6. Occurrence and behaviour of organic substances in European groundwater
authors: Bas van der Grift and Jan Gerritse (TNO, the Netherlands)
Contents
6.1 Introduction
6.2 Sorption and degradation of organic substances in groundwater 6.2.1 Hydrophobic sorption of organic substances
6.2.2 Degradation of organic substances in groundwater 6.3 Aromatic hydrocarbons
6.3.1 Degradation of aromatic hydrocarbons in groundwater 6.3.2 Benzene
6.3.3 Polycyclic aromatic hydrocarbons 6.4 Chlorinated aliphatics
6.4.1 Degradation of volatile chlorinated aliphatics in groundwater 6.4.2 Trichloromethane
6.4.3 Dichloromethane 6.4.4 1,2-Dichloroethane 6.4.5 Trichloroethylene 6.4.6 Tetrachloroethylene 6.4.7 Hexachlorobutadiene
6.4.8 Short chain chlorinated paraffins (SCCPs) C10-13 6.5 Chlorinated aromatics
6.5.1 Hexachlorobenzene 6.5.2 Pentachlorobenzene 6.5.3 Trichlorobenzenes 6.6 Substituted phenols
6.6.1 Pentachlorophenol
6.6.2 Nonylphenols and Octylphenols 6.7 Chlorinated pesticides
6.7.1 Introduction
6.7.2 Leaching to groundwater 6.7.3 Hexachlorocyclohexane
6.7.4 Endosulfan 6.7.5 Alachlor 6.7.6 Atrazine 6.7.7 Simazine 6.7.8 Chlorfenvinphos 6.7.9 Chlorpyrifos 6.7.10 Diuron 6.8 Polar pesticides
6.8.1 Isoproturon 6.8.2 Trifluralin 6.9 Others
6.9.1 Brominated diphenylether
6.9.2 Di(2-ethylhexyl)phthalate (=DEHP) 6.9.3 Tributylin compounds
6.10References
6.1 Introduction
Chapter 6 deals with the behaviour and occurrence of xenobiotic organic compounds in groundwater. Xenobiotic organic compounds (XOCs) can broadly be defined as ‘all organic compounds that are released in any compartment of the environment by the action of man and thereby occur in a concentration that is higher than natural’ (Leisinger, 1983). The number of known XOCs increases with time as better methods for analysis are developed. The occurrence and behaviour of xenobiotic compounds in groundwater environments has mainly been studied in landfill leachate plumes and industrially polluted sites. To date, more than 1000 organic chemicals have been identified in groundwater (Christensen et al., 2001). Even though XOCs typically constitute less than a few per cent of the total dissolved C in groundwater, the fate of XOCs in the affected aquifer is of major concern. This group of pollutants pose a potential health risk (Brown and Donnelly, 1988) and strict drinking water standards are enforced in many countries, with acceptable concentrations often as low as 0.1 µg/l for individual XOCs.
The organic compounds of the WFD pollutants and the EC Priority Substances are subject of this review (Table 1). The state of the art regarding groundwater contamination in the EU with these chemicals will be described. Focus lies on the occurrence and attenuation of these chemicals in groundwater. Attenuation of XOCs is (besides dilution) due to sorption and degradation. For some volatile compounds, volatilization may also occur. However, this is only important at shallow depths and in unsaturated zones, and as such will not be discussed here. In this section, the current knowledge on sorption and degradation in different redox environments is reviewed.
For this review the organic compounds of the WFD pollutants and the EC Priority Substances are divided in functional categories (Table 6.1).
The set-up of this chapter is as follows: paragraph 6.2 gives an overview of recent developments in the knowledge of sorption and degradation of XOC that are specifically related to individual substances. The different categories of XOC are described in paragraphs 6.3 to 6.9. These individual paragraphs start with an overview of the characteristics of the substance category followed by a description of the individual substances belonging to the category. The substances are discussed in a standard format following chemical name, properties, use, environmental exposure and observed groundwater contamination and degradation.
Table 6.1. Functional categories and substances described in this chapter
Category Substances
Aromatic hydrocarbons Benzene Anthracene Fluoranthene Naphthalene Chlorinated aliphatic hydro-
carbons
Trichloromethane 1,2—Dichloroethane Dichloromethane Trichloroethylene Tetrachloroethylene Hexachlorobutadiene C10-13—chloroalkanes Chlorinated aromatic hydro-
carbons
Hexachlorobenzene Pentachlorobenzene Trichlorobenzenes Substituted phenols Pentachlorophenol
Nonylphenols Octylphenols
Chlorinated pesticides Hexachlorocyclohexane Endosulfan
Alachlor Atrazine Simazine Chlorfenvinphos Chlorpyrifos Polar pesticides Diuron
Isoproturon Trifluralin
Others Brominated diphenylether
Di(2—ethylhexyl)phthalate (=DEHP) Tributylin compounds
6.2 Sorption and degradation of organic substances in groundwater
6.2.1 Hydrophobic sorption of organic substances
Sorption of organic compounds is one of the key processes governing their environmental behaviour (Schmidt et al., 2005). Sorption influences transport, in particular in porous media but also across compartment interfaces, bioavailability and, consequently, adverse effects and
biotransformation of organic compounds. The phase transfer process to a solid can involve adsorption occurring on surfaces, and partitioning into a bulk phase. Depending on the type of soil organic matter (SOM), adsorption, e.g., to charcoal-type particles, and/or partitioning into the three-dimensional SOM structure might be of importance. The characterization of organic matter facies in environmental samples, e.g., by using methods from coal petrography, therefore has attracted increasing interest (Kleineidam et al., 1999).
Typically, adsorption dominates at low aqueous concentrations, and partitioning becomes more important at higher concentrations, and nonlinear isotherms are found (Kleineidam et al., 2002).
The difference in reaction rate is also worth mentioning. Adsorption to the particles is a equilibrium process while partitioning into the three-dimensional matrix is a much slower process. This causes ‘aging’ or ‘hystereses’ which means that the sorption relation between the concentration in aqueous phase and in the solid during the time of net adsorption different is than during the period of net desorption. An appropriate general description of sorption therefore needs to take into account both mechanisms. Various forms models have been suggested that usually represent linear combinations of a partitioning isotherm and an
adsorption isotherm, e.g., Langmuir, Freundlich (Accardi-Dey and Gschwend, 2002) or Polanyi (Kleineidam et al., 2002).
Sorption from water is characterized by the solid-water distribution coefficient, Kd, which indicates the ratio of equilibrium concentrations in the solid, Cs, and the aqueous phase, Cw (Kd=Cs/Cw). Note that for nonlinear sorption, Kd depends on the solute concentration. The conventional approach for the determination of Kd is to carry out batch equilibrium studies. By measuring the initial aqueous concentration (Co), equilibrium aqueous concentration (Cw), and solid phase concentration Cs (usually determined from a mass balance), Kd can be calculated.
For many compounds and solids this is a very reliable method. Recently, it was shown that in the derivation of sorption isotherms it is often useful to plot Cs vs. the solubility-normalized aqueous concentration Cw/S (Kleineidam et al., 2002).
For a-priori assessments of the environmental fate of organic compounds, methods that allow a reliable prediction of sorption are mandatory (Schmidt et al., 2005). Numerous methods have therefore been put forward, including linear free energy relationships (LFER) that link the sorption of a compound to a single physico-chemical parameter such as the octanol-water partitioning constant or water solubility (Schwarzenbach et al, 1993). Since these parameters are available for many compounds, this is a straightforward way to estimate sorption
coefficients within a given compound class. However, such one-parameter LFERs need to be carefully checked for validity and misuses in the literature are abundant. A more recent
development is poly-parameter LFERs that capture the relevant molecular interactions between an organic solute and surrounding phases of interest. This approach has been successfully used to predict both adsorption on surfaces from air and partitioning between environmental
compartments, including air, water and organic solvents (Goss and Schwarzenbach, 2001).
