Confounding variables in the environmental toxicology of arsenic

Toxicology 144 (2000) 155 – 162 www.elsevier.com/locate/toxicol

Confounding variables in the environmental toxicology of arsenic T. Gebel * Medical Institute of General Hygiene and En6ironmental Health, Uni6ersity of Goettingen, Windausweg 2, D-37073 Goettingen, Germany

Abstract Arsenic is one of the most important global environmental toxicants. For example, in regions of West Bengal and Inner Mongolia, more than 100 000 persons are chronically exposed to well water often strongly contaminated with As. Unfortunately, a toxicologically safe risk assessment and standard setting, especially for long-term and low-dose exposures to arsenic, is not possible. One reason is that the key mechanism of arsenic’s tumorigenicity still is not elucidated. Experimental data indicate that either DNA repair inhibition or DNA methylation status alteration may be causal explanations. Moreover, when comparing epidemiological data, it cannot be ruled out that the susceptibility to arsenic’s carcinogenicity may be different between Mexican and Taiwanese people. Some other studies indicate that some Andean populations do not develop skin cancer after long-term exposure to As. It is not known yet how this resistance could be mediated. Finally, the situation is even more complicated when taking into consideration that there are several compounds suspected to modulate the chronic environmental toxicity of arsenic, variables that may either enhance or suppress the in vivo genotoxicity and carcinogenicity of the metalloid. Among them are nutritional factors like selenium and zinc as well as drinking water co-contaminants like antimony. Further, yet unidentified factors influencing the body burden and/or the excretion of arsenic are possibly prevailing: preliminary data from own human biomonitoring studies showed a peaking of As in urine samples of non-exposed people which was not caused by elevated exposure to As through seafood consumption. The relevance of these putative confounding variables cannot be finally evaluated yet. Further experimental as well as epidemiological studies are needed to answer these questions. This would help to conduct a toxicologically improved risk assessment, especially for low-dose and long-term exposures to arsenic. © 2000 Elsevier Science Ireland Ltd. All rights reserved. Keywords: Arsenic; Susceptibility; Human biomonitoring

1. Introduction Worldwide, the main reason for a chronic human intoxication with arsenic is the intake of * Tel.: +49-551-394973; fax: + 49-551-394957. E-mail address: [email protected] (T. Gebel)

contaminated drinking water. This contamination originates in As-rich ores like, e.g. pyrite minerals, from which inorganic As leaches into ground and surface water. This problem mainly arose and still arises in developing countries where chemical well water analyses are not be performed periodically. For example, in regions of West Bengal (India)

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and Inner Mongolia (China), more than 100 000 persons are exposed to well water in part highly contaminated with As (Das et al., 1995; Zhang and Chen, 1997). When comparing these exposed people to non-exposed, the internal doses of, e.g. urinary As are elevated from several mg to several hundred mg/l (Lerda, 1994; Das et al., 1995; Hopenhayn-Rich et al., 1996; Aposhian et al., 1997). After long-term exposure to As, increased incidences of diseases mediated by As are the consequence in these populations. The induction of skin and some internal cancers as well as noncarcinogenic effects such as vascular disorders, peripheral neuropathy and diabetes are documented (for review see de Wolff and Edelbroek, 1994). Because of several gaps in knowledge, it is hardly possible to assess the level of exposure at which the chronic toxicity of As becomes epidemiologically relevant. One reason is that in spite of numerous efforts and studies the biochemical key mechanisms of arsenic’s long-term toxicity have not been elucidated. Moreover, when comparing epidemiological studies performed in Mexico and Taiwan, it was hypothesized that the susceptibility to arsenic’s carcinogenicity may be different between the two populations (Dieter, 1991). Additionally, in the Andes, there is evidence that some ethnic groups may have developed skin cancer resistance against As (Vahter et al., 1995; Aposhian et al., 1997). In these regions, high exposures to As are likely for more than 2000 years (Hopenhayn-Rich et al., 1996). The resistance to As may have been selected in these regions because it offered a biological advantage in the case of Chagas disease: the protozoa responsible for the induction of Chagas, Trypanosoma cruzi, are very sensitive to the metalloid. In fact, Chagas disease can be treated with arsenical drugs. However, possible mechanistic causes for the putative resistance and the differences in susceptibility are still unknown. Genetic reasons like enzyme polymorphisms will have to be investigated. As well, simultaneous exposures to further drinking water contents or nutritional factors modulating the effect of arsenic will have to be assessed.

