Review of the toxicology of chlorpyrifos with an emphasis on human exposure and neurodevelopment.

This review examines the large body of toxicological and epidemiological information on human exposures to chlorpyrifos, with an emphasis on the controversial potential for chlorpyrifos to induce neurodevelopmental effects at low doses. The results of this review demonstrate that the use of urinary 3,5,6-trichlorpyridinol (TCPy), a metabolite of chlorpyrifos as a biomarker of nonoccupational exposure is problematic and may overestimate nonoccupational exposures to chlorpyrifos by 10-to 20-fold because of the widespread presence of both TCPy and chlorpyrifos-methyl in the food supply. Current "background" (nonoccupational) levels of exposure to chlorpyrifos are several orders of magnitude lower than those required to inhibit plasma cholinesterase activity, which is a more sensitive target than nervous system cholinesterase. However, several in vitro studies have identified putative neurodevelopmental mechanisms that are altered at concentrations of chlorpyrifos below those that inhibit cholinesterases. Although one human cohort study reported an association between maternal and cord blood chlorpyrifos levels and several measures of neurodevelopment, two other cohort studies that utilized urinary TCPy as a surrogate for chlorpyrifos exposure did not demonstrate an association. Although the weight of the scientific evidence demonstrates that current levels of chlorpyrifos exposure will not have any adverse effects on neurodevelopment that might result from inhibition of nervous system cholinesterases, several recent studies propose alternative mechanisms. Thus, further in vivo investigation on neurodevelopment in an appropriate animal model is needed; additional epidemiological studies may be warranted if a suitable, chlorpyrifos-exposed cohort can be identified and more rigorous measures of exposure are utilized.

Polychlorinated dibenzo-p-dioxins and the human immune system: 4 studies on two Spanish families with increased body burdens of highly chlorinated PCDDs.

The consequences of exposure of people to highly chlorinated polychlorodibenzo-p-dioxins (PCDDs) are much less known than those of TCDD. We report on levels of PCDDs (and PCDFs) in 13 members of two families poisoned by contaminated cooking oil. Originally, all persons displayed chloracne as an early symptom. Persisting hexa- and higher chlorinated PCDDs could be analysed many years after exposure. Highest values found in blood lipids were: OCDD 660,000 pg/g; HpCDD 58,000 pg/g; HxCDDs: 3500 pg/g. None of the participants exhibited increased TCDD levels at the time of study. During a period of 6 years, HpCDD and OCDD disappeared from the blood lipids much faster in persons exposed as children or young adults, than from lipids of their parents. Surface receptors on blood lymphocytes of the members of the two families and the proliferative capacity of these blood cells in the presence of typical stimulants were analysed. Even in family members with the highest body burdens of hexa- to octachlorinated PCDDs we could not detect pronounced changes from a reference population with respect to the immunological markers. Minor deviations of levels of some receptors in a few, but not all, highly exposed persons suggested a similar trend to those reported in previous studies of persons with body burdens of > or =3000 pg TCDD/g blood lipids. An increase in the number of total blood lymphocytes in some subjects exposed as children may have similarity with highly TCDD-exposed children in Seveso.

Kinetics and organotropy of some polyfluorinated dibenzo-p-dioxins and dibenzofurans (PFDD/PFDF) in rats.

While polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/PCDF) and the corresponding polybrominated congeners must be considered as animal teratogens and carcinogens, little information is available on corresponding polyfluorinated compounds (PFDD/PFDF). Kinetic studies on a few fluorinated dibenzo-p-dioxins and dibenzofurans revealed a rapid elimination, suggesting a much lower toxicity than the corresponding polychlorinated and polybrominated congeners. In order to obtain further clues on the possible toxicity, the kinetics and organ distribution (in liver, thymus and adipose tissue) of a PFDD/PFDF-mixture were studied in Wistar rats after intravenous application. The congeners investigated included four of the 2,3,7,8-substituted, and four of the not-2,3,7,8-substituted dibenzo-p-dioxins, as well as two dibenzofurans. The main result of our studies is the finding that the concentration in the thymus of several of the 2,3,7,8-substituted PFDD/PFDF greatly exceeded that in hepatic tissue. An organotropy quite different from that of the other polyhalogenated congeners must be expected, immunosuppressive effects presumably being the predominant ones. Overall, the elimination half-life of all the PFDD/PFDF studied is considerably shorter than that of the corresponding polychlorinated or polybrominated congeners, in the rat, suggesting a much lower toxicity in this species. No information is available for other species, e.g. nonhuman primates or humans.

Analysis of reproductive toxicity and classification of glufosinate-ammonium.

CONCLUSION REGARDING CLASSIFICATION OF GLUFOSINATE-AMMONIUM: Science Partners' Evaluation Group (Evaluation Group) has conducted an independent analysis of the herbicide glufosinate-ammonium (GA) relative to its potential to cause reproductive toxicity in humans. Further, the Evaluation Group has evaluated the implementation of Annex 6 of Commission Directive 2001/59/EC (28th ATP of Council Directive 67/548/EEC) and Council Directive 91/414/EEC, with respect to classification of chemicals posing potential reproductive hazards. After consideration of all information available to us relevant to the potential of glufosinate-ammonium (GA) to cause reproductive toxicity, the Science Partners Evaluation Group concludes that no classification of GA is justified. The following form the basis of this conclusion. There are no human data to suggest that GA causes reproductive toxicity in women or in their conceptus. The issue concerning possible reproductive hazard to humans is raised solely on the basis of positive animal test results that show GA to cause preimplantation or implantation losses in rats. SPECIFICALLY: a. Daily treatment with GA had no detectable effect on the earliest stages of the reproductive sequence including gametogenesis, ovulation, mating and conception; b. Treatment with GA interfered with rat gestation before and at the stage when the conceptus implants into the uterus. This effect occurred at doses of 360 ppm in the feed (corresponding to daily doses of 27.8 mg/kg bw) and above; and c. After implantation, no further effect of GA on prenatal and post-natal development was recognized. Previous concerns that GA might be toxic to embryonic stages after implantation were not supported by the data. Abortions and stillbirth seen were associated with, and regarded as secondary to, maternal toxicity. There was no evidence suggesting the induction of malformations in the offspring. The mechanism underlying this adverse effect in experimental laboratory animals is identified-inhibition of glutamine synthetase. Glutamine is essential to the viability of the embryo. The embryo is dependent on a maternal source of the amino acid. For embryo lethality to occur, a significant reduction of maternal glutamine is required. Such reduction in maternal glutamine depends on a significant inhibition of glutamine synthetase by GA. This can only occur when the mother is exposed to very high levels of GA. SPECIFICALLY: a. The reproductive toxicity of GA is confined to very short, early stages of reproduction, during which the conceptus is dependent on maternal glutamine; and b. In order for the effect to occur, significant reduction in maternal blood glutamine level is required, which in turn depends on a significant inhibition of glutamine synthetase, induced by high levels of GA in the maternal system. There is no evidence for accumulation of GA in the mammalian organism beyond a factor of two and no evidence for its metabolic toxification. To raise a concern in humans, women would have to be exposed to GA during the very limited time frame of preimplantation or implantation and the exposure would have to be to the exceedingly high levels necessary to alter the maternal metabolism and, correspondingly, result in glutamine levels in maternal tissue and blood plasma being drastically reduced. There is no basis to suggest that such exposures would occur under conditions of normal handling and use. SPECIFICALLY: a. Under conditions of normal handling and use, operators would never be exposed to GA levels that could potentially inhibit glutamine synthetase to the extent that this inhibition could impair preimplantation or implantation. b. All acceptable exposure measurements and predictive calculations confirm this conclusion, and in fact demonstrate that reasonably foreseeable exposure of workers would be to levels significantly below the AOEL. c. The evidence is also clear that there is no reproductive toxicity hazard to workers upon reentry tosprayed fields, bystanders, consumers or toddlers. The safety margin compared to the NOAEL in animal studies is sufficiently large to assure protection of the health of workers using GA as well as bystanders, consumers, and toddlers. Pursuant to Annex 6 of Commission Directive 2001/59/EC (28th ATP of Council Directive 67/548/EEC), to justify a classification of category 2 there must be sufficient evidence to produce a strong presumption that human exposure to the substance may result in impaired fertility in humans. It is the conclusion of the Science Partners Evaluation Group that there is no reasonable evidence to suggest a strong presumption of impairment. To the contrary, there is clear evidence demonstrating a strong presumption that exposure to GA would not cause the adverse effect demonstrated in rats. Pursuant to Annex 6 of Commission Directive 2001/59/EC (28th ATP of Council Directive 67/548/EEC), to justify a classification of category 3, there must be sufficient evidence to provide a strong suspicion of impaired fertility in humans. There is no basis to conclude that the animal data demonstrating impaired preimplantation or implantation has any relevance to humans in that the effect found in rats only occurs at levels which would never be experienced by workers under conditions of normal handling and use or by bystanders, consumers, or toddlers.

