Keywords

Introduction

For few decades, the environmental pollution has become the issue of focus among scientists, academicians, environmentalists, and regulatory authorities. There are many kinds of pollutants; some organic contaminants are most dangerous due to their chronic toxicities and carcinogenic nature. The most commonly found organic pollutants are pesticides, phenols, plasticizers, polychlorinated, and polycyclic aromatic hydrocarbons (PAHs) [1–4]. The presence of such pollutants in our earth ecosystem is dangerous and many methods have been reported for their monitoring. It is very important to highlight that the reported methods provide the total concentrations of pollutants but they do not distinguish which mirror images of pollutants are present and which are harmful in the case of chiral contaminants. Therefore, during the last few years, scientists and regulatory authorities are in the demand of the data on concentrations and toxicities of the chiral pollutant mirror images. This is an essential, urgent, and demanding field in the present century. For the preparation of this chapter, we have searched the literature thoroughly and observed that only few groups are working in this area. This chapter describes the role of chirality in the environment, especially the distribution and toxicities of chiral pollutants.

Origin of the Chirality

Before discussing the distribution and toxicities of chiral pollutants, it is essential to discuss some aspects of chirality so that readers can realize the importance of chirality in the environment. Basically, chirality has been derived from the Greek word kheir for handedness and, first of all, used by Lord Kelvin in 1883 [5]. Any molecule deprived of plane, center, and axis of symmetry exists in more than one form; these are called chiral objects or enantiomers; enantiomers that are nonsuperimposable mirror images of each other are called as chiral objects. The role of chirality is very important in our lives and still has not been fully explored. Chirality is found in a wide range of objects starting from elementary constituents of our body structure [6]. There are several examples of the chirality in our environment, i.e., burial chamber mural paintings in Egypt [5], 540 galaxies listed in Carnegie Atlas of Galaxies [7], and helical structures of plants and animals. Briefly, the chirality exists almost everywhere in this universe and is associated with the origin of the earth and life [8].

Chemistry of the Chirality

In 1809, Haüy [9] evolved chemical utility of the chirality, who postulated that, from crystal cleavage observations, a crystal and each constituent space-filling molecules are images of each other in overall shape. In 1848, Pasteur described the different destruction rates of dextro and levo ammonium tartrate by the mould Penicillium glaucum [10]. The tetrahedral arrangement of carbon with different four groups makes the whole pollutant asymmetric in structure, and such pollutant differs in three dimensional configurations and exists in two forms, which are mirror images of each other (Fig. 1). These mirror images are called optical isomers, stereoisomers, enantiomers, enantiomorphs, antipodes, or chiral molecules. The phenomenon of the existence of the enantiomers is called as stereoisomerism or chirality. The 50:50 ratios of the enantiomers are called racemic mixture, which do not rotate the plane-polarized light. The number of the enantiomers may be calculated by 2n, where n is the number of the chiral centers within the respective molecule. In the beginning, the optical isomers were distinguished with (+) and (−) signs or d (dextro) and l (levo), indicating the direction in which the enantiomers rotate the plane-polarized light. (+) or d stands for a rotation to the right (clockwise), whereas (−) or l indicates a rotation to the left (anticlockwise). The main drawback of such an assignment is that one cannot derive the number of chiral centers from it. This is possible when applying the well-known R/S notation given by Cahn and Ingold, which describes the absolute configuration (the spatial arrangement of the substituents) around the asymmetric carbon atom of the pollutant (molecule).