Arsenic, especially in its trivalent chemical species, is a highly bioreactive metalloid covalently binding to sulfhydryl moieties, most preferentially to vicinal dithiols (Basinger and Jones, 1981). As a consequence, in laboratory experiments As was shown to react with and sometimes even inactivate a great number of enzymes and proteins containing thiol moieties (for review see Gebel, 1997). Furthermore, As inhibits enzymes through reaction with dihydrolipoyl groups (for review see Aposhian, 1989). Because of this high reactivity, a wide variety of different biological effects mediated by As have been reported (for review, e.g. see de Wolff and Edelbroek, 1994). Not all of the theoretically possible mechanisms by which the long-term toxicity of As might be modified can be addressed in this paper. The focus shall be drawn on factors on which there is additional epidemiological data to the in vitro studies. The aim is to gather the information available concerning putative confounding variables and possibly relevant modulating factors in the chronic environmental toxicity of As and to point out the gaps in knowledge. In general, several modes of action may be relevant by which arsenic’s action may be modified: first, the rate of enteral absorption of As may be influenced. Second, metabolic methylation of inorganic As may either be suppressed or enhanced. Third, interferences with the process by which As damages DNA may be of importance. Fourth, the absence of protective factors could enhance arsenic’s toxicity. Finally, As could be sequestered and/or inactivated by reaction with certain compounds.

2. Selenium: metabolism and sequestration Nutrition is the main source of selenium exposure of the general population (Kenyon et al., 1997). The contents of the element in drinking water are low. Selenium is well known to protect against arsenic’s toxicity in vitro and in vivo, and vice versa (for review see Kraus and Ganther, 1989; Vadhanakavit et al., 1993; Kenyon et al., 1997). Moreover, in vitro and in vivo selenite was shown to suppress the methylation detoxification

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of As (Styblo et al., 1996; Hall et al., 1997; Kenyon et al., 1997). With respect to acute toxicity and genotoxicity, the non-methylated inorganic As(V) and As(III) are considered to be more toxic than the methylated forms by some orders of magnitude (Moore et al., 1997). Considering this fact, selenite should enhance arsenic’s genotoxicity by methylation inhibition. However, experimental data show the opposite. For example, in human peripheral lymphocytes, Na2SeO3 (sodium selenite) lead to a significant suppression of chromosomal aberrations and sister chromatid exchanges induced by NaAsO2 (sodium arsenite) (Beckman and Nordenson, 1986). In another study, Na2SeO3 (5.6 mg/kg b.w.) was orally administered to mice one hour before, simultaneously and 1 h after NaAsO2 application (2.5 mg/kg b.w.) (Biswas et al., 1999). The As-induced chromosomal aberrations in bone marrow cells were significantly suppressed when Na2SeO3 was applied before, less when given simultaneously, and not significant when applied 1 h after NaAsO2. Two recent studies indicate that arsenic’s genotoxicity may be mechanistically conferred by modifying the cellular methyl pool because of extensive SAM (S-adenosylmethionine) consumption in the metabolism of As (Mass and Wang, 1997; Zhao et al., 1997). After long-term exposure of human A549 cells to As, changes in the DNA methylation status were detected in the promotor region of the p53 gene (Mass and Wang, 1997). An altered expression pattern, i.e. induction of oncogenes and/or suppression of tumor suppressor genes, could be the consequence of this mode of action. If this hypothesis turns out to be crucial for arsenic’s genotoxicity and carcinogenicity, it would become mechanistically plausible that selenite suppresses the toxicity of As through methylation inhibition. However, like As, selenium itself is metabolically methylated, which may lead to changes in the cellular SAM pool. In contrast to As, selenium is considered to be anticarcinogenic (for review see Schrauzer, 1992). A further hypothesis able to explain the protective effect of selenium was offered by one recent study which showed that the element is able to sequester As into lysosomes via precipitation of As2Se (Berry and Galle, 1994). Additional studies

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are necessary to determine the biological significance of this finding. Despite the somewhat contradictory data described above, it can be concluded that selenite suppresses arsenic’s toxicity and should be co-analyzed in the case of elevated exposure to As in human biomonitoring studies. This would offer the ability to assess the putative modulating impact of selenium on the chronic diseases mediated by As.

3. Zinc deficiency and blackfoot disease Blackfoot disease, which is a syndrome associated with gangrene of the feet, is a peripheral vascular disorder endemic in some Taiwanese regions with a formerly high content of As in the drinking water (for review see Engel et al., 1994). Furthermore, in German vintners consuming their ‘house wine’ which was contaminated with arsenical pesticides there was evidence for the induction of gangrenic lesions of the limbs. In contrast, a similar syndrome was not detected after long-term exposure to As in Latin American populations, although durations and levels of exposure were comparable to these in Taiwan and Germany (Zaldivar, 1974; Zaldivar and Guillier, 1977; Wu et al., 1989; Dieter, 1991). However, the typical signs of arsenical skin disease were commonly found in all the three populations. Recently, nutritional zinc deficiency was supposed as possible enhancer of the vascular effects mediated by As. This deficiency could have been caused in Taiwan by malnutrition and by excessive alcohol consumption in the German vintners, respectively (Engel et al., 1994). However, as this hypothesis has not been proved yet, it is still possible that the differing susceptibilities to gangrenic disorders may have other mechanistic reason.

