of TCP production workers were remarkably similar. For all of the early studies, the SMRs for diseases of the circulatory system (ICD-9: ICD 390-459) were approximately 100, meaning that the death rate in the exposed population was nearly the same as that in the general population, controlling for age, race, gender, and calendar year (Fingerhut et al., 1991b; Zober et al., 1990; Bueno de Mesquita et al., 1993; Bertazzi et al., 1989, 1992; Collins et al., 1993; Bond et al., 1989; Coggon et al., 1991). None of the SMRs above 100 were statistically significantly elevated. In the only study of its kind, Flesch-Janys and colleagues (1995) estimated exposure to PCDDs, PCDFs, and total TEQs for all members of the Hamburg chemical worker cohort based on a sample of workers with either serum or adipose tissue exposure measurements. Mortality from all cardiovascular diseases, and specifically from ischemic heart disease, was related to increasing 2,3,7,8-TCDD concentrations. More striking was the positive relationship betweeen Total TEQs and cardiovascular disease as a whole. Mortality from circulatory system diseases among Ranch Hands (SMR = 110, 95% CI = 60-150) (Michalek et al., 1990) and Australian Vietnam veterans (RR = 1.7, 95% CI = 0.9-3.0) (Fett, 1987b) was nonsignificantly elevated. In an update of the Ranch Hand data, the SMR for all circulatory diseases combined among all Ranch Hands was not elevated (SMR = 100, 95% CI = 70-130) (Michalek et al., 1998). However, a significant increase in cardiovascular mortality was noted in enlisted ground crew (SMR=150, 95% CI =1.0-2.2). There was a deficit of deaths from this cause among U.S. Army Vietnam veterans compared to non-Vietnam veterans (Centers for Disease Control Vietnam Experience Study, 1988c). Elevated mortality from circulatory diseases among Seveso residents is considered by the authors to be the result of environmental stresses and possibly other risk factors rather than exposure to 2,3,7,8-TCDD (Bertazzi et al., 1989; Pesatori et al., 1998). The mortality pattern from circulatory and cardiovascular disease among the three zones does not suggest a pattern of effect. Perhaps further followup will clarify the story. Diseases of the circulatory system, particularly heart disease, are the leading causes of death among populations of most developed nations. Leading risk factors include age, cigarette smoking, elevated lipid levels, obesity, hypertension, diabetes, and physical inactivity (Smith et al., 1992). Among the studies that examined mortality from circulatory system diseases, none directly adjusted SMRs for known risk factors or attempted to evaluate jointly the contribution of known risk factors and 2,3,7,8-TCDD to the mortality rates (Fingerhut et al., 1991b; Zober et al., 1990; Bueno de Mesquita et al., 1993; Bertazzi et al., 1989, 1992; Collins et al., 1993; Bond et al., 1989; Coggon et al., 1991). More recent mortality studies attempted to control for known confounders; when possible, most used internal control groups. Still, the picture is inconsistent among the various cohorts. Therefore, given the strong contribution of these risk factors, it is not possible to rule out physical and personal risk factors in the etiology of diseases of the circulatory system and heart in these populations. However, the absence of a "healthy worker effect" for these causes of death suggests that future research be directed specifically at the relationship between circulatory and heart disease and exposure to 2,3,7,8-TCDD. Cross-sectional morbidity studies have not found increases in the prevalence of circulatory or heart disease among TCP workers, Ranch Hands, or U.S. Army Vietnam veterans (Suskind and Hertzberg, 1984; Bond et al., 1987; Moses et al., 1984; Centers for Disease Control Vietnam Experience Study, 1988a; Calvert et al., 1998). In some cross-sectional studies, risk estimates were adjusted for some risk factors, depending on the study (Suskind and Hertzberg, 1984; Poland et al., 1971; Moses et al., 1984; Bond et al., 1983; Centers for Disease Control Vietnam Experience Study 1988a; Roegner et al., 1991). Ranch Hands were the only group to experience marginal differences in diastolic blood pressure, arrhythmias, and peripheral pulse abnormalities after adjusting for selected risk factors (Roegner