According to the WHO’s (1981)
Environmental Health Criteria for manganese, one may estimate for healthy western adults an average oral intake of Mn via food of about 3.6 mg/day and a GI absorption rate not higher than 5%, which would represent a systemic Mn absorption of 180 μg/day. If the liver had no homeostatic control over the systemic Mn status, the Mn absorbed via the GI tract would lead to a concentration of about 40 μg Mn per liter of circulating blood (assuming 4.5 L of blood for an adult). The Toxicological Profile of Manganese (ATSDR, 2008
) reports a range for whole B-Mn from 4 to 15 μg/L. In Europe (Germany), a biological reference value of 15 μg/L is set for whole blood, a threshold which is the presumed 95th percentile of the distribution in the general population (Pesch et al., 2012
). However, a representative study on whole blood manganese in adult European or American men and women is still lacking, let alone in children and newborns. Whole blood contains four major compartments of Mn: 66% of Mn is in the red blood cells, 4.4% in the plasma, 23.2% in the white blood cells, and 6.6% in the platelets (Milne et al., 1990
). Plasma manganese is considered the most readily biologically available fraction in the blood pool for uptake in the organs. In normal adults without increased exposure to Mn, the P-Mn concentration most likely does not exceed 2 μg Mn per liter plasma (Hoet et al., 2011
). To cope with a daily dietary absorption of 180 μg Mn in normal adults, the homeostatic system for Mn must be very effective to maintain the plasma concentration of Mn within a physiologically normal range, which is accomplished by the bile, the main excretion pathway for Mn in humans.
In this symposium, different biomarkers of exposure to Mn were used to characterize individual exposures: B-Mn and P-Mn for exposure in welders, B-Mn and/or H-Mn for environmental exposure to airborne Mn or to drinking water Mn, and the special case of whole blood Mn as a reflection of the newborns’ “exposure in utero” to the Mn status of pregnant women whose B-Mn rises for physiological reasons. Occupational exposure to Mn occurs usually via inhalation of airborne Mn particulate of which the particle size determines to a large extent the bioavailability of Mn and hence the toxicodynamic outcome. This is particularly the case in welders exposed to welding fume aerosols in which the particulate is more than 90% in the respirable fraction and may reach the alveolar compartment of the lungs. A great part of Mn absorbed via this pulmonary route is conveyed with the blood circulation directly to the brain, thus bypassing the homeostatic control of the liver. Nonrespirable particulate, impacted in the upper airways, travels upwards via the mucociliary escalator, is swollen down and eventually handled as oral intake of Mn. This is the reason why inhalation of welding fumes has a much greater potential of causing neurotoxic effects than inhalation of coarser dust particulate. It is not surprising that the health-based preventive measures of ACGIH recommend different TWA-TLVs for Mn inhalation, i.e. 20 μg/m3 for respirable and 200 μg/m3 for inhalable particulate. Chronic overexposure to Mn in the occupational setting may lead to accumulation of this metal in the brain which can be evidenced by MRI and may entail neurotoxic effects as illustrated by the occupational exposure contributions in this symposium. Functional neuroimaging combined with neuropsychological tests showed the potential to elucidate processes of cognitive deficits. Furthermore, neuropsychological cognitive test scores showed significant improvement after Mn exposure ceased or decreased, whereas deficits in most of the motor test scores did not. HEC precision deficit has been shown to be completely reversible only when previous cumulative exposure did not exceed “some critical level”. This suggests differential intrinsic vulnerabilities of the brain loci involved with Mn neurotoxicity.
Despite the evidence of inconsistent findings as to the relationship between whole B-Mn and exposure to Mn, B-Mn is still used as biomarker of occupational Mn exposure. It remains still to be elucidated whether it reflects current, recent or long-term exposure. On a group basis, B-Mn is useful to document internal exposure, however, on the individual basis only P-Mn seems to be a promising biomarker of exposure to Mn in welders. As a biomarker of environmental exposure to Mn, either via air or drinking water, B-Mn also showed unusual characteristics. A recent epidemiological investigation comparing adult residents from Marietta (OH), a town with elevated air-Mn (on average 0.18 μg/m3) due to industrial emissions, and Mount Vernon (OH), a comparison town, showed similar mean B-Mn concentrations, i.e. 9.65 and 9.48 μg/L, respectively. No differences in neuropsychological test scores were found between the two towns. It remains an open question whether the associations between the cumulative exposure index (based on air-Mn) and several neuropsychological outcomes, particularly generalized anxiety, are due to direct neurotoxic effects of air-Mn or to the concern/fear of the Marietta residents about the potential health effects of air pollution.