Recently, it was shown that it is in principle also applicable to soil sorption (Nguyen et al., 2005).
6.2.2 Degradation of organic substances in groundwater
Biodegradation is the most important process for the removal of organic substances from groundwater systems. In the groundwater and attached to the soil matrix a wide variety of
micro-organisms (mainly bacteria and fungi) performs biochemical reactions that decompose complex organic molecules into inorganic minerals. By using specific enzymes the microbes are able to transfer electrons from electron donor substrates to electron acceptors (Figure 6.1).
Through this mineralization process the microbes obtain their energy (e.g. ATP, NADH) and minerals (e.g. C, N, P, S) for life and reproduction (new cells). In many cases the organic xenobiotic compounds can be used as a sole source of carbon and electrons for growth. Aerobic bacteria and fungi use O2 as terminal electron acceptor for the oxidation of organic compounds.
Important alternative electron acceptors that can be available in anoxic groundwater systems are nitrate, manganese(IV)- and iron(III)oxides, sulphate, carbon dioxide and humic acids. Many (poly)halogenated organic pollutants can also serve as electron acceptor and are reduced by
“dehalorespiring” bacteria. Some organic xenobiotic compounds can act as both electron donor and acceptor, yielding a mixture of reduced and oxidised products, in a process called
fermentation. Although complete mineralization of xenobiotic compounds often occurs, this is not always possible when suitable microbes and enzymes that can break down the molecule are not available. In this case the xenobiotic is recalcitrant and may persist in the environment.
Sometimes microbes transform an organic compound without using energy or carbon from it.
This co metabolism can occur under oxic and anoxic conditions. It is mediated by enzymes that micro-organisms use to degrade their natural growth substrates, but have a broad substrate range and also transform the xenobiotic compound. The co metabolic transformation products are often used by other bacteria as growth substrates, resulting in complete degradation.
Figure 6.1 Principle scheme of the microbial biodegradation process
Degradation of xenobiotic compounds in the environment depends on many factors. Obviously, micro-organisms with proper biodegradation capabilities should be present or able to evolve in the polluted environment. When the properties of the xenobiotic deviate much from natural substrates the micro-organisms may not be able to readily decompose the compound.
Hydrophobic compounds may be recalcitrant because they are not “bioavailable” due to low solubility in the groundwater and sorption to the soil matrix. Many xenobiotic compounds are also toxic for the biodegrading bacteria. Finally, the environmental conditions must be suited for
microbial activity. Favourable conditions for pollutant degradation include the availability of nutrients, pH in the range from 5 to 9, temperature between 10 and 30°C and the presence of sufficient water. The availability of electron acceptors is of crucial importance for the
biodegradation of specific xenobiotics (Figure 6.2). Many organic compounds are best degraded in an oxidising environment. One reason is that the amount of energy that bacteria can obtain from the transfer of electrons from a reduced compound to an electron acceptor generally decreases in the order: O2 > NO3
- > Mn4+ > Fe3+ > SO4
2- > CO2. A second reason is that O2 is not only used as terminal electron acceptor, but also as a reactant for the initial conversion of aromatics, such as PAH and BTEX, and aliphatics such as mineral oil. In contrast, oxidised polychlorinated compounds are best dechlorinated in an anaerobic reducing environment. Under these conditions specific anaerobic bacteria can use organic substrates or H2 to fuel reductive dechlorination reactions. The (partially) dechlorinated products can further be degraded when alternative electron acceptors become available.
Figure 6.2. Trends in biodegradation of groundwater pollutants under various redox conditions
Demonstrating the in situ biodegradation of specific contaminants is a critical issue in the assessment of the natural attenuation of organic contaminants in groundwater pollution plumes (Richnowa et al., 2003). In particular, under strictly anoxic conditions, demonstrating the in situ biodegradation of aromatic hydrocarbons may be difficult (Christensen et al., 2000). This is primarily related to three facts: (i) the concentrations of the aromatic hydrocarbons are often low to moderate, although still of concern, (ii) the degradation rates are low (Rügge et al., 1999) and (iii) patterns of electron acceptors used as supporting evidence for the biodegradation of
contaminants are biased in leachate plumes because the dissolved organic carbon (DOC) in the leachate acts as the primary electron donor in the plume.
Numerous studies have been published in literature with data of biodegradation rates of organic compounds in groundwater. In general, a comparison between laboratory and field results indicates that biodegradation constants obtained from laboratory studies are generally higher than field studies (Suarez and Rifai, 1999). The term degradation has been used liberally in the literature. It has been defined as compound disappearance (compared with no disappearance in control experiments), by identification of end products (using e.g. radio-labeled molecules or isotopes), or by revealing degradation pathways, intermediate compounds, and active bacterial species or enzymes. The different definitions and perceptions of what degradation is makes it difficult to compare degradation results from different studies (Christensen et al., 2001).
6.3 Aromatic hydrocarbons
6.3.1 Degradation of aromatic hydrocarbons in groundwater
The aromatic hydrocarbons generally degrade readily under aerobic conditions, but very sparingly under anaerobic conditions. Aerobic transformation of PAHs associated with soil and groundwater often leads to rapid depletion of dissolved oxygen and this eventually decreases the redox potential (Eh) to sulphate-reducing, or even methanogenic conditions. Recent
observations in a number of natural attenuation studies suggest a significant potential for anaerobic degradation, especially for TEX (Christensen, 2002). However, the first order degradation rates observed under unspecified anaerobic conditions (Rifai et al., 1995; Suarez and Rifai, 1999) are typically 1 or 2 orders of magnitude lower than rates reported for aerobic conditions (Nielsen et al., 1996a).
Aerobic biodegradation of aromatic hydrocarbons
Aerobic biodegradation of aromatic hydrocarbons takes place via oxidation with molecular oxygen (O2) as a reactant. In this approach, organisms invest biochemical energy to form very reactive species that can interact with the organic compounds via mechanisms not operating in abiotic dark environments. The goal of these initial reactions is to transform the xenobiotic compound into a product(s) which is structurally more similar to chemicals which micro- organisms are used to metabolise. After one or several initial transformations, the resulting chemical product(s) are funnelled into in the common metabolic pathways and are fully mineralised to CO2 and water. Aerobic biodegradation of benzene is a good example of this pathway. Initially the aromatic hydrocarbon is oxidized to catechol by a dioxygenase (Figure 6.3). The aromatic ring of catechol is subsequently cleaved by a catechol-dioxygenase, between the two hydroxyl substituents (ortho cleavage) or adjacent to them (meta cleavage) leaving non- aromatic organic acid(s). Further conversion via several enzymatic reactions results in
formation of acetyl-CoA and succinate (ortho cleavage) or acetaldehyde and pyruvate (meta cleavage), which are readily mineralized via the tricarboxylic acid cycle. Catechol and its substituted derivatives are also produced in the metabolism of numerous natural aromatic compounds.
Figure 6.3. Pathway aerobic degradation benzene.
Anaerobic degradation of aromatic hydrocarbons
Anaerobic pathways of aromatic hydrocarbons biodegradation are important because these compounds are frequently found under conditions where the potential use of oxygen quickly exceeds the supply. These conditions are often found in contaminated groundwater. The lack of molecular oxygen as reactant for aromatic ring activation and cleavage explains why anaerobic degradation of aromatic hydrocarbons is more difficult than aerobic conversion. In many cases, degradation of benzene and PAH in anoxic groundwater systems appears to occur at very low or even undetectable rates. Nevertheless, microcosm and field studies have indicated that
biodegradation of diverse aromatic compounds, including benzene and naphthalene, can occur under different redox conditions, coupled to nitrate reduction, iron(III)oxide reduction, sulphate reduction or methanogenesis as terminal electron accepting processes (Bianchin et at., 2006;
Grbić-Galić, 1987; Lovley, 2000). Even the degradation of higher condensed PAH such as fluoranthene has been reported (Meckenstock et al., 2004). A negative correlation between the molecular mass of the compound and its degradation rate was observed under similar
experimental conditions in batch experiments (Meckenstock et al., 2004). So far, except for naphthalene, no degradation of unsubstituted PAH has been found under methanogenic
conditions, although compounds with polar substituents such as naphthol (isomer not specified) can be metabolized (Mihelcic & Luthy 1988).