et al., 1991). The animal data suggest that at high levels of 2,3,7,8-TCDD, the vascular system, cardiac muscle, and valves and function may be affected by exposure (Kociba et al., 1978; Buu-Hoļ et al., 1972; Brewster et al., 1988; Hermansky et al., 1988; Kelling et al., 1987; Canga et al., 1988). However, with the exception of the long-term feeding study (Kociba et al., 1978), the exposures in animals were single high doses and the human exposures (except Seveso) were chronic, medium to high doses. In summary, the animal studies suggest that 2,3,7,8-TCDD causes pathologic changes that may lead to later circulatory system disease. However, long-term studies of mature and aged animals have not been carried out to evaluate these hypotheses and to correlate the results of the animal with the human studies. Few epidemiologic studies were designed to control for many of the risk factors known to cause circulatory system and heart disease, but a consistent absence of the healthy worker effect for circulatory disorders and heart disease in numerous mortality studies, and the positive relationship observed in one study between total TEQs and circulatory diseases, suggest the need for additional research in this area. These studies should also include methods to quantify subject exposure to 2,3,7,8-TCDD and control of confounders related to circulatory diseases.
The available epidemiologic studies on immunologic function in humans relative to exposure to 2,3,7,8-TCDD do not describe a consistent pattern of effects among the examined populations. Two studies of German workers, one exposed to 2,3,7,8-TCDD and the other to 2,3,7,8-tetrabrominated dioxin and furan, observed dose-related increases of complements C3 or C4 (Zober et al., 1992; Ott et al., 1994), while the Ranch Hands continue to exhibit elevations in IgA (Roegner et al., 1991; Grubbs et al., 1995). Other studies of groups with documented exposure to 2,3,7,8-TCDD have not examined complement components to any great extent or observed significant changes in IgA. Suggestions of immunosuppression in a small group of exposed workers as a result of a single test should be confirmed in other cohorts of similarly exposed populations (Tonn et al., 1996). Comprehensive evaluation of immunologic status and function of the NIOSH, Ranch Hand, and Hamburg chemical worker cohorts found no consistent differences between exposed and unexposed groups for lymphocyte subpopulations, response to mitogen stimulation, or rates of infection (Halperin et al., 1998; Michalek et al., 1999; Jung et al., 1998; Ernst et al., 1998). However, in a single study, T-cell response to Inferon-( in TCDD-exposed workers was negative in isolated peripheral blood mononuclear lymphocytes and positive in diluted whole blood (Ernst et al., 1998). More comprehensive evaluations of immunologic function with respect to exposure to 2,3,7,8-TCDD and related compounds are necessary to assess more definitively the relationships observed in nonhuman species. Longitudinal studies of the maturing human immunologic system may provide the greatest insight, particularly because animal studies have found significant results in immature animals, and human breast milk is a source of 2,3,7,8-TCDD and other related compounds. Additional studies of highly exposed adults may also shed light on the effects of long-term chronic exposures. Therefore, there appears to be too little information to suggest definitively that 2,3,7,8-TCDD, at the levels observed, causes long-term adverse effects in adult humans.
The Vietnam Experience Study found a significant relationship between service in Vietnam and sperm abnormalities, whereas the Ranch Hand study did not confirm these results when exposure was defined by both cohort status and 2,3,7,8-TCDD levels. However, the data on alterations in male reproductive hormone levels associated with occupational exposure to 2,3,7,8-TCDD emphasize that further research in these areas is required.
Two published reports of infertility patients (Mayani et al., 1997; Pauwels et al., 1999) have raised the potential for an association between endometriosis and TCDD/TEQ exposure. These studies are small and of limited power. Studies of women from Seveso and New York State are currently underway and will add to the database on this outcome.
