Microbial transformation of highly persistent chlorinated pesticides and industrial chemicals.

Organic pollutants can be transformed, both in unsaturated and saturated areas of the soil, by means of biologically mediated reactions. The potential of soil microorganisms to clean up polluted soils is enormous. However, soil systems are highly heterogeneous with respect to the spatial distribution of substrates, nutrients and microorganisms, and also with respect to various phases and interfaces (e.g. water, air, minerals, organic matter). To come to the development of appropriate soil bioremediation techniques, comprehensive knowledge is required about the biological and physico- chemical processes and parameters that determine the rate, the nature and the degree of degradation of organic compounds within a certain time scale in soil. The most important processes and parameters that influence the behavior of pollutants in soil are shown in Figure 7.1.<img src="/wda/abstracts/i2207_1.gif" height="400" width="600"/>Several aspects, which may arise when assessing the application of microorganisms in soil remediation. techniques, have been studied in this thesis.A major part of this thesis concerns the environmental conditions, that are essential for biodegradation. The biodegradability of a pollutant depends not only on the molecular structure of this pollutant, but also on the occurrence of the appropriate enviromnental conditions. The biodegradation of many enviromnental pollutants has been studied primarily under aerobic conditions, since these conditions are prevailing in polluted locations and easy to maintain. However, the anaerobic degradation of pollutants forms a relatively new potential of interesting reactions, which may be used for bioremediation purposes.For instance, β-hexachlorocyclohexane (β-HCH), a waste product from the production of the pesticide lindane, is considered extremely recalcitrant under aerobic conditions.This may be due to the spatial arrangement of the chlorine atoms in this HCH isomer. Disappearance of β-HCH under anaerobic conditions has been reported, but product identifications and mass balances were lacking in these studies. In Chapter 2, the biodegradation of β-HCH under anaerobic conditions is described. It is shown that β-HCH can be dechlorinated under methanogenic conditions. The observed dechlorination products of β-HCH, benzene and chlorobenzene, have individually been found for the anaerobic biodegradation of γ-HCH, but they were never reported to be formed simultaneously from any HCH isomer by microorganisms.The dechlorination products of β-HCH, benzene and chlorobenzene, accumulate under the conditions studied. The formation of these products, which are potentially toxic, introduces a new problem since chlorobenzene is recalcitrant under methanogenic conditions and anaerobic benzene degradation has been observed in only a few cases. However, many studies have shown that these compounds are degradable in the presence of molecular oxygen. Therefore a bioremediation process in which subsequently anaerobic and aerobic conditions are created would lead to a total mineralization of β-HCH.Such a sequential change from anaerobic to aerobic conditions also applies to the decontamination of soils polluted with polychlorinated benzenes (CBs) and biphenyls (PCBs). These compounds, when substituted with five or more chlorine atoms, generally persist under aerobic conditions. However, anaerobic reductive dechlorination of these compounds has abundantly been shown. The products of this dechlorination process are important for several reasons. From a toxicological point of view, because ortho -chlorinated PCBs are considered 10-100 times less toxic than non ortho -chlorinated ones (coplanar PCBs), but also from a biotechnological point of view, since the dechlorination products are the substrates in an eventually following aerobic treatment. Microbial reductive dechlorination of PCBs is mainly limited to meta - and para -dechlorination, and therefore reduces the toxicity of the PCBs. Dechlorination of CBs and PCBs also increases the water solubility and therefore the mobility of these compounds in the soil. This may be advantageous in the case of an ex situ bioremediation. However, since this process has been shown to occur in undisturbed polluted sediment and will also occur in biological soil treatments in situ , one should prevent the pollutant to move away from the polluted site.The dechlorination of CBs and PCBs is carried out by different groups of microorganisms, originating from different locations. Chapter 3 describes the dechlorination of these compounds in a methanogenic consortium, enriched from a mixture of Polluted sediments on 1,2,4-trichlorobenzene (TCB). CBs appear to be extensively dechlorinated by this consortium, with 1,2- and 1,3-dichlorobenzene (DCB) as well as chlorobenzene as the final products. This is in contrast with the performance of previously reported enrichments, which have been obtained mainly using hexa- and pentachlorobenzene as substrates for dechlorination. Such cultures show a lesser extent of dechlorination and produce 1,3,5-TCB or 1,3-TCB as final products, which are problematic compounds for an eventually following aerobic mineralization. Chapter 3 also shows that, whereas mono- through pentachlorobenzenes are dechlorinated within 2 