The current information on the anaerobic degradation of polycyclic aromatic hydrocarbons shows definitely that at least naphthalene and other two-ring PAHs (Meckenstock et al., 2004) can be used as the sole source of carbon and energy under sulfate-, iron-, and nitrate-reducing conditions. Several anaerobic micro-organisms which can grow with aromatic hydrocarbons as sole carbon and energy source have been isolated in the last years. Studies on their degradation pathways resulted in detailed information on the biochemistry of transformation mechanisms.
Nevertheless, the anaerobic degradation mechanisms of benzene and PAH are not completely resolved. Anaerobic pathways for breakdown of the aromatic ring are different and quite distinct from the aerobic pathways (Karthikeyan & Bhandari, 2001). Aromatic ring cleavage occurs via reductive transformations. To this end the aromatic ring is first destabilised. Activation reactions may involve hydroxylation, carboxylation, methylation, fumarate addition or acetyl-CoA addition. Subsequently channelling reactions transform the xenobiotic compound to central metabolic intermediates, followed by reduction of the aromatic ring and hydrolytic ring cleavage.
The resulting acids are further mineralised.
It is uncertain if the organisms isolated so far really have a major function in the field or they are merely organisms that grew fastest under the applied cultivation conditions. Molecular
studies might elucidate in the future which strains or microbial communities are the dominating degrading organisms in contaminated aquifers. An important challenge is to identify the limitations of biodegradation in contaminated aquifers. The question why some contaminants are only poorly degraded in situ although the organisms that could degrade the compounds can be isolated from the aquifer and degradation activities can be detected by compound-specific metabolites or stable isotope fractionation is largely unsolved.
6.3.2 Benzene
Chemical Name: Benzene (C6H6)
CAS Number: 71-43-2
Properties: Solubility in water: 1800 mg.l-1 at 25ºC; vapour pressure: 9970 (Pa) at 20°C; log KOW: 2.13 (calculated); KOC 134.15 (calculated - this value is located within the range of experimentally determinations).
Use: benzene is used for printing & lithography, paint, rubber, dry cleaning, adhesives &
coatings, detergents, extraction and rectification, preparation and use of inks in the graphic arts industries, as a thinner for paints and as a degreasing agent. In the tire industry and in shoe factories, benzene is used extensively.
It is also used as a raw material in the synthesis of styrene (polystyrene plastics and synthetic rubber), phenol (phenolic resins), cyclohexane (nylon), aniline, maleic anhydride (polyester resins), alkylbenzenes (detergents), chlorobenzenes, and other products used in the production of drugs, dyes, insecticides, and plastics.
Environmental exposure and observed groundwater contamination: Benzene will enter the atmosphere primarily from fugitive emissions and exhaust connected with its use in gasoline.
Another important source is emissions associated with its production and use as an industrial intermediate. In addition, there are discharges into water from industrial effluents and losses during spills. Benzene is also released from its indirect production in coke ovens; from nonferrous metal manufacture, ore mining, wood processing, coal mining and textile
manufacture. Although most public drinking water supplies are free of benzene or contain <0.3 ppb, exposure can be very high from consumption of contaminated sources drawn from wells contaminated by leaky gasoline storage tanks, landfills, etc.
The contamination of the groundwater with benzene as point sources is seen at many sites, especially those associated with petrochemical industry like gasoline stations, former gaswork site and landfills. The associated range in groundwater concentrations is large, ranging from the detection limit to it’s solubility. The highest concentrations are found near pure fuel NAPL’s.
The concentrations are then dependent on the solubility of the substance. When chemicals are released into the environment from a mixture like a petroleum hydrocarbon fuel, the water
solubilities of the chemicals are typically far lower than their published solubilities. For example, the solubility of benzene is around 1800 mg/L, but typical maximum benzene concentrations resulting from equilibrium between gasoline and water are only 20 - 40 mg/L.
This occurs because the concentration (or effective solubility) depends on the abundance of the chemical in the fuel.
Benzene is observed in landfill leachates in a concentration range of 0.2-1630 µg/l (Kjeldsen, 2002).
Degradation: Suarez & Rifai (1999) presented a comprehensive review of rates of
biodegradation of fuel hydrocarbons and chlorinated solvent is groundwater obtained from field and laboratory studies. Figure 6.4 shows their results as first-order biodegradation rates for benzene. The difference between lab-studies and field experiments and aerobic and anaerobic conditions is obvious. Biodegradation rates obtained from laboratory studies are generally higher than field rates. This result can be explained because laboratory studies maintain
favourable ambient conditions for biodegradation. Benzene shows a big difference (three orders of magnitude) between aerobic and anaerobic coefficients. The data in Figure 6.4 indicate that the range of reported anaerobic degradation rate constants for benzene is very wide. The reason is most likely that in the absence of oxygen benzene biodegradation can be coupled to a variety of electron-accepting conditions (Coates et al., 2002). However, median benzene degradation value under anaerobic field conditions is essentially zero.
10
1
0.1
0.01
0.001
0.0001
0.00001
in situ studies lab studies in situ studies lab studies anaerobic rates
first-order biodegradation rate (1/day)
aerobic rates
maximum
P75
P25 median
minimum Legend
42
59 23
3
Figure 6.4. First-order biodegradation rates for benzene (Suarez & Rifai,1999).
In 11 landfill leachate studies of XOC’s degradation under anaerobic conditions benzene was reported as not degraded while in 1 study it was reported as degraded (Christensen, 2002). Baun et al. (2003) also showed that benzene was not removed in an anaerobic part of a landfill leachate plume influenced aquifer.
US-EPA (1999) reported as most likely first order biodegradation rate for benzene of 0 (1/day) under anaerobic conditions, together with a maximum of 0.071 and a standard deviation of 0.0152. This estimate is based on 25 field and laboratory studies with different temperature, pH and redox conditions.
Very little is known about the micro-organisms responsible for anaerobic benzene degradation.
The bacteria involved in benzene degradation under methanogenic or sulphate-reducing conditions are unknown. Molecular analysis of dominant microbial populations in Fe(III) reducing sediments demonstrated that Geobacter species may be involved the anaerobic oxidation of benzene to CO2 (Lovley, 2000). However, the known Geobacter species that are available in pure culture have not been shown to oxidise benzene. The only known bacteria that are capable of anaerobic growth on benzene are two Dechloromonas strains, RCB and JJ (Coates et al., 2001). Dechloromonas sp. strain RCB degrades benzene, toluene, ethylbenzene and xylene compounds aerobically or anaerobically with (per)chlorate or nitrate as electron acceptor. Benzene is completely oxidised to CO2 (Figure 6.5). The crucial steps of anaerobic benzene degradation by this bacterium are a hydroxylation and carboxylation reaction (Chakraborty and Coates, 2005). The authors suggest that all anaerobic benzene-degrading bacteria may use this pathway, regardless of their terminal electron acceptor. This is in agreement with the detection of phenol and benzoate as intermediates of anaerobic benzene degradation under various terminal electron accepting conditions (Caldwell and Suflita, 2000;
Grbić-Galić and Vogel, 1987).
Figure 6.5. Proposed mechanism of anaerobic benzene degradation (Chakraborty and Coates, 2005).
6.3.3 Polycyclic aromatic hydrocarbons
Chemical Name: Polycyclic aromatic hydrocarbons (PAHs) form a group of compounds consisting of two or more fused aromatic rings.