Fig. 1
figure 1

Enantiomers of o,p′-DDT (1,1,1-trichloro-2-(o-chlorophenyl)-2-(p′-chlorophenyl)ethane)

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Environmental Aspects of the Chirality

The most notorious environmental pollutants are pesticides, which are about 25% chiral molecules [11]. Polyaromatic hydrocarbons may also be chiral pollutants. It has been observed that one of the enantiomers of the chiral pollutant may be more toxic, and, hence, both enantiomers may have different toxicities [5, 12]. This is an important information to scientists when performing environmental analyses. Biological transformation of the chiral pollutants can be stereoselective; uptake, metabolism, and excretion of enantiomers may be very different [12, 13]. Therefore, the enantiomeric composition of the chiral pollutants may be changed through these processes. Metabolites of the chiral compounds are often chiral as well. Moreover, some of the achiral pollutants degrade into the chiral metabolites. For example, γ-hexachlorocyclohexane (γ-HCH) and atrazine degrade into γ-pentachlorocyclohexene (γ-PCCH) and 2-chloro-4-ethylamino-6-(1-hydroxy-2-methylethyl-2-amino)-1,3,5-triazine racemic mixtures, respectively. It has also been reported that the enantiomers may react at different rates with achiral molecules in the presence of chiral catalysts [4]. Since constituents of living organisms are usually chiral, there are greater chances of the chiral pollutants to react at different rates. To predict the exact chiral pollution load determination of enantioselective toxicities and concentrations of the enantiomers is thus required and an essential need.

Distribution of the Chiral Pollutants in the Environment

Both point and nonpoint sources are major sources for pollution in the environment. The most commonly found chiral pollutants are given in Table 1. These compounds are widely distributed in our ecosystem. Marine water has been reported as polluted due to heptachlor exo-epoxide (a metabolite of heptachlor); α-, β-, and γ-HCH; toxaphene; and phenoxyalkanoic acid herbicides. There are only few papers published [5] on the ground water contamination by pesticides and other toxic organic pollutants [3]. Weigel [14] reported on the presence of several drugs in waste water. Buser et al. [15] identified ibuprofen, a nonsteroidal anti-inflammatory drug, in waste and river waters. Recently, Ali et al. [16] reviewed the literature on the distribution of drugs in the environment.

Table 1 List of some common chiral pollutants [4]
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Vetter et al. [17] determined toxaphene in Canadian lake sediments and chloroborane congeners in the sediment from the toxaphene-treated Yukon lake [18]. Rappe et al. [19] reported the presence of chiral pesticides in the sediment of the Baltic Sea, and Benicka et al. [20] identified PCBs in sediments of a river. Wong et al. [21] looked into the enantiomeric ratios of eight PCB species in the sediments from selected sites in the USA. Biselli et al. [22] carried out a comprehensive study on the distribution of the chiral musks in sediments of various waste water plants. Moisey et al. [23] determined the concentrations of α-, β-, and γ-HCH isomers and enantiomers in sediments obtained from the North Sea. Aigner et al. [24] described the enantiomeric ratio of the pesticide chlordane in the soil of Midwestern USA. The pesticides detected in these samples were chlordane, heptachlor, and heptachlor exo-epoxide. Wiberg et al. [25] described organochlorine pesticides in 32 agricultural and 3 cemetery soils from Alabama. Lewis et al. [11] detected the dichlorprop pesticide in Brazilian soils, and White et al. [26] identified cis- and trans-chlordanes in the soil of a green house unit.

Besides water, soil, and sediments, chiral pollutants have also been detected in the atmosphere. The concentrations of chiral pollutants being found varied from place to place. Ridal et al. [27] detected α-HCH in air above the water surface of Ontario Lake. Ulrich and Hites [28] reported the presence of chlordane in air samples near the Great Lakes. Aigner et al. [24] reported enantiomeric ratios of the chlordane pesticide in the air of Midwestern USA. Similarly, Bidleman et al. [29] collected air samples from Corn Belt, South Carolina, and Alabama areas. The authors reported the presence of cis-chlordane, trans-chlordane, heptachlor, and heptachlor exo-epoxide in these samples. Other authors who described the presence of chiral pesticides in air samples are Wiberg et al. [30] (chlordane) and Buser and Müller [31] (heptachlor and chlordane).