This section includes all developmental effects except those postnatal developmental effects which are covered by outcome (thyroid, neurobehavioral outcomes, and AST and ALT). Outcomes related to fertility are also reported (e.g., time to pregnancy, birth rates, semen changes).
A variety of study designs, including case-control, ecologic, cross-sectional, and historical cohort designs, have addressed the issue of 2,3,7,8-TCDD and reproductive effects in humans. Unfortunately, the different criteria for case definitions across studies make it difficult to compare the results. In addition, the method of case ascertainment for certain endpoints influences the rate of events observed. The Vietnam Experience Study substudies of veteran-reported birth defects compared with those identified through hospital records demonstrated that rates of self-reported outcomes differed by exposure status. Moreover, predictive value of self-reported events was poor in both cohorts. In contrast, rates of birth defects in the Ranch Hand study were similarly reported by the Ranch Hands and controls. Both groups underreported 7% of birth defects in children conceived prior to their SEA tour and 14% after their tour of duty.
With the exception of the finding that Vietnam veterans were more than twice as likely to have low sperm concentrations (OR = 2.3, 95% CI = 1.2-4.3), no effect measure greater than 2 was noted in any of these investigations. This is not surprising, given the limitations of the studies, particularly with regard to exposure misclassification. Therefore, the trends across these studies carry more import than "statistically significant" results.
Although these studies have restricted inclusion of reproductive events to those that occurred after exposure to 2,3,7,8-TCDD was suspected, no study has evaluated 2,3,7,8-TCDD levels at the time of the outcome. Determination of a 2,3,7,8- TCDD dose-response relationship with adverse reproductive outcomes would not be valid unless individual 2,3,7,8-TCDD levels were available. The recent Ranch Hand study estimated the 2,3,7,8-TCDD levels at the time of the developmental outcome, which is an important contribution toward understanding this phenomenon. However, with regard to early losses, this analysis would not be able to address those occurring very early in gestation, before recognition of the pregnancy by the woman or her physician, or their subsequent effect on identification of adverse outcomes identified later in pregnancy or at birth..
A growing body of animal research described in Chapter 5 lends biological plausibility to the association between dioxin and most of the reproductive endpoints evaluated in these studies, with the notable exception of molar pregnancies. There is growing evidence that dioxin affects testis and accessory gland weight, testicular morphology, spermatogenesis, and fertility in males. A model for a paternally mediated dioxin effect on congenital malformations has not been reported; however, increased interest in this area (Olshan and Mattison, 1994) may yield more information on this topic. Among female animals, the primary reproductive endpoints that have been examined include decreased fertility and pregnancy loss. The mechanism by which 2,3,7,8-TCDD causes adverse reproductive and developmental effects has not been well described, although considerable insight has been gained from research focusing on the Ah receptor. Although the Ah receptor has been linked with birth defects in several mouse strains, it appears that the mechanism of effect may be dependent on the outcome evaluated, as well as other dioxin congeners to which the population is exposed. Clearly, these relationships in humans have not been adequately investigated. The discovery in the Times Beach, Missouri; CDC; and Ranch Hand studies that self-reported dioxin exposure and exposure indices developed from the analyses of 2,3,7,8-TCDD in soil and military records are poorly correlated with serum 2,3,7,8-TCDD levels aids in understanding the inconsistencies of the research to date regarding 2,3,7,8-TCDD and effects. Thus, because of the likelihood of exposure misclassification in those studies lacking direct measures of exposure, the findings have historically been severely limited.