Naphthalene (C10H8) Anthracene (C14H10) Fluoranthene (C16H10);
CAS Number: Naphthalene: 91-20-3; Anthracene: 120-12-7; Fluoranthene: 206-44-0 Properties:
Naphthalene: Molecular mass: 128.17 g/mol; Solubility in water: mg.l-1 at 25ºC; vapour pressure: 9970 (Pa) at 20°C; log KOW: 2.13 (calculated); KOC 134.15
Anthracene: Molecular mass: 178.23 g/mol; Solubility in water: 0.041 mg.l-1 at 25ºC; vapour pressure: 0.0008 (Pa) at 20°C; log KOW: 4.54 (measured at 20°C); KOC: measured values ranging from 2600 to 8600
Fluoranthene: Molecular mass: 202.26 g/mol; Solubility in water: 0.265 (between 0.22 and 0.265 mg/L at 25°C); vapour pressure: 0.0008 (Pa) at 20°C; log KOW: 5.16; KOC: 144544 Use: PAH’s are mainly used as intermediates in the manufacture of alkyd and polyester resins, dyes (fluorescent), pharmaceuticals, pigments and agrochemicals. PAHs are also formed by pyrolysis of fossil fuels and are constituents of coal tar, oil products, tobacco smoke, automobile exhaust gas, incinerated waste, industrial effluents and urban air.
Environmental exposure and observed groundwater contamination: Due to their wide
distribution and because of their toxic, mutagenic and carcinogenic properties the environmental pollution by PAHs has aroused global concern. Combustion sources are thought to account for over 90% of the environmental burden of PAHs. In particular stationary point sources account for around 90% of these inputs (UNEP, 2002, Howsam and Jones, 1998). Inputs to the
atmosphere are dominated by emissions associated with residential heating (coal, wood, oil and gas burning) and industrial processes such as coke manufacture. Recent estimations of XOC atmospheric deposition rates to soils in the Netherlands are made by Duyzer & Vonk, 2003. For Naphthalene, Anthracene and Fluoranthene they report values of 0.0253, 0.0044 and 0.099 kg/km2.jr. Non-combustion processes such as the production and use of creosote and coal-tar (and the remediation of sites contaminated with these substances), though poorly quantified, are potentially very significant primary and secondary sources. The significance of these inputs may be set to increase within Europe as legislative restrictions on emissions from combustion
processes continue to be enacted.
Time trend analysis using a variety of depositional media (sediment cores, polar ice and peat) has shown that developed countries have succeeded in reducing PAH emissions from a wide range of sources through technological applications and legislation in addition to a marked shift away from the use of coal and coal products. Nations undergoing rapid industrialisation may well prove to be an increasingly significant source of PAH in global terms as their populations continue to grow as the number of mobile and industrial sources increase.
The contamination of the groundwater with PAHs is seen at many sites, especially those
associated with petrochemical industry like gasoline stations. Low aqueous solubility’s of PAHs and high octanol-water partition coefficients (KOW) often result in their accumulation in soils and sediments to levels several orders of magnitude above aqueous concentrations.
Naphthalene is also observed in landfill leachates in a concentration range of 0.1-260 µg/l (Kjeldsen, 2002).
Degradation: Persistence of the PAHs in the environment depends on many factors (Van Agteren et.al., 1998). Bioavailability and biodegradation of PAHs generally decreases with the number of aromatic rings. The low molecular weight PAHs are most easily degraded. The reported half-lives of naphthalene, anthracene and benzo(e)pyrene under aerobic conditions in acclimated sediments range from 5 to 9, 43 to 280, and 83 to 21,000 hours, respectively. In non- acclimated sediments the half-lives were 10 to 400 times higher. For higher molecular weight PAHs, half-lives are up to several years in soils/sediments have been reported.
The availability of oxygen is an important factor for biodegradation. Under aerobic conditions PAHs can be degraded by bacteria, fungi, algae and yeasts. In soil systems bacteria and fungi are most likely the micro-organisms responsible for PAH biodegradation. Many aerobic bacteria have been isolated that can use naphthalene, anthracene or fluoranthene as the sole source(s) for growth and energy. The aerobic bacterial metabolisms involves dioxygenases which incorporate molecular oxygen in one of the aromatic rings. The resulting dihydroxynapthalene,
dihydroxyanthracene and dihydroxyfluoranthene are ring-cleaved by a second dioxygenase and further transformed to mono-aromatic acids or catechol, after which complete mineralisation can occur. Figure 6.6 shows the initial reactions of the aerobic naphthalene degradation pathway.
Figure 6.6. Proposed pathway for the initial reactions of the aerobic bacterial degradation of naphthalene (http://umbbd.ahc.umn.edu/index.html).
Fungi cannot use PAHs as a sole source of carbon and energy for growth but transform these compounds cometabolically. The initial step of fungal metabolism involves the action of cytochrome P450-like monooxygenases or extracellular ligninases and peroxidises. These enzymes partially oxidise the aromatic ring, resulting in the formation epoxides and/or hydroxylated aromatic compounds. The further transformations of these compounds have not been studied in detail. In many cases they are released in the environment where they serve as
substrates for growth of various aerobic bacteria. The specific contributions of bacterial and fungal (co)metabolism to PAHs degradation in groundwater systems is currently unknown.
The first indications on anaerobic PAH degradation of PAH-compounds are from Mihelcic and Luthy who reported degradation of naphthalene, naphtol and acenaphthene in microcosms under denitrification conditions (Mihelcic and Luthy, 1988). In the presence of nitrate, naphtol, naphthalene and acenapthene were degraded from initial aqueous-phase concentrations of 8, 7 and 0.4 mg/l to non-detectable levels in 16, 45 and 40 days, respectively. Napthol was also degraded in the absence of nitrate but naphthalene and acenaphthene were not. Subsequent studies revealed that the oxidation of PAH can be coupled to the reduction of sulphate, iron(III) oxides and manganese(IV) oxide (Langenhoff et al., 1996; Meckenstock et al., 2004). It has been demonstrated that a variety of PAHs, including naphthalene, fluoranthene and anthracene, can be degraded anaerobically (Chang et al., 2002; Coates et al., 1997; Meckenstock et al., 2004). PAH degradation rates in microcosm studies were in the range from 0.03 to 37 µM/day, while higher rates from 120 to 384 µM/day were observed in soil column systems. Most research on anaerobic PAH degradation was done with naphthalene as a growth substrate. It remains unclear whether three or more-ring PAHs are also used as growth substrates or they are only co-metabolised. Oxidation of naphthalene to CO2 was coupled to sulphate, nitrate,
manganese (IV) or iron(III) as terminal electron acceptor. Early studies on bacterial enrichment cultures suggested that the initial reaction of the sulphate-coupled naphthalene oxidation was a hydroxylation to naphtol (Bedessem et al., 1997) or a carboxylation to 2-naphthoic acid (Zhang and Young, 1997). Pure cultures of denitrifying and sulphate-reducing bacteria that can grow on naphthalene have been isolated and their degradation pathways have partially been resolved. A pure culture of the sulphate-reducing bacterium strain NaphS2 was able to grow on naphthalene and 2-naphthoic acid, but not on naphtol (Galushko et al., 1999). This is in agreement with the carboxylation pathway. The doubling time of strain NaphS2 was between 6 and 8 days.
Recently however, Safinowski and Meckenstock found evidence that neither hydroxylation nor carboxylation but methylation is the initial reaction in anaerobic naphthalene degradation by a sulphate-reducing enrichment culture (Safinowski and Meckenstock, 2006). They indicated that the biochemical degradation pathway of naphthalene is a methylation to 2-methylnaphthalene, with subsequent fumarate addition, oxidation to the central metabolite 2-naphthoic acid, followed by ring reduction and cleavage (Figure 6.7). The authors proposed that methylation is a general mechanism of activation reactions in anaerobic degradation of unsubstituted aromatic hydrocarbons.