Moreover, chiral xenobiotics have been routinely identified in earth’s biota. Different chiral ratios of different pollutants were detected in various organs of seals, Eider ducks, polar bears, whales, pelagic zooplankton, arctic cod, sea birds, fishes, bivalves, crayfish, water snakes, barn swallows, sheep, roe deer, and even humans [32–40]. For a ready reference, Tables 2 and 3 describe the distribution of chiral pollutants in water and biota of our ecosystem. The different chiral ratios of toxaphene, cis- and trans-chlordane, and heptachlor exo-epoxide in air, sediment, soil, and plants are given in Tables 4–7.

Table 2 Chiral pollutants in various water resources [4]
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Table 3 Chiral pollutants in different biota [4]
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Table 4 Chiral ratio of toxaphene congeners b7-1001 and b7-923 in sediments from the Canadian Hanson Lake [4]
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Table 5 Chiral ratio of cis- and trans-chlordane pesticides in soil samples [4]
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Table 6 Chiral ratio of chlordane, heptachlor, and heptachlor exo-epoxide in air from the Corn Belt region, South Carolina, and Alabama [4]
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Table 7 Chiral ratio of cis- and trans-chlordane in different plants [4]
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Toxicities of Chiral Pollutants

Only little information is available on the enantioselective toxicity of pollutants. Basically, differences in the bioaffinity of the enantiomers to a binding site on an enzyme or receptor surface are responsible for different toxicities. Such differences may reveal in terms of distribution rates, compound’s metabolism, and excretion; in antagonistic actions relative to each other; or in their individually different tissue-specific toxicological properties. The enantioselective toxicities of chiral pollutants are discussed in the following sections.

Enantioselective Toxicities of Pesticides

Möller et al. [41] described the different carcinogenic potencies and growth stimulation of α-HCH enantiomers in primary rat hepatocytes by reporting 100% cell death in the presence of 3.0 × 10−4 M (+)-α-HCH. Contrarily, (−)-α-HCH only induced 75% toxicity at the same concentration. By using concentrations of 5.0 × 10−5 M of both enantiomers, significant increases in mitosis occurred in the presence of the (+)-α-HCH enantiomer (factor 2.4) as compared with a 1.7-fold stimulation by (−)-α-HCH enantiomer. Concentration-dependent cell death rates observed in primary cultures of rat hepatocytes treated with (+)- or (−)-α-HCH and the corresponding stimulation of mitosis in these cells are depicted in Fig. 2.

Fig. 2
figure 2

Differences in the biological effects of the (+)- and (−)-enantiomers of α-HCH (α-hexachlorocyclohexane). (a) Toxicity in primary cultures of rat hepatocytes and (b) effects on the stimulation of mitosis in these cells (from [41] with reprint permission of Eco-Informa Press)

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Miyazaki et al. [42] studied the enantioselective toxicities of cyclodiene pesticides (e.g., chlordiene, chlordiene epoxide, and heptachlor exo-epoxide) on male German cockroach insects (Blattella germanica). The authors reported that (+)-chlordiene, (−)-chlordiene epoxide, and (+)-heptachlor exo-epoxide enantiomers exhibited stronger toxicity when compared to their corresponding antipodes (Table 8). The toxicity was expressed in percent of dead animals 24 h after the application of the compound. From the results obtained, it can be concluded that (+)-chlordiene, (−)-chlordiene epoxide, and (+)-heptachlor exo-epoxide are more toxic in these insects than their enantiomers. The LD28.6 value of (+)-chlordiene was 129, while LD50 values of (−)-chlordiene epoxide and its racemate were 76 and 157, respectively, indicating the differences of enantioselective toxicities of these insecticides.

