PBBs toxic effects

Robertson, L.W., A. Parkinson, S. Bandiera, L. Lambert, J. Merrill, and S.H. Safe. 1984. PCBs and PBBs biologic and toxic effects on C57BL/6J and DBA/2J inbred mice. Toxicology 31 191-206. 

This section will describe clinical practice and research concerning methods for reducing toxic effects of exposure to PBBs and PBDEs. However, because some of the treatments discussed may be experimental and unproven, this section should not be used as a guide for treatment of exposures to PBBs and PBDEs. When specific exposures have occurred, poison control centers and medical toxicologists should be consulted for medical advice. No texts were found that provided specific information about treatment following exposures to PBBs and PBDEs. 

Information on toxic effects of acute-duration exposure to PBBs by routes other than oral are limited to data on hepatic, renal, dermal, and ocular effects of inhalation and dermal exposure in rats or rabbits (Millischer et al. 1980 Needham et al. 1982 Norris et al. 1975a Waritz et al. 1977), but these data may not be reliable due to study limitations and possible delayed lethality. Quantitative data for inhalation and dermal absorption of PBBs are lacking. Studies of inhalation and dermal absorption following exposure to soil containing PBBs (i.e., bioavailability studies) would be useful for assessing risk at a hazardous waste... 

The mechanism of toxic action of PBBs is not completely understood and no methods exist to block the toxic response due to exposure to PBBs. A more complete characterization of the cytosolic Ah receptor protein, to which some PBB congeners are thought to bind, and understanding of physiological effects of receptor blockage would be useful for the possible identification of blockers of Ah receptor-mediated toxic effects. Further studies aimed at elucidating the nonreceptor-mediated mechanism of action of some PBBs would also be valuable. 

Some weeks later FDA investigators and Farm Bureau people found a partly filled bag of Firemaster at the feed mill. Within the month (May 1971 ) action (tolerance) levels were established by the FDA for PBB contamination. The Michigan Department of Agriculture began quarantining contaminated farms—a process that went on until the end of 1975- Another six months saw the original MACs for PBBs further lowered to 0.3 ppm for meats and dairy products, and 0.05 ppm for eggs and feed, as the toxic effects in animals were noted at lesser concentrations, and the routine detection of PBBs at these lower concentrations became feasible. 

PBBs cause immuno-suppression at levels that also cause a number of the other toxic effects described [5],... 

Several countries have been given restrictions on the use of compounds because of potential toxic effects in humans. In the European Community, the use of tris(2,3-dibromopropyl) phosphate (EC Directive 76/769/EEC) and tris(l-aziridinyl)phosphine oxide (EC Directive 83/264/EEC) in textiles has been banned. In 1977, the U.S. Consumer Product Safety Commission banned the use of tris(2,3-dibromopropyl) phosphate in children s clothing (ICPS, 1995). The European Community has also banned the use of PBBs in textiles (EC Directive 83/264/EEC). Several countries have either taken or proposed regulatory actions on PBBs. In addition, controls on the emissions of dioxins and furans from municipal solid waste incinerators have been implemented in the United Kingdom under the Environmental Protection Act (1990). Germany has developed rules for the maximum content of selected 2,3,7, and 8 substituted polychlorinated dibenzo-pura-dioxins and dibenzofurans in products. Recently, the European Commission has issued a proposal to ban the production and use of PentaBDE. In U.S. A., on the other hand, there are currently no regulations on PBDE production or use. PBBs have not been used widely in Europe and also in U.S.A. the production of the main mixture, hexabromobiphenyl (Firemaster BP-6), ceased in 1974, after the Michigan disaster. ... 

Such epidemiological studies entail cohorts with moderately elevated PbB with no history of chelation treatment or undefined earlier exposures. Children sustaining excessive exposures above medical guidelines during such serial studies would be removed. By contrast, high-risk children presenting with PbB levels requiring that same level of intervention appear to show that maximum or peak Pb is the exposure metric with the closest association with toxic effects (Chisolm, 1982 Chisolm and Harrison, 1956). 