Despite the results of scientific research as described above, PAHs have often been found to be recalcitrant under anaerobic conditions. In 8 landfill leachate studies of XOCs degradation under anaerobic conditions naphthalene was reported as not degraded while not one study reported it as degraded (Christensen, 2002). US-EPA (1999) reported as most likely first order biodegradation rate for Naphthalene of 0 (1/day) under anaerobic conditions, together with a maximum of 0.03 and a standard deviation of 0.00791. This estimate is based on 18 field and laboratory studies with different temperature, pH and redox conditions.
Figure 6.7. Proposed pathway for anaerobic naphthalene degradation (after Meckenstock et al., 2004 and Safinowski and Meckenstock, 2006).
+ CH3-X + X
6.4 Chlorinated aliphatics
6.4.1 Degradation of volatile chlorinated aliphatics in groundwater
The chlorinated aliphatic compounds are a diverse group of chemicals. They are produced on a large scale for various industrial activities, but may also have a natural origin. The chlorinated aliphatics are relatively mobile in groundwater and belong, together with the low molecular weight hydrocarbons (BTEX, alkanes) to the most frequently detected pollutants in the subsurface (Committee on Intrinsic Remediation, National Research Council, 2000). The chlorinated aliphatics include many toxic and carcinogenic compounds and are considered to belong to the most serious groundwater pollutants worldwide.
Different mechanisms are involved in the microbial degradation of chlorinated aliphatic compounds (Bradley, 2003). Oxidative degradation pathways dominate in oxygen and nitrate containing groundwater. Anaerobic oxidation of chlorinated alphatics with iron (III) or manganese(IV) as electron acceptor is also an important natural attenuation process. The end products of (anaerobic) oxidation of chlorinated aliphatics are CO2, H2O and chloride. Under more reducing sulfidogenic and methanogenic conditions, reductive pathways are dominant.
Reductive dechlorination reactions usually involve the substitution of a chloride (Cl-) for a proton (H+) on the aliphatic compound. Sequential dechlorination reactions result in the production of partially or completely dechlorinated aliphatic compounds. Chlorinated alkanes can also be dechlorinated through removal of two chlorides in one step (dihaloelimination) from two adjacent carbon atoms. This results in the formation of a double C=C bond. During
reductive dechlorination the chlorinated aliphatics act as an electron acceptor. This implies that the reductive dechlorination processes are fuelled by an electron donor such as H2, naturally occurring carbon compounds or petroleum hydrocarbons (Christensen, 2002). Fermentative pathways have been proposed as a third mechanism for anaerobic degradation of a number of chlorinated aliphatic pollutants. Although the information on fermentative degradation is still relatively limited, this mechanism may play a role in groundwater with limiting amounts of electron acceptors and electron donor substrates.
The relative rates of oxidative degradation of chlorinated carbon compounds generally decrease with increasing degree of chlorination. In contrast, reductive dechlorination of poly-chlorinated compounds proceeds more readily than that of low or mono-chlorinated analogues. Therefore the poly-chlorinated aliphatics are best dechlorinated in anaerobic, reducing groundwater, whereas their (partially) dechlorinated are more rapidly degraded under oxic conditions.
6.4.2 Trichloromethane
Chemical Name: trichloromethane, chloroform, methane trichloride (CHCl3) CAS Number: 67-66-6
Properties: Trichloromethane: Molecular mass: 119.4 g/mol; Solubility in water: 8.2 g.l-1 at 25ºC; vapour pressure: 155 (mmHg) at 20°C; log KOW: 1.97; KOC: 63.4 - 86.7
Use: Trichloromethane is the technical name for chloroform. Chloroform is mainly used in the European Union to the extent of 90-95 % as feedstock for the manufacture of other chemicals.
The main application is the manufacture of HCFC 22 and through this substance; chloroform is also an important building block for fluorinated polymers and copolymers. Other applications as feedstock are in the synthesis of dyestuffs, pharmaceutical products and pesticides. Chloroform is also used as a solvent, for example in the extraction of penicillin and
other antibiotics, and for pesticides, fats, oils, rubbers, alkaloids and waxes. In the past, it is used as a anaesthetic. Trichloromethane is also degradation product from carbon tetrachloride.
Environmental exposure and observed groundwater contamination: Trichloromethane is observed in landfill leachates in a concentration range of 1-70 µg/l (Kjeldsen, 2002).
Degradation: There are currently no micro-organisms known that are capable of using trichloromethane as a sole substrate for growth and/or energy production. Nevertheless,
trichloromethane can be transformed cometabolically by various aerobic and anaerobic bacteria.
Under aerobic conditions trichloromethane is effectively be oxidised by monooxygenase enzymes of bacteria which use growth substrates as methane, butane or ammonia (Arp et al., 2001). The soluble methane monooxygenases of methanotrophic or butane-grown bacteria oxidise trichloromethane to various metabolites, which are further degraded to CO2 and HCl.
The relative importance of aerobic cometabolic degradation of trichloromethane in natural groundwater systems is largely unknown. Obviously, this process requires the combined presence of oxygen and oxygenase-inducing substrates in sufficient quantities.
Under anoxic conditions trichloromethane is cometabolically transformed by various anaerobic bacteria. Examples are certain species of methanogenic, sulphate reducing, nitrate reducing, fermentative and specific dechlorinating bacteria. They form a variety of products, including reductive the dechlorination products dichloromethane and chloromethane, CO2, and other unidentified products (Mohn and Tiedje, 1992; Yu and Smith, 2000). Other anaerobic transformation products of trichloromethane are carbon disulphide and carbon monoxide.
In one landfill leachate plume study of XOC’s degradation under nitrate reducing conditions and one study under aerobic conditions trichloromethane was reported as not degraded (Christensen, 2002).
On a site contaminated with trichloromethane and 1,2-dichloroethane, trichloromethane rapidly degraded to dichloromethane near the source area where the groundwater was anaerobic (Cox et al., 1998). Further down-gradient from the source area the groundwater redox conditions became oxic and the remaining dichloromethane was cometabolised in the presence of methane with a half-life of 143 days.
US-EPA (1999) reported as most likely first order biodegradation rate for trichloromethane of 0.0315 (1/day) under anaerobic conditions, together with a maximum of 0.25 and a standard deviation of 0.0884. This estimate is based on 11 field and laboratory studies with temperature >
15°C, pH <6 to 8 and undefined redox conditions.
6.4.3 Dichloromethane
Chemical Name: Dichloromethane, methylene dichloride, methylene chloride (C2H2Cl2) CAS Number: 75-09-2
Properties: Dichloromethane: Molecular mass: 84.9 g/mol; Solubility in water: 13.2 g.l-1 at 25ºC; vapour pressure: 350 (mm Hg) at 20°C; log KOW: 1.25 (calculated); KOC: 1.68.
Use: Dichloromethane is used as: solvent, degreasing agent, paint remover, pressure mediator in aerosols, degradation product of tetra- and trichloromethane
According to the Euro Chlor (1999), the uses of dichloromethane are:
• For the pharmaceutical industry (30%): where dichloromethane is used as solvent for chemical reactions, purification and isolation of intermediates or products.
• For paint stripping (19%): Dichloromethane based paint strippers normally consist of 70-90% dichloromethane along with other organic solvents, e.g. ethanol, surfactants, emulsifiers and alkaline and/or acid activators. These products have several crucial advantages over other coating removal methods, as example, non flammable, reasonable price, universally suitable for all types of coatings, fast acting at room temperature, etc.
• For aerosols (9%): This application began in the mid-1970 to replace CFC.
Dichloromethane is not a propellant itself, but contributes to package homogeneity through its good solvency and reduces the flammability of the propellant hydrocarbon mixture.
• For adhesives (10%): This application uses dichloromethane as a replacement of 1,1,1- trichloroethane as a solvent.
• For other applications (32%) including metal degreasing, foam blowing,
chemicalprocessing (polyurethanes, polycarbonates), secondary refrigerant medium, etc.