Table 8 Chiral toxicities of chlordiene, chlordiene epoxide, and heptachlor exo-epoxide insecticide enantiomers in the German cockroach Blattella germanica (% of dead animals after 24 h) (according to [42])
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Furthermore, Miyazaki et al. [43] reported the different enantioselective toxicities of heptachlor and 2-chloroheptachlor pesticides on the same cockroach species. LD50 values for these pesticides were calculated after 24 h (Table 9). It has been reported that only heptachlor and 2-chloroheptachlor showed toxicities, while 3-chloroheptachlor was nontoxic. LD50 values for (−)-, (+)-, and (±)-heptachlor were 5.32, 3.38, and 2.64, respectively. On the other hand, LD50 values calculated for (−)-, (+)-, and (±)-2-chloroheptachlor were 100, 50, and 20, respectively. Therefore, it may be concluded that the toxicities of the (+)-enantiomers of heptachlor and 2-chloroheptachlor are greater than that of their corresponding (−)-enantiomers. Based on these results, the theoretical LD50 values of the racemic mixtures of heptachlor and 2-chloroheptachlor should be 4.35 and 75.0, respectively. However, the observed LD50 values are lower than the theoretical values (Table 9). Therefore, it may be concluded that the toxic potency of one enantiomer is being increased due to the presence of the other.

Table 9 Chiral toxicities of heptachlor and 2-chloroheptachlor enantiomers in the German cockroach Blattella germanica (% of dead animals after 24 h) (according to [43])
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McBlain and Lewin [44] reported (−)-o,p′-DDT as a more active estrogen-mimic species in rats than the (+)-enantiomer. Hoekstra et al. [45] described a yeast-based assay to assess the enantiomer-specific transcriptional activity of o,p′-DDT via interaction with the human estrogen receptor (hER). While the (−)-enantiomer strongly induced measureable hER activity, the corresponding potency of the (+)-o,p′-DDT was negligible. However, high concentrations of the (+)-enantiomer influenced (decreased) the transcriptional activity of the (−)-o,p′-DDT. The dose-dependent reporter gene (β-galactosidase) activity is shown in Fig. 3.

Fig. 3
figure 3

Differences in enantiomer-specific transcriptional activity of o,p′-DDT via interaction with the human estrogen receptor (hER). Binding affinities of the (+)- and (−)-enantiomers and the racemic mixture were measured via β-galactosidase activity (from [45] with reprint permission of Elsevier)

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Miyazaki et al. [46, 47] reported on the enantioselective differences in the toxicities of methamidophos (O,S-dimethyl phosphoramidothiodate) and acetaphate (O,S-dimethyl-N-acetylphosphoramidothiodate) to houseflies. In houseflies, the (+)-enantiomers are more potent than their (−)-counterparts. By contrast, the (−)-enantiomers were found more toxic to German cockroaches (B. germanica), albeit LD50 values were close for both enantiomers. In addition, the (−)-enantiomer resulting from sulfoxidation of propaphos was found more potently inhibiting cockroaches and―at the biochemical level―the bovine erythrocyte acetylcholinesterase (AChE) when compared with its (+)-enantiomer [47]. Furthermore, the authors studied the toxicities of these two enantiomers on houseflies and green leaf hoppers and reported only little differences in the toxicities in these insects. Phosphor-containing pesticides were introduced in the 1950s for insect control in fruits, vegetables, and other crops.

Lang et al. [48] described the conversion of atrazine into the racemic mixture of 2-chloro-4-ethylamino-6-(1-hydroxy-2-methylethyl-2-amino)-1,3,5-triazine by liver microsomes of rats, pigs, and humans. The authors reported on the dominance of the R-enantiomer in humans, while the higher concentrations of the S-enantiomer were observed in rats and pigs. A species-dependent enantioselective formation of this metabolite with S/R ratios of 76:24 in rats, 49:51 in pigs, and 28:72 in humans was stated. Similarly, trans-nonachlor, a major constituent of technical chlordane, is achiral, and the replacement of chlorine substituents by another atom or group can produce a chiral derivative. Further, malathion usually is biotransformed into racemic malaxon that exhibits anti-AChE (insecticidal) activity. For bovine erythrocyte cholinesterase, the antagonistic activity of the R-enantiomer is 22 times greater than for the S-enantiomer [49, 50].