days, hexachlorobenzene (HCB) and several PCBs are dechlorinated after a lag phase of about two weeks. Dechlorination of the latter compounds occurs via a different pathway, in which only chlorines with two adjacent chlorines are removed. Chapter 3 also shows that dechlorination of highly chlorinated benzenes can be directed to the formation of chlorobenzene as the only product by the addition of 2-bromoethanesulfonic acid (BrES), a methanogenic inhibitor. These observations show that different dechlorinating activities are present in the consortium. They are probably carried out by different physiological groups of microorganisms. In contrast, Chapter 4 describes a different methanogenic consortium which has the capacity to dechlorinate both CBs and PCBs with the same specificity toward chlorine substution pattern of these compounds. This and the fact that similar lag phases for the dechlorination of CBs and PCBs were observed, shows that the dechlorination of these compounds may be carried out by the same type of microorganisms.Chapter 5 applies to the environmental conditions which are necessary for a fast dechlorination of PCBs. Anaerobic reductive dechlorination of PCBs is a slow process, in which lag periods of weeks to months are commonly observed before dechlorination starts. Also the maintenance and subcultivation of PCB dechlorinating microbial consortia are extremely difficult. The optimal enviromnental conditions for the dechlorination process are not known. This is partly due to our lack of knowledge about the physiology of microbial reductive dehalogenation in general and about dechlorination of PCBs in particular. In Chapter 5, the lag phase of the reductive dechlorination of 2,3,4,5-tetrachlorobiphenyl (TeCBP), performed by the consortium described in Chapter 4, was reduced by means of different types of additions to the medium. Sterile Rhine sand appears to shorten the lag phase and to be essential for the maintenance of the consortium. A similar positive effect of sand was also observed in other dechlorinating consortia (Chapter 2 and 3) and for the degradation of toluene in a manganese reducing consortium.The mechanism of the reduction of the lag phase by Rhine sand is not known. The Rhine sand that we added to our medium, may contain additional trace metals, necessary for dechlorination. Although the organic carbon content in the Rhine sand we used is very low (&lt;0.05% w/w), another explanation may be that the organic matter served as a supply of growth factors, like vitamins, for the dechlorinating bacteria. Sorption of 2,3,4,5-TeCBP to the sand might also improve the contact between the PCBs and the dechlorinating cells, if these cells tend to settle out or grow on the sand surface (adhesion).The addition of anaerobic granular sludge (AGS), either autoclaved or not, also reduced the lag phase before dechlorination of 2,3,4,5-TeCBP drastically. The decrease of the lag phase, achieved by the addition of AGS, may be caused by factors, originating from microorganisms in the AGS. However, the addition of the redox mediator vitamin B _{<font size="-2">12</font>} , Which is abundantly present in AGS, had little effect on the lag phase. Reduction of the lag phase for PCB dechlorination was also observed by Abramowicz et al. who added fluid thioglycolate medium with beef extract (FTMBE) to polluted Hudson River sediments. In fact, the use of undefined additions such as AGS and FTMBE illustrates how little is known about the mechanism of reductive dechlorination of PCBs. It is of practical importance that a relatively cheap and easily available material such as AGS also reduces the lag phase for dechlorination in a methanogenic enrichment, which is capable of dechlorinating the complex PCB mixture Aroclor 1260 (Chapter 5). It may be worth trying to introduce AGS into methanogenic river sediments polluted with PCBs, to stimulate in situ dechlorination.Another aspect one should consider when taking biodegradation into practice is the availability of the pollutant for biodegradation. This is illustrated in Chapter 6, which deals with the biodegradation of pentachlorophenol (PCP) in different soil types by inoculated Rhodococcus chlorophenolicus. After an initially high mineralization rate of the PCP in the inoculated soils, this rate decreased during time and levelled off, leaving behind a residual amount of PCP. Such a non- degraded residue has been observed in many soil biodegradation studies and is referred to as a fraction that is not available for biotransformation, i.e. not bioavailable. Limited bioavailability is considered as one of the most important problems impeding the development of successful bioremediation techniques.The rate at which microorganisms can degrade pollutants during bioremediation depends on the intrinsic metabolic activity of the cell, and on the transfer of the pollutant to the cell. Thus, bioavailability of a pollutant is determined by the rate of mass transfer, relative to the activity of the microbial cells. As bacteria generally degrade pollutants to concentrations which are much lower than the residual concentrations observed in soils, it appears that mass transfer is the limiting factor for further degradation. This was confirmed by Bosma et al. who quantified bioavailability by modelling the effect of both mass transfer and the intrinsic activity of the microbial