Environmental exposure and observed groundwater contamination: Dichloromethane is observed in landfill leachate in a concentrations range of 1.0-827 µg/l (Kjeldsen, 2002).
Degradation: US-EPA (1999) reported as most likely first order biodegradation rate for Dichloromethane of 0.0076 (1/day) under anaerobic conditions. This estimate is based on one field studies under pH neutral, methanogenic conditions.
Dichloromethane can serve as a growth substrate for a number of aerobic and anaerobic bacteria. In addition, various cometabolic transformations have been documented under oxic and anoxic conditions. Aerobic cometabolism involves the oxidation of dichloromethane by enzymes such as methane monooxygenases. Reduced coenzymes of anaerobic methanogenic or acetogenic bacteria cometabolise dichloromethane to chloromethane.
Aerobic facultative methylotrophic bacteria that are able to grow on dichloromethane can easily be isolated from groundwater systems. These bacteria transform dichloromethane to HCl and S- chloromethylglutathione (Fetzner, 1998). The latter compound hydrolyzes to glutathione, chloride and formaldehyde, which is a central metabolite for methylotrophic growth (Figure 6.8). Growth of methylotrophs on dichloromethane with nitrate as electron acceptor has also been described. No scientific literature on iron(III) or manganese(IV) coupled anaerobic oxidation of dichloromethane has been found.
Figure 6.8 Initial reaction of the aerobic degradation of dichloromethane methylotrophic bacteria (http://umbbd.ahc.umn.edu/index.html).
In methanogenic microcosms dichloromethane was degraded to mainly CO2 and CH4
(Freedman and Gossett, 1991). CO2 and acetate were the major degradation products when methanogenesis was inhibited. The enrichment culture could use dichloromethane as the sole growth substrate. The acetogenic bacterium Dehalobacterium formicoaceticum is able to grow on dichlormethane, forming formate, acetate and chloride as end products (Mägli et al., 1998).
US-EPA (1999) reported a order biodegradation rate for dichloromethane of 0.0064 (1/day) under anaerobic conditions, based on 1 field study at 15°C, pH 6-8 and un defined redox conditions.
6.4.4 1,2-Dichloroethane
Chemical Name: 1,2-Dichloroethane: 1,2-DCE; 1,2-Ethylene dichloride; 1,2-Ethylidene dichloride; 2-Dichloroethane; alpha, beta-dichloroethane; borer sol; Brocide; destruxol borer- sol; di-chlor-mulsion; dichloremulsion; Dutch liquid; Dutch oil; EDC; Ethane dichloride;
Ethylene chloride; Ethylene dichloride; Freon 150; glycol dichloride; sym-dichloroethane;
(C2H4Cl2).
CAS Number: 107-06-2
Properties: 1,2-dichloroethane: Molecular mass: 98.96 g/mol; Solubility in water: 8.0 g.l-1 at 25ºC; vapour pressure: 60 (mm Hg) at 20°C; log KOW: 1.48 ; KOC:19-152
Use: Intermediate for production of vinyl chloride, PVC and other chlorinated chemicals; lead scavenger; fumigant; solvent for cosmetics and drugs; carpet cleaner; photography; photocopy.
More than 95 % of the manufactured dichloroethane is transformed into vinyl chloride.
Less than 5 % of the manufactured EDC is used as :
• Raw material for ethyleneamines, trichloroethylene, perchloroethylene
• Extraction and cleaning solvent
• Lead scavenger for gasoline
Environmental exposure and observed groundwater contamination: Dichloroethane is observed in landfill leachates in concentrations <6 µg/l (Kjeldsen, 2002).
Degradation: 1,2-Dichloroethane serves as a growth substrate for a number of aerobic bacteria.
In most of the known species, such as Xanthobacter, Ancylobacter, Rhodococcus and Mycobacterium strains the initial transformation step is a hydrolytic dehalogenation yielding chloroethanol and HCl as products (Janssen et al., 2005). In contrast, Pseudomonas sp. strain DCA1 uses a monooxygenase-mediated transformation yielding the unstable 1,2-
dichloroethanol, which spontaneously decomposes to chloroacetaldehyde (Hage and Hartmans, 1999). Both aerobic pathways completely mineralize chloroacetaldehyde via chloroacetic acid and glycolic acid to CO2 (Figure 6.9). The oxidation of 1,2-dichloroethane is also mediated by monooxygenase enzymes of methanotrophic bacteria, but this is a cometabolic process in which the methanotrophs do not use 1,2-dichloroethane as a growth substrate.
Figure 6.9. Different pathways of aerobic 1,2-dichloroethane degradation (Hage and Hartmans, 1999).
Under anaerobic conditions the following pathways are used for the microbial transformation of 1,2-dichloroethane (Gerritse et al., 1999). Cometabolic dechlorination to mainly ethene is carried out by various methanogenic, acetogenic or sulphate reducing bacteria. Vinyl chloride, chloroethane and ethane may be formed as by products. Anaerobic bacteria capable of
halorespiration using 1,2-dichloroethane as an electron acceptor have also been described (Grostern and Edwards, 2006). The nitrate-coupled oxidation of 1,2-dichloroethane to CO2 and HCl, and the fermentation of 1,2-dichloroethane to ethene, CO2 and HCl are additional
biotransformation pathways under anaerobic conditions (Figure 6.10). There are currently no indications that iron(III) or manganese(IV) coupled anaerobic oxidation of 1,2-dichloroethane is an important process in anaerobic groundwater systems.
1,2-dichloroethane Cl Cl
H C C H H H
H H
C C
H H
ethene electron donor
2[H]
2HCl
H H
C C 1/3 CO2
H H
5/6 ethene 2HCl
2CO2
12/3[H]
12[H]
electron acceptor
NO3- O2
2HCl
Figure 6.10. 1,2-Dichloroethane biotransformation pathways (Gerritse et al., 1999).
US-EPA (1999) reported a first order biodegradation rate for 1,2-dichloroethane of 0.0076 (1/day) under anaerobic conditions. This estimate, corresponding to a half-life of 91 days, is based on one field study under pH neutral, methanogenic conditions. Bosma et al., (1997) reported half-lives over a wide range from less than 1 to over 30 years in an anaerobic aquifer, heavily contaminated with 1,2-dichloroethane. Abiotic transformation of 1,2-dichloroethane, including alkaline hydrolysis to vinyl chloride and hydrolytic substitution to ethylene glycol, also occurs under anaerobic conditions. The half-lives of these processes are generally, in the order of 10 to 70 years (Barbash and Reinhard, 1989; Jeffers et al., 1989). This indicates that microbial transformation is usually the dominating degradation process in groundwater systems.
6.4.5 Trichloroethylene
Chemical Name: Trichloroethylene: 1-chloro-2,2-dichloroethylene; 1,1-dichloro-2-
chloroethylene; 1,2,2-trichloroethylene; 1,1,2-trichloroethene; anamenth; benzinol; blacosolv;
blancosolv; cecolene; chlorilen; chlorylea; chorylen; circosolv; crawhaspol; densinfluat; dow- tri; dukeron; fleck-flip; flock flip; fluate; lanadin; lethurin; narcogen; narkogen; narkosoid;
Nialk; perm-a-chlor; perm-a-clor; petzinol; philex; threthylene; Triad; Trial; triasol; triklone;
Triol; tri-plus; tri-plus m; vestrol; vitran; trieline; Acetylene trichloride; Ethinyl trichloride;
triciene; Westrosol; Chlorylen; Gemalgene; Germalgene; Tri-Clene; Trielene; Trilene;
Trichloran; Trichloren; Algylen; Trimar; Triline; Trethylene; Trichlorathane; Acetylene
trichloride; Ethinyl trichloride; ethylene trichloride; TCE; Tri; trichloroethene; triciene (C2HCl3) CAS Number: 79-01-6
Properties: Molecular mass: 131.39 g/mol; Solubility in water: 1.1 g.l-1 at 25ºC; vapour pressure: 60 (Hg) at 20°C; log KOW: 2.29; log KOC (5 % OC): 2,1 (calculated).