Enantioselective Toxicities of Polychlorinated Biphenyls

Polychlorinated biphenyls (PCBs) are the most notorious class of chlorinated chiral pollutants. Although their use has been banned in many countries since the 1970s, these compounds still represent an important class of priority pollutants due to their long persistence, bioaccumulation, and toxicity [51]. About 209 PCB congeners are known, and out of them 78 are chiral in nature. Again out of these 78 PCB congeners only 19 form stable enantiomers (atropisomers) [52]. Different toxicities of these chiral PCBs have been described in terms of porphyria, teratogenicity, endocrine and reproductive malfunctions, etc. It has been reported that non-ortho coplanar PCBs exhibit the highest toxicities followed by the moderately toxic mono-ortho coplanar congeners, while di-ortho-substituted PCBs turned out to be less toxic. Ahlborg et al. [53] presented a toxic equivalency model to describe the toxicities of PCB congeners. The authors calculated the toxic equivalency factors (TEFs) for individual PCBs. Each PCB has been assigned a TEF value based on its toxicity relative to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), which has by definition a TEF of 1.00 (100%). Püttman et al. [54] reported PCB139 and PCB197 congeners as effective inducers of drug-metabolizing enzymes (e.g., cytochrome P-450 monooxygenases, CYPs; N-demethylase; and aldrin epoxidase). The authors described the (+)-enantiomer of PCB139 as the stronger inducer in comparison to the (−)-enantiomer. Contrarily, the racemic mixtures of PCB197 and its individual enantiomers are only weak inducers of these enzymes. Furthermore, in 1991, the same group reported activities related to the induction of CYPs, that is, the activity of ethoxyresorufin-O-deethylase (EROD) and benzphetamine-N-demethylase (BPDM) [55]. They demonstrated that EROD activity is induced to much greater extent by (+)-enantiomers of all of the congeners studied with no activities of the (−)-enantiomers of PCB88 and PCB197. The effects of the enantiomers of PCB88, PCB139, and PCB197 on the induction of total CYP enzymes and EROD and BPDM activities are shown in Figs. 4–6, respectively.

Fig. 4
figure 4

Effects of the enantiomers of (a) PCB88, (b) PCB139, and (c) PCB197 congeners on the induction of total cytochrome P-450 enzymes (CYPs); (triangle) (+)-enantiomer, (filled triangle) (−)-enantiomer, and (circle) racemic mixture (from [55] with reprint permission of Elsevier)

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Fig. 5
figure 5

Effects of the enantiomers of (a) PCB88, (b) PCB139, and (c) PCB197 congeners on the induction of ethoxyresorufin-O-deethylase (EROD) activity; (triangle) (+)-enantiomer, (filled triangle) (−)-enantiomer, and (circle) racemic mixture (from [55] with reprint permission of Elsevier)

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Fig. 6
figure 6

Effects of the enantiomers of (a) PCB88, (b) PCB139, and (c) PCB197 congeners on the induction of benzphetamine-N-demethylase (BPDM) activity; (triangle) (+)-enantiomer, (filled triangle) (−)-enantiomer, and (circle) racemic mixture (from [55] with reprint permission of Elsevier)

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The effects of the enantiomers of PCB88, PCB139, and PCB197 on the accumulation of protoporphyrin and uroporphyrin (URO) chick embryo liver cell cultures are summarized in Table 10. The results indicate that URO accumulation occurred only at high concentrations (i.e., ≥1.0 μM) of PCB88 and PCB197, but at low concentrations of the PCB139 congener [≥0.034 μM for (+)-PCB139 and ≥0.34 μM for (−)-PCB139]. The strongest URO accumulation occurred with PCB139, with 64% URO generated by the (+)-enantiomer and 47% URO generated by the (−)-enantiomer at the highest concentration tested.