The last two contributions of this symposium deal with cognitive performance in six studies on school-aged children and one study on postnatal neurodevelopment in relation to environmental Mn exposure. Increased air-Mn was shown in two children studies, one from Mexico (PM10, median 0.13 μg/m3) and one from Brazil (PM2.5, median 0.11 μg/m3). Adjusted Full Scale IQ was significantly and inversely associated with H-Mn, but not with B-Mn. The other four studies investigated children whose drinking water was contaminated with Mn, one from China (W-Mn: 241–346 μg/L), two from Bangladesh [W-Mn: mean 795 μg/L (4-3908) and mean 726 μg/L (40–3442)], and one from Quebec [(W-Mn: mean 97 μg/L (1–2700)]. Although these studies showed inverse relation between cognitive performance and W-Mn and/or biomarkers of exposure, they are difficult to compare as the cognitive tests used differed and the relationships with regard to the biomarker of exposure (either B-Mn or H-Mn) were inconsistent. An analysis of three children studies combined (Mexico, Brazil, Quebec) was possible as the studies used H-Mn as a biomarker of Mn exposure and had a measure of Full IQ. The study outcome showed an overall decrease in Full IQ of 2.62 points for a 10-fold increase in H-Mn and a significant (p = 0.04) interaction term H-Mn × sex with a greater loss of Full IQ in girls compared to boys. Taken together, these studies suggest a negative impact of excess environmental Mn exposure on children’s cognitive development, however, consistency as to the association with variables reflecting internal Mn dose needs confirmation. The contribution dealing with early childhood Mn exposure effects on neurodevelopment fills a gap in our current knowledge. The period following fetal life is characterized by rapid brain development and the initiation of the response to sensory input. That the early postnatal period may be critical as to Mn exposure is a valid research hypothesis. However, one should bear in mind that the fetus’ environmental exposure to Mn occurs in the womb of the mother and is subjected to physiological demands of Mn for fetal and early postnatal development, e.g. skeleton and brain. Whole B-Mn is increased in pregnant women and the newborn, but the question remains whether the physiologically active fraction of Mn in the blood in the fetal stage is the Mn in the red blood cells, representing the bulk of the blood Mn, or a free circulating form of Mn transiently elevated because of the pregnancy status. Nevertheless, this prospective study showed an inverted U-shaped association between 12-month B-Mn and concurrent mental development, suggesting that both low and high B-Mn levels may have adverse neurological effects.
Environmental exposure to airborne Mn in adults did not seem to reach a sufficient level to entail a critical concentration of Mn accumulation in the brain causing neurotoxic effects. Recently developed pharmacokinetic models for non-human primates and humans (Schroeter et al., 2011
) seem to indicate that long-term exposure to respirable particulate above 20 μg Mn/m3
is expected to lead to Mn accumulation in the globus pallidus. It is thus clear that for airborne Mn only occupational exposures may exceed such an exposure threshold, whereas in the environmental air-Mn exposure studies the air-Mn levels are at least two orders of magnitude lower. With respect to the children’s exposure to Mn in drinking water, a daily consumption of 1 L would likely constitute a health risk in those individuals drinking water containing more than 2000 μg Mn/L. According to the Toxicological Profile of Manganese (ATSDR, 2008
), the Estimated Safe and Adequate Daily Intake for Mn in children of 4–10 years old has been estimated 1–2 mg/day. In none of the drinking water studies does W-Mn exceed the average concentration of 1 mg/L and in the Quebec study the mean W-Mn did not reach even 0.1 mg/L. With regard to the W-Mn studies in children it remains an open question to what extent potential confounders (e.g. demographic variables) or effect modifying factors (e.g. Fe-deficiency, possible greater GI absorption rate for W-Mn) still influence the outcome of the multiple regression models. Finally, with regard to H-Mn, a prerequisite for its usefulness as a biomarker of W-Mn or air-Mn exposure is the demonstration of unequivocal toxic-kinetic relationships between the external exposure, the Mn dose in the organism (internal exposure) and the excretion rate of Mn with type and growth of hair.