The available evidence for PbBO as a dose metric in Pb toxicology and epidemiology documents that it (1) is a cumulative but not inert dose/exposure metric which both serves to quantify stored Pb and, in metabolicaUy diverse settings, released Pb as an endogenous source of systemic Pb exposure (2) may be a better Pb exposure correlate with various toxic effects than PbB in settings where resorptive releases are significant or long term, e.g., in retired Pb workers or in children, older children, or adults whose Pb-exposure histories foretell sizeable stores of bone Pb and (3) is likely a measure of the source of much of the Pb that is released on provocative chelation and that better reflects potential toxicity risk than PbB. However, the relative robustness of PbBO in dose—response relationships versus other measures of toxic dose, e.g., PbB or chelatable Pb, wiU vary case to case. 

A critical question in fetal Pb toxicokinetics is how one best measures dose—response relationships among various exposure biomarkers as well as relationships governing dose—toxic response relationships. It is now accepted that bone Pb is a better biomarker in constmcting dose—toxic response relationships for a variety of toxic effects than are indicators such as PbB. This was demonstrated in Chapter 13 describing cardiovascular effects of Pb. 

Discussing PbB and bone Pb as biomarkers of ongoing or immediately toxic exposure requires that these measures be kept distinct from measures of body lead burden. These measures have been the source of considerable confusion in the environmental health literature. Body lead burden, the total amount of body Pb regardless of relative immediate mobilizability, may not be as sensitive a measure of concurrent exposure risk as a measure such as PbB in cases where circulating Pb in vivo is available to have immediate toxic impacts in target tissues and where the time for toxic response is short. PbB can rise considerably and induce toxic effects in sensitive organs... 

The role of bone as the principal body depository for Pb makes it the principal biomarker of body lead burden. By comparison, the total Pb content of whole blood is considerably less (Rabinowitz et al., 1976). Given current evidence, however, the kinetic and metabolic mobility of bone Pb is sufficient to qualify this Pb reservoir as a biomarker of Pb exposure and a biomarker relevant for much longer periods than PbB. Some recent epidemiological studies, like those for toxic effects described in the toxicity chapters of this book, indicate associations of bone Pb with long-term chronic adverse effects in human populations while PbB provided negative results. 

Table 22.1 sets forth dose—toxic response relationships for young children in terms of PbB thresholds for adverse effects as exposures increase. Here, thresholds are taken to mean those lowest PbB levels that have been reliably associated in the international literature with toxic effects. Identifying some empirical threshold on the basis of the low point of ranges in regression or other statistical analyses should not be taken to mean an actual threshold exists (see above on lack of thresholds for Pb in humans). 

The full spectrum of toxic effects ranges from subtle disturbances in multiple organs and systems at the low end of exposure, <10 p-g/dl, to death or profound brain damage at the high, extreme concentrations >100 p.g/dl. Both are increasingly likely at PbB values above 100 p.g/dl. For ease of presentation, Table 22.1 notes the specific new effects in terms of a lowest PbB value associated with those toxic endpoints. However, as PbB increases in the table, the number of effects and the severity of effects first identified at lower associations will increase. Finally, the lowest levels of PbB empirically linked to adverse health effects in children are not thresholds below which the specified effects do not occur. 

Table 22.3 presents the full-spectrum of dose—toxic response relationships for lead toxicity in adults in terms of lowest reported adverse effect level thresholds. As with the earlier tables, only the lowest values for PbB in the associations specific for the indicated range are identified. It is also understood with these adult tabulations that as PbB values rise above lowest levels of determined associations, those toxic effects increase in severity and multiplicity. Toxicity criteria for reliability and validity parallel those enumerated for the earlier childhood tables. A number of dose-responses scaled to dose/exposure are identified in the table. In many cases, including hematotoxicity, peripheral neurotoxicity, and nephrotoxicity, effects are qualitatively similar but occur at higher empirically measured thresholds. 

The available hterature, as described in the adverse effects chapters, allows one to conclude that bone Pb is a better dose metric for long-term, chronic toxic effects in adults than are single measurements of PbB. In fact, that hterature shows associations of trabecular and/or cortical bone Pb with a number of toxic endpoints when no links are shown with PbB levels. This comparative set of outcomes is a strong argument for using bone Pb or combined bone Pb and PbB as exposure metrics. 

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