Table 10 Concentrations of enantiomers of PCB88, PCB139, and PCB197 and induced accumulation of protoporphyrin and uroporphyrin in chick embryo liver cell culture [55]
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PCB methyl sulfones (MeSO2-PCBs) are metabolites of PCBs generated via the mercapturic acid pathway. Cleavage of the sulfur–carbon bond in the cysteine moiety, methylation, and oxidation of the methyl sulfide has been described by Bakke and Gustafsson [56]. These metabolites are more persistent and less hydrophobic than their corresponding parents, which make them long-lasting contaminants of the biosphere. Several MeSO2-PCBs have been shown to strongly induce CYP activity such as CYP2B2, 3A2, and 2C6. A study on the influence of MeSO2-PCBs in the reproduction of minks (Mustela vison) indicated that MeSO2-PCB and 1,1-dichloro-2-(o-chlorophenyl)-2-(p′-chlorophenyl)ethylene (DDE) methyl sulfone mixtures increased the litter size in these animals [57]. A strong respiratory distress and alterations in the immune status of yusho patients in Japan have been related to MeSO2-PCBs [58]. There are several reports describing the toxicities of MeSO2-PCBs, but, unfortunately, no report is available on the enantioselective toxicities of MeSO2-PCBs. Furthermore, a Swedish and German collaborative project started in 1997 with the aim of studying enantioselective accumulation of MeSO2-PCBs in the liver of humans and rats [36, 59], but data on the enantioselective toxicities are still missing.

Enantioselective Toxicities of Polycyclic Aromatic Hydrocarbons

Among the most toxic PAHs in the environment are β-naphthoflavone, benzo[a]pyrene (BP), and anthracene and their derivatives. Toxicities of the racemic metabolites of PAHs are known for a long time, but only few reports are available on their enantioselective cytotoxic, mutagenic, and carcinogenic effects. In 1977, Levin et al. [60] studied the carcinogenic activity of trans-7,8-dihydroxy-7,8-dihydrobenzo[a]pyrene (BP-7,8-dihydrodiol) on CD-1 mouse skin with 50, 100, and 200 nmols of each of the (+)- and (−)-enantiomers. It was observed that the (−)-enantiomer was more toxic than the (+)-enantiomer at all concentrations. The maximum tumor formation was observed after 21 weeks, with a 5–10 times higher carcinogenic activity of the (−)-enantiomer in comparison to the (+)-enantiomer. Furthermore, the same authors described the effect of BP-7,8-dihydrodiol on the skin of newborn mice, concluding the greater toxicity of the (−)-enantiomer. While the parent BP is not a carcinogen by itself, its metabolic products such as diol-epoxides are highly potent [61–66]. Two possible diastereoisomers originating from trans-BP-7,8-dihydrodiol were characterized [67], and one of them was shown to be an ultimate carcinogen in newborn mice [68, 69]. The diastereomer BP-7β,8α-dihydrodiol-9α,10α-epoxide (anti-BPDE) has been characterized as potent mutagen in bacteria and certain mammalian cells [70–72]. Wood et al. [73, 74] reported the (+)-enantiomer [i.e., (+)-BP-7β,8α- diol-9α,10α-epoxide, (+)-anti-BPDE] being four times more toxic in Chinese hamster cells than its (−)-antipode [(−)-anti-BPDE]. Slaga et al. [75] also studied the enantioselective carcinogenesis of anti-BPDE on mouse skin. According to the authors, the carcinogenic potency of the (+)- and (−)-anti-BPDE enantiomers were 60% and 2%, respectively. The results of this study are depicted in Fig. 7.