Miscarriages were investigated in several studies with different designs and varied patterns of parental exposure. Events were generally ascertained by self- or spousal report. When case ascertainment was through medical records, such as in the Ranch Hand study or the Vietnamese investigations, the events are by definition restricted to those miscarriages that were clinically recognized. Research in the area of pregnancy loss indicates that 30%-50% of all conceptions are lost prior to or during implantation (Hertig et al., 1959). The rate of loss between implantation and expected first menstrual period ranges from 22% to 30% (Wilcox et al., 1988; Sweeney et al., 1988). Thus it is clear that restriction of the examination of pregnancy loss to those events that are ascertained through medical records, or even self-reports, results in excluding a large proportion of the outcome of interest. In studies of environmental factors and spontaneous abortion where information is lacking concerning conception, "the conflation of different doses with different effects can mislead" (Kline et al., 1989). Because of these discrepancies, it would not be.
meaningful to pool the results of the research on the association between dioxin exposure and miscarriage to judge the "consistency" of the association. Overall, it must be acknowledged that the data compiled to date are inadequate to address this issue. To simply enumerate and compare the number of "positive" versus "negative" studies to ascertain consistency in the research would be inappropriate. The reasons for this have been described above in detail, with emphasis on the high (40%-50%) exposure misclassification that has been documented in the majority of these investigations, the small sample sizes evaluated, lack of data on dioxin levels at the time of conception, and the unknown impact of early pregnancy loss on identification of birth defects. The animal and human evidence for a 2,3,7,8-TCDD-pregnancy loss relationship is sufficiently suggestive to warrant further investigation. Several studies of various designs and populations have demonstrated weak but consistent associations (Reggiani, 1978; Hatch, 1984b; Constable and Hatch, 1985; Huong et al., 1989; Phuong et al., 1989a; Stellman et al., 1988), whereas others have not (Townsend et al., 1982; Smith et al., 1982; Cutting et al., 1970; Kunstadter, 1982; Report to the Minister for Veterans' Affairs, 1983; Stockbauer et al., 1988; Erickson et al., 1984; Centers for Disease Control Vietnam Experience Study, 1989). (These studies include those that should be restricted to the assessment of military service in Vietnam and reproductive events.) Only one study, an analysis of the Ranch Hand developmental data (Wolfe et al., 1995), has used biological measurements and estimated the levels present around the time of conception. This study did note a modest increase in recognized spontaneous abortions and stillbirths at the low, but not the high, level. Thus, the Ranch Hand study leaves several questions unanswered, including the determination of a dose-response level and the impact of very early pregnancy losses on rates of recognized fetal death and birth defects that survive long enough to be "counted."
The confusing evidence regarding the relationship between dioxin exposure and birth defects suffers not only from the limitations described above for the studies of miscarriage but also from the lack of power to evaluate specific types of malformations. To increase the power to detect a potential relationship, the studies have combined all birth defects together and calculated an odds ratio for total birth defects. Given evidence for etiologic heterogeneity among subgroups of birth defects (Khoury, 1989), it is probable that this approach dilutes the effect measure. These studies also should more carefully examine type of parental exposure, i.e., paternal, maternal, or both. Timing of exposures and potential biological mechanisms for birth defects are different for maternal and paternal exposures. The field of paternally mediated effects is rather new, but future research may assist in the interpretation of these results (Olshan and Mattison, 1994). If dioxin exposure is related to malformations among the offspring conceived after paternal service in Vietnam, then the effect must occur either premeiotically, or anew with continuing circulating levels of TCDD. Some animal studies have found that spermatogonia and spermatocytes (premeiotic spermatogenic cells) were able to repair DNA after exposure to toxic agents, whereas spermatids and spermatozoa did not have this capability (Lee and Dixon, 1978). A few studies (Hatch, 1984a; Constable and Hatch, 1985; Huong et al., 1989; Phuong et al., 1989a), including those investigations of the Yusho and Yu Cheng incidents (Rogan, 1982; Yen et al., 1989; Rogan et al., 1988) have suggested an association. Many studies have failed to find a relationship between dioxin and birth defects (Townsend et al., 1982; Smith et al., 1982; Cutting et al., 1970; Kunstadter, 1982; Stockbauer et al., 1988; Erickson et al., 1984; Centers for Disease Control Vietnam Experience Study, 1989; Mastroiacovo et al., 1988). Again, however, in view of the serious limitations of these studies of 2,3,7,8-TCDD and developmental and reproductive events, it must be concluded that the relationship between paternal dioxin exposure and congenital malformations remains unknown. However, if military service in Vietnam is the exposure of interest, there is only modest evidence to support an association with birth defects. Most of the data on grouped birth defects have very small numbers.