4. Antimony: metabolism and genotoxicity Arsenic and antimony can be found as environmental co-contaminants resulting in environmental and occupational coexposure to man (Crecelius et al., 1974; Atadshanov et al., 1982;

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Riss et al., 1990; Li and Thornton, 1993; Gebel et al., 1995). For example, ores like fahlerze containing As can be co-contaminated with antimony (Gebel et al., 1995), copper ores can contain both arsenic and antimony as contaminants (Crecelius et al., 1974). In most regions where As was found in elevated environmental amounts, it has not been investigated whether an additional coexposure to antimony was present. A coexposure to antimony may be relevant because the element is known to be able to modify arsenic’s action. It was shown in experiments with rat liver cytosol that 5 mM SbCl3 completely inhibited the metabolic methylation of As (Buchet and Lauwerys, 1985; Bailly et al., 1991). Furthermore, it could be demonstrated in own investigations that trivalent antimony is able to suppress the arsenicinduced induction of sister chromatid exchanges and micronuclei (Gebel et al., 1997; Gebel, 1998). The health consequences of a long-term, simultaneously elevated internal exposure to antimony and arsenic to man have not been investigated yet. With respect to the data described above, it is not known whether antimony may enhance or suppress arsenic’s in vivo toxicity (Gebel, 1999). Further investigations assessing these coexposures and their putative in vivo effects are necessary.

in human liver samples could not be detected in contrast to liver samples of several animal species. In the human samples, As methylation seemed to be mediated non-enzymatically (Zakharyan and Aposhian, 1999). Contradictorily, enzymatic As methylation was found in Chang human hepatocytes by the same authors (Zakharyan et al., 1999). Apart from putative metabolic polymorphisms, varying cancer susceptibilities may also be caused by an adapted cellular tolerance to As. Namely, it is possible that skin cancer resistance to As could be mediated by membrane pumps, i.e. the MRP proteins (multidrug resistance associated proteins) belonging to the family of the ABC transporters. The gain of an adapted cellular tolerance to arsenicals and antimonials was documented for bacteria (E. coli ), protozoa (Leishmania) and mammalian V79 cells (Rosen et al., 1988; Mukhopadhyay et al., 1996; Wang et al., 1996). On the molecular basis, this resistance is mediated by ATP-dependent membrane pumps (Rosen et al., 1988; Mukhopadhyay et al., 1996). In mammalian cells, MRP transporters are likely candidates to confer this resistance (Cole et al., 1994; Ishikawa et al., 1996).

6. Seafood and other 5. Susceptibility: metabolism and tolerance Possible genetic differences in the susceptibility to As can only be discussed in short, because research on this topic is still in its infancy. Apart from seafood (see below), the main environmental source of As is inorganic, pentavalent As(V). In mammals As(V) is metabolized via reduction and glutathione (GSH) conjugation to As(III), which subsequently is mono- and dimethylated (Thompson, 1993). This process can be considered as detoxication (Moore et al., 1994; Vahter, 1999). There is some recent evidence that human GST polymorphisms may indirectly influence the extent of As methylation (Chiou et al., 1997). Little is known about As-methylating enzyme polymorphisms in man because these enzymes have not been identified yet. There is one study reporting that methylating enzyme activity

Although seafood is not a variable which modulates arsenic’s toxicity itself, it is an important factor of influence on the body burden with As and has to be discussed in short here. Seafood, but not freshwater fish, contains comparatively high amounts of As. A high portion of the seafood-borne As is either bound to proteins and sugars or can be found as arsenobetaine and arsenocholine (Phillips, 1990; Buchet et al., 1994). Arsenobetaine is the major source of As in seafood. After ingestion and absorption, all these seafood-borne As compounds release inorganic arsenic in only minute amounts, thus, their toxic potentials are not significant. However, few percent of the seafood-borne As is dimethylarsinic acid (DMA) which has a low yet not irrelevant toxicological significance. The highly toxic inorganic arsenicals, i.e. As(V) and As(III), are pre-