Fig. 7
figure 7

Skin tumor-initiating activities of (circle) (+)-BP-7β,8α-diol-9α,10α-epoxide [(+)-anti-BPDE], (filled triangle) (−)-BP-7α,8β-diol-9β,10β-epoxide [(−)-anti-BPDE], and (triangle) racemic BP-7,8-diol-9,10-epoxide [(±)-anti-BPDE] in female CD-1 mice. A dose of 200 nmol of each compound was applied. (a) Percentage of papillomas; (b) Papillomas per mouse (from [75] with permission of American Association for Cancer Research―AACR)

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The stereoselective metabolism of BP toward its ultimate carcinogen occurs as follows: BP→BP-7,8-oxide→trans-BP-7,8-dihydrodiol→BP-7,8-diol-9,10-epoxide (BPDE). Chang et al. [76] reported that the (−)-enantiomer of BP-4,5-oxide was 1.5–5.5-fold more mutagenic than the (+)-enantiomer in bacterial strains of Salmonella typhimurium (TA98, TA100, TA1537, TA1538) and in Chinese hamster V79 cells. The authors reported when mixtures of the enantiomers were studied in V79 cells, synergistic cytotoxic and mutagenic responses could be observed. The most cytotoxic and mutagenic effects occurred with a 3:1 mixture of the (−)- and (+)-enantiomers of BP-4,5-oxide. Levin et al. [77] described that the (+)-BP-7,8-oxide showed greater enantioselective toxicity in the skin of newborn mice (see Table 11). It is interesting to note that the tumor formation potencies of BP-7,8-oxide were in the order: racemic mixture > (+)-enantiomer > (−)-enantiomer. The higher toxicity of the racemic mixture might be the result of catalytic interferences between the enantiomers.

Table 11 Chiral tumorigenicity of BP-7,8-oxide enantiomers (according to [77])
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Wood et al. [78] studied the enantioselective toxicities of four isomers of chrysene-1,2-diol-3,4-epoxide in bacterial (histidine dependent) strains of S. typhimurium and in mammalian (Chinese hamster V79) cells. In strain TA98 of S. typhimurium, the (−)-anti-chyrsene-1,2-diol-3,4-epoxide was 5–10 times more toxic compared to the other three isomers. However, in strain TA100 of these bacteria and in Chinese hamster V79 cells, (+)-anti-chyrsene-1,2-diol-3,4-epoxide was the most mutagenic diol-epoxide and about 5–40 times more active than the other three optical isomers. Furthermore, the same group studied the enantioselective toxicities of trans-1,2-dihydroxy-1,2-dihydrochrysene (chrysene-1,2-dihydrodiol) and chrysene-1,2-diol-3,4-epoxides in two different mouse tumor models [79]. In the animals, the skin, pulmonary, and hepatic carcinogenicity of these chiral pollutants was investigated. Skin carcinogenicity is presented in Table 12. Table 13 summarizes the extent of pulmonary and hepatic tumor formation in newborn mice.

Table 12 Chiral tumorigenicity of trans-1,2-dihydroxy-1,2-dihydrochrysene (chrysene-1,2-dihydrodiol) and anti-chrysene-1,2-diol-3,4-epoxide enantiomers in mouse skin (according to [79])
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Table 13 Chiral pulmonary and hepatic carcinogenicity of trans-1,2-dihydroxy-1,2-dihydrochrysene (chrysene-1,2-dihydrodiol) and syn- and anti-chrysene-1,2-diol-3,4-epoxide enantiomers in newborn mice (according to [79])
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Table 12 shows that only 3% tumors were found when (+)-enantiomer of chrysene-1,2-dihydrodiol was injected, while 67% carcinogenicity was observed with the (−)-enantiomer. Contrarily, the tumor-initiating activity of (+)-, (−)-, and (±)-chrysene-1,2-diol-3,4-epoxides (at a dose of 1.2 μmol each) was 21%, 13%, and 25%, respectively. Again, the toxicity of the racemic mixture was greater than the effects seen for (−)- and (+)-enantiomers, which might be due to the interference of both enantiomers with each other. A perusal of Table 13 indicates that again the (−)-trans-chrysene-1,2-dihydrodiol is more toxic than its (+)-antipode in both male and female mice.