Finnish investigators examined the association of enamel hypomineralization of permanent first molars in 6-7-year-old children and TEQ exposure through breast feeding (these teeth develop during the first 2 years of life) (Alaluusua et al., 1996; Alaluusua et al., 1999). These data present interesting findings. Unfortunately, the presentation of the results is incomplete, so the potential biological significance cannot be assessed. This would be an interesting outcome to examine in additional studies.
Sex ratio at birth was significantly depressed in a group of 17 children in zone A, Seveso, in the years shortly following the industrial accident. This pattern disappeared a few years later. Other investigators who have studied developmental outcomes examined their data and wrote short reports or letters to the editor on this topic. The decrease in sex ratio was not supported in other groups examined. These analyses provided limited data for one or more of the following reasons: limited exposure data, no adjustment for other risk factors/confounders, assumption of a gold standard of 51.4% males (and not having comparison groups), or small numbers. Although the initial finding in Seveso is of interest, examination of other data sets does not support an alteration in sex ratio in exposed parent(s) at this time. Because this outcome can easily be collected in studies of developmental effects, more thorough examination of this outcome could be useful.
Growth measures include endpoints such as intrauterine growth retardation (IUGR), low birth weight, and postnatal growth. Available evidence does not support an association between paternal dioxin level and low birth weight (Centers for Disease Control Vietnam Experience Study, 1989; Wolfe et al., 1992b, 1995; Michalek et al., 1998). In the Rotterdam study, decrements in length (but not other measures of growth) were observed early (months 3-7 postnatally) and associated with PCB levels, but disappeared with increasing age (Patandin et al., 1998). The Finnish data (Vartiainen et al., 1998) are interesting because birth weight did decrease in males with increasing TEQ, but the lack of detail on the statistical analyses makes interpretation difficult.
Additional reproductive outcomes that were evaluated in a subset of the studies include molar pregnancies (in the Vietnamese studies), neonatal and infant death, and childhood cancer and mortality. Mainly because of small sample sizes, it is difficult to reach conclusions regarding neonatal, infant, and child mortality and childhood cancers. Recently, the increased risk for neonatal death observed in the Ranch Hand study, the only study with individual TCDD levels, was investigated. Some changes were observed in the Ranch Hand study for preterm birth and neonatal death, but these did not follow an exposure-response relationship (Michalek et al., 1998). In conclusion, the research to date has been successful in resolving some confusion surrounding the evidence for an association of dioxin exposure and various developmental and reproductive endpoints in humans. High occurrence of exposure misclassification, differences in case definitions across studies, and small sample sizes have severely limited the power of these studies to address these questions. Additional research that includes a measure of dioxin level at the time of conception for both the father and mother is necessary if the effect of dioxins on the spectrum of reproductive outcomes is to be understood.
The following section reviews endpoints that were described in groups shortly after exposure to 2,3,7,8-TCDD but were not observed as chronic effects in studies conducted many years after exposure ceased. Also reviewed are endpoints observed as long-term effects in single studies.
Dermatologic conditions other than chloracne, such as hyperpigmentation, hypertrichosis, and eyelid cysts, have been related to exposure to 2,3,7,8-TCDD in early case reports (Ashe and Suskind, 1950; Suskind et al., 1953; Bleiberg et al., 1964; Poland et al., 1971; Bauer et al., 1961; Goldman, 1972; Jirasek et al., 1974; Oliver, 1975). However, these conditions may have been acute effects of 2,3,7,8-TCDD exposure that resolved over time or may be residual effects of chloracne, because they appear to occur more frequently in individuals with persistent chloracne (Suskind and Hertzberg, 1984). These conditions were not observed in studies in which the cohorts were examined years after cessation of exposure, in individuals with the potential for high exposure, or in those with high adipose or serum 2,3,7,8-TCDD levels (Moses et al., 1984; Webb, 1989; Roegner et al., 1991; Burton et al. 1998). Actinic keratosis, Peyronie's disease, and basal cell carcinoma may not be associated with 2,3,7,8-TCDD. All three conditions were observed in only one study (Suskind and Hertzberg, 1984; Lathrop et al., 1984) and were not observed in studies of individuals with similar potential for exposure (Ott et al., 1994).