Use: Trichloroethylene is used as a solvent and in manufacturing to clean grease from
machinery. According to the European Chlorinated Solvent Association (ECSA), the major use of trichloroethylene (more than 80 %) is for vapour degreasing and cleaning of metal parts. The introduction of more efficient and closed degreasing equipment has significantly reduced the consumption of TRI during the last decade, and consequently the emissions in the environment and in the workplace. Trichloroethylene is used as a replacement for 1,1,1-trichloroethane (a substance which was phased out in the developed world under the Montreal Protocol at the end of 1995), in vapour degreasing systems. It is not, however, generally recommended or used as a replacement in emissive applications such as cold cleaning.
Trichloroethylene is also used in adhesives and for synthesis in the chemical industry (HFCs for example) and as solvent for various products.
Environmental exposure and observed groundwater contamination: It is one of the most abundant groundwater pollutants.
Degradation: There are no aerobic or anaerobic bacteria known that can use trichloroethylene as a sole substrate for growth and energy production. Under aerobic conditions trichloroethylene is degraded cometabolically by a wide variety of bacteria expressing non-specific monooxygenase or dioxygenase enzymes. Examples are bacteria oxidising methane, ethene, propane, propene, toluene, phenol, ammonium, isoprene and vinyl chloride (Bradley, 2003). Oxidation of trichloroethylene can yield many different compounds, as illustrated in Figure 6.11.
Trichloroethylene epoxide is a major intermediate, which may be detected in groundwater during active aerobic cometabolic degradation. This unstable intermediate spontaneously degrades in dichloroacetate, carbon monoxide, glyoxylate, or formate. The number, type, and proportion of products depend on the local environment. Oxygenase expressing bacteria initiate the pathway, but other organisms may also carry out later steps.
Anaerobic halorespiring bacteria catalyse the reductive dechlorination of trichloroethylene to cis-1,2-dichloroethylene, vinyl chloride, and ethylene (Figure 6.12). Trans-1,2-dichloroethylene and 1,1-dichloroethylene may also be formed in low quantities. During the anaerobic
dechlorination process the intermediates accumulate sequentially in groundwater. This is significant, because especially vinyl chloride is a potent human carcinogen. The reductive dechlorination processes especially in methanogenic and sulphate-reducing groundwater systems. The less chlorinated intermediates, dichloroethylene and vinyl chloride are mainly dechlorinated reductively under methanogenic conditions. They can however also be degraded through anaerobic oxidative or fermentative pathways (Figure 6.13). Although the bacteria which are responsible for these processes are unknown, they may play an important role in the natural attenuation of chlorinated ethylenes in nitrate and iron-reducing groundwater systems.
Figure 6.11. Initial reactions of the cometabolic degradation of trichloroethylene by various aerobic bacteria (http://umbbd.ahc.umn.edu/index.html).
Figure 6.12. Complete reductive dechlorination of chloroethylenes (chloroethenes) under anaerobic reducing conditions (http://umbbd.ahc.umn.edu/index.html).
CO2(CH4?)
VC Fermentative Acetate
acetogenesis
Reductive dechlorination/
Chlororespiration
Methanogenic Humic acid-
reducing NO3-reducing Mn(IV)-reducing
Fe(III)-reducing SO4-reducing
CO2
CO2
CO2 + CH4
Ethene SO4-reducing Methanogenic
CO2
Ethane
Direct oxidation
Fermentation
CO2
Chloroethanol
Figure 6.13. Pathways for anaerobic degradation of vinyl chloride. The dashed lines indicate hypothetic pathways (adapted from Bradley, 2003).
US-EPA (1999) reported as most likely first order biodegradation rate for trichloroethylene of 0.0016 (1/day) under anaerobic conditions, together with a maximum of 0.04 and a standard deviation of 0.00889. This estimate is based on a large number of field studies under a variety of conditions. Higher rates with a mean of 0.003 (1/day), with a maximum of 0.023 and standard deviation of 0.005, were reported for reductive dechlorination of trichloroethylene under anaerobic field in situ conditions (Suarez and Rifai, 1999). This corresponds to a half-life of about 230 days. Under aerobic “cometabolism conditions” in situ studies indicate
trichloroethylene decay rates of 0.948 (1/day), corresponding to a half-life of less than 1 day.
6.4.6 Tetrachloroethylene
Chemical Name: Tetrachloroethylene; tetrachloroethene; 1,1,2,2-Tetrachloroethylene; carbon dichloride; Ethylene tetrachloride; PERC; Perchloroethylene; PCE; PERK; (C2Cl4)
CAS Number: 127-18-4
Properties: Molecular mass: 165.83 g/mol; Solubility in water: 0.1 g.l-1 at 25ºC; vapour pressure: 14 (Hg) at 20°C; log KOW: 2.53-2.88; log KOC (5 % OC): 2.3-2.4
Use: Tetrachloroethylene, also known industrially as perchloroethylene (PCE), is a good solvent used to clean machinery, electronic parts, and clothing.
Environmental exposure and observed groundwater contamination: Tetrachloroethylene is a suspected carcinogen one of the most abundant environmental pollutants of groundwater.
Degradation: Tetrachloroethylene can not be used as a sole substrate for growth of aerobic or anaerobic bacteria. Until recently it was believed to be recalcitrant under oxic conditions, until it was discovered that it can be cometabilically oxidised by the toluene-o-xylene monooxygenase of a Pseudomonas stutzeri strain (Ryoo et al., 2000). Nevertheless, tetrachloroethylene is considered very resistant to oxidative degradation and often found recalcitrant in oxic environments. The anaerobic oxidation of tetrachloroethylene, for example with nitrate, manganese or iron(III) as electron acceptor, has never been reported.
In contrast, the reductive dechlorination of tetrachloroethene to trichloroethene readily occurs in anaerobic reducing groundwater. Methanogenic and sulphate reducing conditions favor this dechlorination, but in a predominantly iron(III) reducing environment the reaction also occurs.
Certain methanogenic, acetogens and sulphate reducing bacteria can slowly cometabolise tetra- and trichloroethylene reductively. Specific dehalorespiring bacteria, such Sulfurospirillum, Dehalobacter, Desulfitobacterium, Desulfuromonas and Dehalococcoides, are however considered to be responsible for chloroethylene dechlorination in contaminated groundwater systems. Most organisms studied convert tetrachloroethylene to trichloroethylene or cis-1,2- dichloroethylene. Dehalococcoides ethenogenes species are the only reported bacteria that can completely dechlorinate tetrachloroethylene to ethelyene.
US-EPA (1999) reported as most likely first order dechlorination rate for tetrachloroethylene to trichloroethylene of 0.00186 (1/day) under anaerobic conditions, together with a maximum of 0.071 and a standard deviation of 0.0223. This estimate is based on a number of field studies under a variety of conditions. A higher rate, 0.010 (1/day) with a maximum of 0.080 and a standard deviation of 0.022 was reported for reductive dechlorination of tretrachloroethylene in field and in situ studies (Suarez and Rifai, 1999). This corresponds to a half-life of about 70 days. Under oxic in situ conditions the degradation rate constant of 0.000 confirmed the persistence of tetrachloroethylene under these conditions.
6.4.7 Hexachlorobutadiene
Chemical Name: 1,1,2,3,4,4-hexachloro-1,3-butadiene (C4Cl6)
CAS Number: 87-68-3
Properties: Hexachlorobutadiene is a clear colourless liquid; Molecular weight: 260.76 g/mol;
Melting point - 21 ºC; Boiling point: 215 ºC; Vapour pressure: 20 Pa at 20 º C; Log KOW: 4.78;
Solubility in water: 3.20 mg/L; Henry’s law constant: 1044 Pa m3/mol.