Benz[a]anthracene(BA)-3,4-diol-1,2-epoxide (BADE) results from regio- and stereoselective metabolism of BA. The tumorigenic activities of the (+)- and (−)-enantiomers of trans-3,4-dihydroxy-3,4-dihydrobenz[a]anthracene (BA-3,4-dihydrodiol) and racemic diastereomers of BADE were studied in newborn Swiss-Webster mice [80]. Furthermore, Tang et al. determined the tumorigenic potencies of racemic syn- and anti-7,12-dimethylbenz[a]anthracene(DMBA)-3,4-diol-1,2-epoxides (DMBADE) via the two-stage initiation–promotion protocol in mouse skin [81]. They observed that both syn- and anti-DMBADE were active tumor initiators and that the occurrence of papillomas was dependent on the dose of carcinogen applied.

Enantioselective Toxicities of Algae Toxins

Tetrodotoxin and saxitoxin are major marine toxic chiral pollutants. The toxic effects of saxitoxin have been observed in some part of the world, and, hence, sometimes filter-feeding shell fish industries have been affected. It has also been reported that the marine environment is rich with (−)-enantiomer of saxitoxin. Other chiral neurotoxins (anatoxin, homoanatoxin, etc.) are produced in the aquatic environment as well [82]. Therefore, sometimes, water becomes toxic due to the presence of these compounds, and several reports have been published on the death of cattle and dogs due to intoxication [83, 84]. Accordingly, the presence of these toxic chiral pollutants may be health threatening for humans as well. Unfortunately, no reports have been published on the enantioselective toxicities of these toxins yet.

Enantioselective Toxicities of Drugs and Pharmaceuticals

A myriad of different drugs are being used by people, and among them a great part is chiral in nature. Therefore, the presence of such types of drug enantiomers in the environment may be problematic and hazardous. Weigel [14] reported on the presence of several drugs in aquatic environments at high concentrations. Kümmerer [85] reported the presence of several drugs in surface, ground, and drinking water. In 1960, thalidomide [(R,S)-N-(2,6-dioxo-3-piperidyl) phthalimide] was introduced as a sedative drug in Europe, and, unfortunately, teratogenic effects of this drug occurred in embryos due to the highly toxic S-enantiomer [86]. Ifosfamide is a cyclophosphamide analog, which possesses toxicity that is enantioselective in nature. Masurel et al. [87] studied the enantioselective toxicity of this drug in rats. The authors injected the racemate and the enantiomers separately into nontumor-bearing rats at doses of 550–650 mg/kg. The mean weight loss (at highest dose) was 30%, 20%, and 17% for the (+)- and (−)-enantiomers and the racemic mixture of ifosfamide, respectively. Furthermore, the authors observed signs of acute bladder toxicity, as blood was reported in the urine of rats when (−)-ifosfamide was injected. Similarly, there are several other drugs whose enantiomers are selectively toxic. l-DOPA has long been introduced for the treatment of Parkinson’s disease, albeit d-DOPA turned out to be toxic [88, 89]. In 1986, Domino [90] described enantioselective opioid hallucinogen interactions of N,N-dimethyltryptamine and lysergic acid N,N-diethylamide in rats.

Conclusions

It is clear from this chapter that the chirality plays an important role in environmental toxicology affecting our lives. The chiral pollutants are widely distributed. Enantioselective toxicities have been reported for several xenobiotics in the earth ecosystem. In spite of this, scientists have only rarely been attracted toward this problem, and only few groups are addressing this issue. Therefore, there is still an urgent need to more comprehensively explore the enantioselective toxicities of chiral pollutants. The existing toxicological data of pollutants, mostly pertaining to their racemic forms, must be refined in terms of enantioselective toxicities. Even achiral pollutants are sometimes metabolized into chiral follow-up products, and, therefore, the study of these chiral species is a demanding field. Analysis of the specific toxicities of the chiral pollutants is essential and may be useful for controlling certain adverse health effects and diseases. In summary, the role of chirality in environmental toxicology is a burning area and needs more attention of the world’s scientists for the welfare of human beings.