A number of studies reported elevated liver enzymes, particularly AST and ALT, among individuals who were being exposed at the time of the measurement (May, 1973) or whose exposure was within a few years of the measurement (Jirasek et al., 1974; Mocarelli et al., 1986; Caramaschi et al., 1981). Follow-up studies or longitudinal analyses of exposed cohorts suggest that the increase in enzyme level resolves over time (Mocarelli et al., 1986; Assennato et al., 1989; Pazderova-Vejlupkova et al., 1981; May, 1982). In studies of exposed populations tested many years after exposure ceased, levels of AST and ALT were within normal range (Calvert et al., 1992; Webb et al., 1989; Roegner et al., 1991; Suskind and Hertzberg, 1984; Moses et al., 1984). D-glucaric acid was tested in a number of 2,3,7,8-TCDD-exposed populations as an indicator of enzyme induction (Ideo et al., 1985; Martin, 1984; Roegner et al., 1991; Calvert et al., 1992). Shortly after the TCP reactor release, D-glucaric acid levels in Seveso children were elevated (Ideo et al., 1985). No other studies of exposed groups tested 5 to 37 years after exposure ceased found elevations of this enzyme (Martin, 1984; Roegner et al., 1991; Calvert et al., 1992). These data suggest that certain hepatic enzymes are increased as a response to high, exogenous exposure to 2,3,7,8-TCDD. Once the exposure ends, the enzyme levels seem to decrease over time, as observed in the Seveso populations (Mocarelli et al., 1986; Ideo et al., 1985). Additional evidence of the acute nature of AST, ALT, and D-glucaric acid elevations is demonstrated by the lack of such increases in studies of highly exposed groups conducted long after exposure ceased (Calvert et al., 1992; Roegner et al., 1991; Grubbs et al., 1995; Martin, 1984). As in the case of ALT, AST, and D-glucaric acid, hepatomegaly appears to be a condition reported in case reports after high exposure to 2,3,7,8-TCDD-contaminated chemicals, particularly among TCP production workers examined after a TCP reactor explosion and among Seveso residents (Ashe and Suskind, 1950; Suskind et al., 1953; Jirasek et al., 1974; Reggiani, 1980a). Later studies conducted after exposure ceased failed to find excess dose-related hepatomegaly in the exposed populations (Bond et al., 1983; Suskind and Hertzberg, 1984; Moses et al., 1984; Calvert et al., 1992; Centers for Disease Control Vietnam Experience Study, 1988a; Roegner et al., 1991; Webb et al., 1989; Hoffman et al., 1986). However, the absence of an effect in cross-sectional studies does not confirm the lack of an effect in the past.
Early case reports suggest that exposure to 2,3,7,8-TCDD chemicals may cause temporary respiratory irritation (Zack and Suskind, 1980) and tracheobronchitis (Goldman, 1972). The data from two cross-sectional medical studies provide weak evidence of slightly decreased lung function among exposed individuals (Suskind and Hertzberg, 1984; Roegner et al., 1991). In these studies, the effects may be due more to smoking (Roegner et al., 1991) or to a substantial age difference between the exposed and unexposed groups (Suskind and Hertzberg, 1984). One study of highly exposed TCP production workers found no relationship between serum 2,3,7,8-TCDD levels and chronic obstructive pulmonary disease, bronchitis, or decreased pulmonary function (Calvert et al., 1992). In conclusion, case reports indicate that intense acute exposure to 2,3,7,8-TCDD can produce respiratory irritation. However, the findings from controlled epidemiologic studies conducted many years after exposure do not convincingly support an association between 2,3,7,8- TCDD exposure and chronic effects on the respiratory system.