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vailing in even minor portions in fish and shellfish. When performing biomonitoring studies, the hydride atomic absorption spectrometry (AAS) is the analytical method recommended to determine the As species of toxicological relevance to man, i.e. As(V), As(III), MMA, and DMA. Organoarsenicals and trimethylated As compounds in seafood are not detected with this technique (Crecelius, 1977; Buchet et al., 1980). However, consumption of marine fish or shellfish leads to a short-term increase in the toxicologically relevant As detected by hydride AAS, which is mainly caused by DMA. Own investigations showed that a possible further yet unidentified confounder on the internal doses of As may exist: four of in all 300 German subjects, who participated in a human biomonitoring study, showed a strong peaking of As in 24-h-urine. The As levels of these four subjects were found lying above 40 mg As/l, the maximum value was 132 mg As/l urine (Table 1). Long-term exposures resulting in As urine levels above 40 mg As/l are judged as toxicologically critical (Ewers and Brockhaus, 1987). Three further subjects showed urine levels above 15 mg As/l, in which case a re-check is recommended by the German national health agency. This shall serve to identify potential sources of elevated exposure and to exclude a long-term exposure increase (Ewers and Brockhaus, 1987). Table 1 Peaking As levels in 24-h-urine samples of volunteersa Proband

1 2 3 4 5 6 7 a

Arsenic in urine (mg As/l) 1st sampling

2nd sampling

3rd sampling

24.2 15.3 20.6 132.0 16.1 47.6 17.5

10.8 75.9 41.1 2.0 19.7 10.9 9.7

– 5.7 11.6 – 10.8 – –

The analytical determination was performed with atomic absorption spectrometry/hydride technique. Elevated levels (\ 15 mg As/l urine) are given in bold letters.

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In all, two rechecks were performed with intervals of several weeks till the resampled urine of each of the seven probands showed arsenic levels below 15 mg As/l (Table 1). Because of the quite short biological half-life of As in man of 3–4 days (Tam et al., 1979; Buchet et al., 1980; Pomroy et al., 1980), it is not known yet whether the peaking of urinary As was short-term or long-term. Such peaking of urinary As was documented by others (e.g. Vahter and Lind, 1986; Schmid et al., 1996; Trepka et al., 1996) and was suggested to be caused by enhanced consumption of seafood. Indeed, several studies showed that seafood consumption is associated to an increase in urinary As detected with the AAS hydride technique (e.g. Vahter and Lind, 1986; Polissar et al., 1990; Gebel et al., 1998). Remarkably, our questionnaire evaluation revealed that the seven subjects described in Table 1 by us were low consumers of seafood. One of the seven subjects had eaten fish the day before the first urine sampling (subject no. 1 in the Table 1), the other six had not eaten seafood for the last 10 days before each sampling. It is known that urinary As returns to its basal level 3 days after the consumption of a seafood meal (Arbouine and Wilson, 1992). Thus, the peaking of urinary As could not have been caused by seafood in six of these seven subjects. Elevated occupational exposures to As could be excluded because of the questionnaire data. As well, an elevated intake of As via drinking water or mineral water could not be made responsible for the peaking because the threshold limits in Germany are low (10 and 50 mg As/l, respectively) and well surveyed. There may be other, presumably dietary, sources of As, which need to be identified. Unfortunately, there is no current nutrition source known leading to such a marked increase in the exposure to toxicologically relevant As. Another explanation could be that the intake of food containing arsenic-chelating/-binding compounds may influence the enteral absorption and/or the excretion rate of the metalloid. In an extensive human biomonitoring study carried out with volunteers of the general German population, depending on the age group evaluated 15–16% of in all 4731 urine samples showed As levels higher than 15 mg As/l. As mentioned

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above, this level of urinary As can be judged as upper threshold of the non-elevated internal dose range (Ewers and Brockhaus, 1987). In that study, elevated internal doses of cadmium, chromium, copper and mercury were found far less often in frequencies ranging from 1 to 3.5%, respectively (Krause et al., 1996). We believe that these data as well as our preliminary results justify the need to carry out further studies to investigate the putative existence of a hitherto unidentified confounder for the internal doses of As in the general population.

7. Conclusion The epidemiological data, mainly collected from the extensive epidemiological studies in Taiwan, suggest that there may be no elevated tumor risk consuming drinking water with As levels below 100 mg As/l (Brown and Chen, 1995). Experimental data as well indicate that arsenic’s carcinogenicity may not be induced linearly but underlie a sublinear dose-response (Rudel et al., 1996). By the help of additional studies, it is quite possible that a toxicological re-evaluation in the future may define an exposure level below which the carcinogenic action of As is not relevant. This would mean that toxicologically based ‘threshold’ levels could be defined although As is classified as human carcinogen. However, before such re-evaluation can be performed, it will be necessary to carry out additional studies to assess the relevance of enzyme polymorphisms in the metabolism of As and to investigate the human in vivo adaptation of cellular As tolerance. Furthermore, in case of elevated exposure to arsenicals it seems recommended to simultaneously monitor each of the putative confounding variables which were addressed in this paper. Environmental monitoring can be difficult in the case of analysis of alimentary factors because nutrition duplicate studies have to be carried out; these studies are of high analytical expenditure. In general, it seems more feasible to perform a human biomonitoring which offers the advantage to assess the internal doses to all the respective compounds on an individual basis.

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