cells. Bioavailability of a pollutant is controlled by a number of physico-chemical processes, such as sorption, diffusion and dissolution. Especially in old polluted sites, part of the pollutants appears to be inaccessible to degrading bacteria. This was shown for a soil contaminated with chlorophenols for over 40 years, in which bioremediation through composting resulted in a residual concentration of about 40 mg/kg. No further degradation was observed when the soil was inoculated with PCP mineralizing bacteria. However, <sup><font size="-2">14</font></SUP>C-labelled PCP which was freshly added to this soil led immediately to a complete mineralization of the added PCP. Similar results were reported for two soils contaminated with polycyclic aromatic hydrocarbons and for the reductive dechlorination of hexachlorobenzene in sediment. Rijnaarts et al. showed that biodegradation rates of α-hexachlorocyclohexane in soil could be increased by breaking up the soil particles.Such observations indicate a reduced bioavailability due to a so-called "ageing" of the pollutant in the soil. This may be caused by different processes. The first process is a slow diffusion of the pollutant into the deeper and smaller pores of the soil particles and absorption into organic matter as visualized in Figure 7.2. The pollutant can only reach the degrading microorganisms - which are not small enough to penetrate into the deeper pores - by means of diffusion.<img src="/wda/abstracts/i2207_2.gif" height="455" width="600"/>A second process is the occurrence of the pollution as solid particles or as non-aqueous phase liquids (NAPLs) with surrounding semi-rigid films, from which mass transfer into the water phase is limited. Here, bioavailability is controlled by the actual dissolution rate. The third process is bound residue formation and involves the incorporation of pollutants into humic compounds via oxidative coupling reactions. This process especially applies to naphtolic compounds, chlorinated phenols, benzoic acids and anilines, due to their similarity to natural organic compounds. Oxidative coupling can be catalyzed by inorganic materials, such as sesquioxides, clay minerals, oxides and oxohydroxides of iron, silica, and allophane, but also by peroxidases and phenol monooxygenases. Bound residue formation alters the original chemical and biological activity of the pollutants, which therefore become less available, less toxic and less mobile in the soil system. Chapter 6 shows that PCP, freshly added to a sandy and a peaty soil, can be partly recovered as extractable organic halogen (EOX) after 4 months of incubation. However, a chloride mass balance reveals that in the peaty soil, not all measured EOX can be ascribed to PCP, its transformation product pentachloroanisole, or any other possible low molecular weight organic chlorinated compound (MW &lt; 512). This indicates that bound residue formation may have taken place in a time period of 4 months.The use of the oxidative enzymes has been proposed as a method to immobilize organic pollutants into the humic material of the soil by stimulating bound residue formation. However, uncontrollable release and distribution of the pollutant cannot be guaranteed. Macromolecular components of dissolved organic matter may form colloids which serve as carriers facilitating the transport of contaminants that bind to the organic matter. Furthermore, fungal oxidative enzymes have also been shown to catalyze unwanted reactions, like the formation of highly toxic chlorinated dibenzodioxins from coupling reactions with chlorinated phenols as substrate.It can be concluded that bioavailability is more a physico-chemical problem rather than a microbiological problem. Solutions have therefore to be sought in physico-chemical techniques to increase the transport of the pollutants to the degrading bacteria. These techniques may comprise pulverizing soil particles, increasing the water content of the soil (slurry systems), heating the soil, and using detergents to increase solubility and/or desorption.In this thesis, the ever increasing capability of microorganisms to degrade chlorinated pesticides and industrial chemicals is demonstrated, provided that the appropriate enviromnental conditions are created. Despite this enormous potential, the limited bioavailability of pollutants in old contaminated sites and strict soil quality standards, largely hamper the application of microorganisms in soil remediation techniques. The mentioned techniques to increase the availability of pollutants have economical or enviromnental drawbacks. Alternative remediation techniques like incineration or solvent-extraction yields soils without any biological activity, which' destiny is mainly to be used as raw material for building purposes.On the other hand, one should ask what the consequences can be of the slow release of the residual amount of a pollutant from a soil that has been biologically treated. (Eco)toxicological assessment of such soil may serve the development of guidelines for the urgency to remove residual pollutants from such soils. Such an approach would stimulate the use of in situ bioremediation techniques, which eventually prevents polluted soil from being excavated. Of course, a continuous monitoring of the soil quality and an appropriate groundwater management of run-off water are essential