The results of case reports and epidemiologic studies demonstrate that exposure to 2,3,7,8-TCDD-contaminated materials is associated with symptoms referable to the central and peripheral nervous systems shortly following exposure and, in some cases, lasting many years (Filippini et al., 1981; Ashe and Suskind, 1950; Moses et al., 1984). Overall, however, the neurologic status of workers, community residents, and Vietnam veterans exposed to 2,3,7,8- TCDD and evaluated from 5 to 37 years after last exposure appears to be normal (Centers for Disease Control of Vietnam Experience Study, 1988a; Lathrop et al., 1984; Sweeney et al., 1993). The data suggest that, although exposure to 2,3,7,8-TCDD may have been extensive as in exposed workers, Ranch Hands, and Seveso residents, the effects described in case reports may have been transient (Filippini et al., 1981; Lathrop et al., 1984; Centers for Disease Control Vietnam Experience Study, 1988a,b; Assennato et al., 1989; Alderfer et al., 1992; Sweeney et al., 1993). The findings of recent studies suggest that in adults there are no long-term neurologic effects caused by even high exposure to 2,3,7,8-TCDD-contaminated materials, but there is very little information with which to examine the effects of exposure on the developing human neurologic system.
In rats and mice, exposure to 2,3,7,8-TCDD has been clearly shown to alter porphyrin metabolism (Goldstein et al., 1973; Smith et al., 1981; Jones and Chelsky, 1986; DeVerneiul et al., 1983; Cantoni et al., 1981; Goldstein et al., 1982). Whether 2,3,7,8-TCDD is associated with porphyrin changes in humans, particularly PCT, is a subject of unresolved debate. It has been suggested that the PCT and elevated urinary porphyrins observed in the New Jersey and Czechoslovakian workers during the years of operation of the plants were the result of exposure to hexachlorobenzene, which was produced at the same time as TCP (Pazderova-Vejlupkova et al., 1981; Jones and Chelsky, 1986). These statements have not been corroborated with strong studies. In the follow-up studies, urinary porphyrin levels of these TCP production workers were not elevated (Pazderova-Vejlupkova et al., 1981; Poland et al., 1971) or did not differ from levels in the control group (Calvert et al., 1993). Doss et al. (1984) also described transient elevations in coproporphyrins among 22 Seveso residents exposed to 2,3,7,8-TCDD. Because 2,3,7,8-TCDD is a porphyrigen in rats and mice, it has been of interest to determine whether exposures may have contributed to the observed changes in porphyrin levels in human populations. The NIOSH study could not address the question of etiology or transient porphyria, but it did not find porphyria in highly exposed workers many years after their occupational exposure. However, although porphyria was not found after 2,3,7,8-TCDD exposure, it may be an outcome of exposure to PCBs when there is a co-exposure to chlorophenols, as suggested by the study of pentachlorophenol workers (Hryhorczuk et al., 1998). Further research is recommended to examine these findings.
Many effects of 2,3,7,8-TCDD exposure in animals resemble signs of thyroid dysfunction or significant alterations of thyroid-related hormones. In the few human studies that examined the relationship between 2,3,7,8-TCDD exposure and hormone concentrations in adults, the results are mostly equivocal (Centers for Disease Control Vietnam Experience Study, 1988a; Roegner et al., 1991; Grubbs et al., 1995; Suskind and Hertzberg, 1984). However, concentrations of thyroid binding globulin (TBG) appear to be positively correlated with current levels of 2,3,7,8-TCDD in the BASF accident cohort (Ott et al., 1994). Little additional information on thyroid hormone levels has been reported for production workers and none for Seveso residents, two groups with documented high serum 2,3,7,8-TCDD levels.
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