Reproductive toxicology: the science today.

Reproductive toxicology is concerned with chemical or physical agents interfering with fertility in both gender. Adverse effects may be induced directly, especially in adult males by damaging the semen producing epithelium (e.g., DBCP), or indirectly, predominantly by interfering with sex hormonal homeostasis. Many critical events must occur during well-defined periods of prenatal and early postnatal development of the reproductive system. Most of such differentiation processes, several of which in the male critically depend on inducing influences of androgens, cannot take place at later stages, and lack of "imprinting" will result in irreversible defects or dysfunctions. These processes might be disturbed by interfering agents (e.g., by anti-androgens: feminization), provided that the exposure is high enough. Several of the processes known to be essential for male development can also be altered in females by exposure to a large excess of androgens (masculinization). Essential processes required for normal male development include: 1) androgen-dependent differentiation of the male phenotype during late embryonic development, 2) differentiation of the male secondary sex organs during the fetal period, 3) formation of a fixed number of Sertoli cells during the perinatal period, 4) imprinting of male sexual behavior in defined brain areas during the perinatal period, 5) imprinting of the pulsatile GnRH regulation of hypophysial hormone formation in both gender via the hypothalamico-hypophysial axis, and 6) differentiation of the male organism during puberty. Many effects on fertility can be induced on the adult organism. Besides a direct action on the receptors, inhibition of the feed back mechanism that guarantees sex hormonal homeostasis is another mode of action. Many synthetic steroid compounds exhibit effects on more than one receptor, thus causing a complex situation. This must also be taken into account when analyzing possible effects of "ecohormones." Adverse hormonal actions are well established from experience in clinical and experimental medicine, using either natural or synthetic sex hormones, or enzyme inhibitors. Possible effects of "environmental" agents either mimicking or inhibiting sex hormonal actions are less well studied in clinical trials. Because of considerable species differences in hormonal effects, especially in pharmacokinetics, data of animal studies are of limited predictive value for extrapolations in preventive hazard minimization (but may be useful for revealing possible mechanisms of action). Data of in-vitro studies are even less suitable for extrapolations. It may be doubted that exposure of the general population to "ecohormones" or "xenohormones" is sufficient to induce clear-cut clinical effects. Adverse effects induced by, e.g., greatly unbalanced diets or after accidental overdoses cannot be excluded.

Tissue concentrations and induction of a hepatic monooxygenase in male Wistar rats after repeated doses of defined polychlorinated dibenzo-p-dioxin and dibenzofuran (PCDDs and PCDFs) mixtures.