Use: Hexachlorobutadiene is used mainly as an intermediate in the manufacture of rubber compounds. It is also used in the production of lubricants, as a fluid for gyroscopes, as a heat transfer liquid, as solvent, fungicide and in hydraulic fluids. Hexachlorobutadiene is mainly formed as a by-product during the manufacture of chlorinated hydrocarbons such as tri- and tetrachloroethene and tetrachloromethane. The world annual production was estimated to be 10,000 tonnes in 1982. There is probably no commercial production of HCBD any more (Van de Plassche and Schwegler, 2005).
Environmental exposure and observed groundwater contamination: Emission of HCBD to the environment can occur via:
1. Emission from production of chlorinated hydrocarbons (unintentional release);
2. Emission from disposal of waste from the production of chlorinated hydrocarbons containing HCBD;
3. Emission from the remaining commercial uses;
4. Emission from magnesium production.
Once hexachlorobutadiene is released into the environment, intercompartmental transport will occur chiefly by volatilization from water and soil, adsorption to particulate matter in water and air, and subsequent sedimentation or deposition (Vermeire, 1994). US-EPA reports levels ranging from 0.043 – 0.35 mg/kg in soil (US-EPA, 2003). Hexachlorobutadiene does not migrate rapidly in soil and accumulates in sediment. Therefore, it is not a major thread for contamination of groundwater. Hydrolysis does not occur.
Degradation: Very little information is available on the bacterial transformation of
hexachlorobutadiene in soils, sediments or water. There are currently no bacteria known that can mineralize hexachlorobutadiene. This compound has been considered recalcitrant for a period of 3 years under conditions with oxygen or nitrate as the major electron acceptor (Bosma et al., 1994; Van Agteren et al., 1998). Tabak et al. (1981) observed complete loss of
hexachlorobutadiene (5 and 10 mg/l) in 7 days. The flasks were incubated under an air atmosphere, but the addition of yeast extract and the fact that they were not shaken may have created anaerobic conditions in the water phase. In methanogenic soil columns with Rhine sediment hexachlorobutadiene was reductively dechlorinated via pentachlorobutadiene to mainly 1,2,3,4-tetrachloro-1,3-butadiene (>90%) and traces (<5%) of a trichlorobutadiene (Bosma et al., 1994). This process occurred after an acclimation period of 4 months. This substance – which is known as a antifungal agent according to the authors - may be degraded further aerobically. No half lives are presented.
6.4.8 Short chain chlorinated paraffins (SCCPs) C10-13
Chemical Name: Polychlorinated alkanes (CxH(2x-y+2)Cly), in the case of SCCPs alkanes with C10-13. They are manufactured by chlorination of liquid n-alkanes or paraffin wax and contain from 30 to 70% chlorine. The products are often divided in three groups depending on chain
length: short chain (C10 – C13), medium (C14 –C17) and long (C18 – C30) chain lengths. Only the SCCPs are usually considered as priority substances.
CAS Number: 85535-84-8
C10H17Cl5 C13H22Cl6
Properties: They are largely depending on the chlorine content. Based on the EU risk
assessment, these are the values for SCCPs: water solubility: 150 to 470 µg.l-1 at 20°C; vapour pressure: 1.6 x 10-4 mm Hg (0.021 Pa) at 40°C (with predicted range for C10-30% chlorine to C13-70% chlorine of 3.3 x 10-4 to 6.0 x 10-8 mm Hg at 20°C); log KOW: 4.39 - 8.69.
Uses: Chlorinated paraffins, complex mixtures of straight chain chlorinated hydrocarbon molecules with a range of chain lengths (short C10-13, intermediate C14-17 and long C18-30) and degrees of chlorination (between 40 - 70 % weight basis) were first produced as extreme pressure additives around 1930. Over 200 commercial formulations with a range of physical and chemical properties exist which make them useful in a wide range of applications, such as secondary plasticizers in PVC and other plastics (C14-17), extreme pressure additives (all chain lenghts), flame retardants (all chain lenghts), sealants (all chain lenghts) and paints (all chain lenghts). The chlorinated paraffins also impart a number of technical benefits, of which the most significant is the enhancement of flame retardant properties and extreme pressure lubrication.
The widespread, numerous contemporary uses of chlorinated paraffins result in the major source of environmental contamination, particularly to the aquatic environment. They may be released into the environment from improperly disposed metal-working fluids containing chlorinated paraffins or from polymers containing chlorinated paraffins. Loss of chlorinated paraffins by leaching from paints and coatings may also contribute to environmental contamination. Risk assessment and risk management showed a need for regulating the Short Chained Chlorinated Paraffins (SCCPs) and restrict their use in main fields of current applications. OSPAR and the EU–Programme on Existing Chemicals suggested these measures and legislation has been issued by the EU-Commission (Directive 2002/45/EC, banning the use of SCCPs in metalworking fluids and fat-liquoring of leather).
Environmental exposure and observed groundwater contamination: CPs may be released into the environment from improperly disposed metal-working fluids or polymers containing chlorinated paraffins. Loss of chlorinated paraffins by leaching from paints and coatings may also contribute to environmental contamination.
Degradation: Chlorinated paraffins are not readily biodegradable. Short chain CPs with less than 50% (w/w) chlorine content seem to be degraded under aerobic conditions. The more highly chlorinated and the medium and long chain products are degraded more slowly. Certain bacteria have also been shown to dechlorinate SCCPs with high chlorine contents in a cometabolic process. A Rhodococcus strain was able to use various chlorinated paraffins with chlorine content <50% as the sole carbon and energy source (Allpress and Gowland, 1999). Therefore
biodegradation of these compounds may also be expected to occur in an oxic environment.
Chlorinated parrafins with chlorine content >58% were not used as a growth substrate by this bacterium. No information on the anaerobic biodegradation of SCCPs is available (European Commission, 2005). SCCP residues have been detected in sediment cores dating back to the 1920s and 1930s. From observations like this, it can be concluded that those components degraded very slowly.
6.5 Chlorinated aromatics
6.5.1 Hexachlorobenzene
Chemical Name: Hexaclorobenzene (C6Cl6) CAS Number: 118-74-1
Properties: Solubility in water: 50 µg.l-1 at 20°C; vapour pressure: 1.09 x 10-5 mm Hg at 20°C;
log KOW: 3.93-6.42; KOC: 36,308 (3,000-180,000).
Uses: Hexachlorobenzene (HCB) is a fungicide that was first introduced in 1945 for seed treatments of grain crops (UNEP, 2002). HCB is also a by-product of the manufacture of industrial chemicals including carbon tetrachloride, perchlorethylene, trichloroethylene and pentachlorbenzene. It is a known impurity in several pesticide formulations, including
pentachlorophenol and dicloram and may be present as an impurity in others. HCB may still be found in the food chain from its former use as a pesticide, however, the main source of release into the environment is now probably as a by-product of certain industrial processes such as aluminium smelting and some older technologies of production of perchloroethylene and vinyl chloride monomer. Some aluminium smelters use hexachloroethane (HCE) gas to remove hydrogen gas from the molten aluminium before it solidifies. However, HCE gas can form HCB in the smelter. Manufacture of vinyl chloride monomer and volatile halocarbons is known to produce HCB as a by-product. Another potential source of HCB is the usage of
pentachlorophenol (PCP), which can be contaminated with HCB during manufacture, and the electrolytic production of magnesium and aluminium.
The use of HCB in such applications was discontinued in many countries in the 1970s owing to concerns about adverse effects on the environment and human health. HCB was also a by- product of the manufacture of industrial chemicals including carbon tetrachloride,
perchlorethylene, trichloroethylene and pentachlorbenzene, this problem already does not exist in main countries of Europe. It is a known impurity in several pesticide formulations, including pentachlorophenol and dicloram and may be present as an impurity in others.
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