Two groups of male Wistar rats were treated 16 times (every 3rd day) subcutaneously with a defined mixture of polychlorinated dibenzo-p-dioxins (PCDDs) or of polychlorinated dibenzofurans (PCDFs). These mixtures contained no measurable amount of 2,3,7,8-TCDD. Each single dose was calculated to contain either 57 ng I-TEq (international 2,3,7,8-T4CDD toxicity equivalencies)/kg body weight of the PCDD mixture or 39 ng I-TEq/kg body weight of the PCDF mixture. Both mixtures contained a large excess of non-2,3,7,8-substituted congeners. The activities of ethoxyresorufin O-deethylase (EROD) in liver microsomes were correlated with the corresponding concentrations of PCDDs or PCDFs in hepatic tissue. Data were compared with results obtained after single injections of 2,3,7,8-tetrachlorodibenzo-p-dioxin (2,3,7,8-T4CDD). As expected, a complex kinetic situation resulted, because of the different tissue distributions and elimination half-lives of the various congeners: (1) 2,3,7,8-substituted PCDDs: the time course of the concentrations in liver and adipose tissue was similar for all congeners, the levels increased during the treatment period and decreased after treatment. Tissue concentrations of all 2,3,7,8-substituted PCDDs were considerably higher in liver than in adipose tissue. The liver/adipose tissue concentration ratios increased with the degree of chlorination. The ratio of 1,2,3,7,8-P5CDD was much lower than those of all other 2,3,7,8-substituted congeners. (2) 2,3,7,8-substituted PCDFs: 1,2,3,7,8-P5CDF was rapidly eliminated from liver and adipose tissue while 2,3,4,7,8-P5CDF largely persisted after the treatment period in both tissues. 2,3,7,8-T4CDF was eliminated even more rapidly than 1,2,3,7,8-P5CDF and could not be detected after treatment in both tissues. Time courses of the concentrations of 2,3,4,7,8-P5CDF, H6CDFs, H7CDFs and OCDF in liver and adipose tissue were similar: the levels of all congeners increased during the treatment period but no clear-cut decrease was observed within 34 days after the last treatment. Tissue concentrations of all 2,3,7,8-substituted PCDFs were higher in liver than in adipose tissue. The liver/adipose tissue concentration ratios increased with the degree of chlorination. The ratios of 2,3,7,8-T4CDF and 1,2,3,7,8-P5CDF were much lower than those of all other 2,3,7,8-substituted congeners. (3) non-2,3,7,8-substituted PCDDs and PCDFs: a number of non-2,3,7,8-substituted PCDD and PCDF congeners were found in both tissues in concentrations below 1 ng/g. In adipose tissue nearly all congeners were found during the treatment period showing a decrease after the treatment. In liver samples, many higher chlorinated PCDF congeners (with >4 chlorine atoms) could be detected. Most of those substituted in three of the four 2, 3, 7 and 8-positions persisted after treatment. In contrast, only one 1,4,6,9-substituted isomer of each PCDD homologue group was found during treatment with high recoveries after the third injection, but a rapid decline occurred already during the treatment period. (4) EROD activity: a good linear relationship (when using a double-log plot) between the EROD activities and the hepatic concentrations (ng I-TEq/g tissue) was found both in the PCDD-treated (r2= 85.8%) and in the PCDF-treated group (r2=87.3%). A similar correlation (r2=95.6%) was observed in rats treated with 2,3,7,8-TCDD alone (concentration range in liver tissue: 0.2 to 9.7 ng/g wet weight). The concentration-response curves for both the PCDD and PCDF mixtures run parallel to the curve for 2,3,7,8-T4CDD. However, the inductive potency of the PCDD or PCDF mixture was approximately 3-fold or 4-fold lower, respectively, compared with the inductive potency of 2,3,7,8-T4CDD. Thus, the I-TE factors overestimated the potency of the mixtures in the concentration range tested and taking EROD induction as an end point.