Treatment of Contaminated Soil

Anthropogenic activities have resulted in contaminated soils covering significant areas of land. In the Eighties people recognised the size and the consequences of this problem. Initially, treatment and remediation processes were developed mainly by the industry. Very often those approaches were very pragmatic and there was a lack of a scientific basis for the processes, and a need for optimisation and further development. This gap has been closed in the meantime, but there is of course a need for further investigations.

The relationship of our society towards soil and its functions can be seen high?lighted on the photograph showing the festive ceremony of the turning of the first sod in Leipzig, Germany. There is the pleasure to have finally achieved the use of a site for the construction of an industrial production plant. The participants have equipped themselves with high yellow rubber boots. They obviously associated soil with dirt and mud. Shoes and trousers had to be protected against the envi?ronmental medium soil in its natural form. So far concerning the introduction of a highlight for soil consciousness in politics and society.

Twelve years ago, when the investigations within the research centre “treatment of contaminated soils” began, it was the common consensus that decontamination of contaminated soils is preferred to protection measures like cover systems. Meanwhile the situation has changed thoroughly, because in Germany the federal Soil Protection Act (BBodSchG) and the federal Soil Protection and Contaminated Sites Ordinance (BBodSchV) came into force. The opinion of practitioners is that there is no preference of decontamination to protection measures. With few exceptions, cheaper methods of remediation are available in accordance with the law, so that the title of the paper is somewhat rhetorical.

Soil and sediment pollution is a serious environmental problem in The Nether?lands. Over the past 10 to 15 years, awareness of the problem, and the policy and strategy to tackle the problem have radically changed. Initially the approach to tackle the problem of polluted soils was primarily focused on the clean-up of soil after excavation. This approach was consistent with government policy of that time, whose aim was the maintenance or restoration of soil multifunctionality. The result was the development and application of intensive and relatively expensive methods such as thermal treatment and soil washing/wet classification, that were capable of achieving the target values set for clean soil. At that time biological treatment methods and in-situ treatment were not considered feasible. However, inventories revealed that the polluted sites in The Netherlands numbered approxi?mately 100,000 and contained more than 200 million tons of contaminated soil. Moreover, the problems associated with cleaning up these sites were found to be highly diverse. It became clear that for both technical and financial reasons, it would be impossible to clean up all these sites to the target values. This realisation resulted in a shift in policy. Rather than try to clean-up all sites completely, the approach shifted to reducing the risk to humans and the ecosystem. Due to this shift in policy, other techniques which were not able to clean-up the soil to the target values were becoming important for practical application. The result was, among other things, an increasing application of in-situ treatment techniques, biological treatment techniques and isolation methods. However, this approach did not fully solve the problem, mainly due to the high costs. This has recently led to a second shift in soil remediation policy. This second shift has resulted in a remedia?tion approach that now also includes a direct link with the intended use of the remediated site. Also, the financial basis for the remediation activity has been broadened by integration of soil remediation into other social and industrial activi?ties such as building, residing, and rural development. Altering the system of fi?nancing will increase the market dynamics of remediation activities.

To determine the contaminating substances it is necessary to validate the analysis of contaminated soil, and it is also necessary to monitor the change in composition of such soil during decontamination procedures by using suitable analytical methods.

In the majority of cases the currently contaminated and problem sites result from a pollution with mineral oils and their derived products. For the investigation and the registration of such mineral oil contaminated materials, as well as for the selection of suitable measures of treatment and remediation, the physico-chemical analysis is a major requirement.

Chemical analysis takes a major part in the treatment of contaminated soils. Results from analyses are called in for the determination and subsequent assessment of contaminated soils. Furthermore they give information about the success of a soil treatment measure and ultimately indicate the aim of the remediation measure.

Nuclear Magnetic Resonance Spectroscopy (NMR) is the most prominent method to study structural features of many types of chemical substances, especially organic compounds. Consequently this method has been a standard tool to gain insight to the structures of fulvic and humic acids and other humic fractions for many years. From the viewpoint of NMR, two different targets will be discussed: the investigation of soluble compounds with high resolution NMR and the study of insoluble material with Solid State NMR. Good compilations of both types of investigation are given elsewhere (Nanny et al. 1997; Wershaw and Mikita 1987). For technical and practical reasons, this choice has important consequences for studies of fulvic and humic acids. This contribution concentrates on Solid State NMR studies.

The use of stable isotopes, especially 13C, is well established for investigation of metabolic pathways or biosysntheses, as a tool in diagnostic medicine and to generally trace the fate of organic compounds in the environment. Excellent techniques to monitor the isotope in labelled compounds include nuclear magnetic resonance (13C-NMR-spectroscopy) (Marshall 1983; Breitmaier and Voelter 1987) and (combustion) isotope-ratio-monitoring mass spectrometry (IRM-MS) (Merrit et al. 1994; Wong et al. 1995; Newman 1996; Brenna 1999).

The contamination of soil from former industrial sites varies across all kinds of inorganic and organic chemical products, and thus different analytical instruments (SITE 1996; Overton et al. 1996; Matz 1993b) and procedures are required to determine the toxic potential of a soil sample. The contaminating chemicals vary over an extremely wide range of volatility and solubility, they are typically distributed inhomogeneously and the soil material is heterogeneous. This variability prompts the following question, that should more often be asked by the decision-makers dealing with the treatment of large amounts of contaminated soil:

Ecotoxicology has a long history of engaging in the development of biological assessment tools applicable to ecological hazard and risk evaluation. The Proceedings of the 1st International Conference “Environmental Bioassay Techniques and their Application” contained 67 papers, but only two were dealing with soils (Munawar et al. 1989). While numerous compilations of aquatic and soil test methods are readily available for evaluating toxicity as an ecological effect, fewer test methods are readily accessible for evaluating soil contamination. Here, as a result of a 12 year-study, we present biological assessment tools applicable to soil contamination evaluation, with a central focus on methods that are currently available for evaluating soil toxicity. Soil toxicity is one potential indicator of adverse effects associated with soils, caused by chemicals released to the environment.

It is the aim of the Federal Soil Protection Act (Bundes-Bodenschutzgesetz, BBodSchG 1998) to protect or to restore the functions of the soil in a sustainable way. According to the BBodSchG (1998) the natural (ecological) functions of the soil include: the basis for life and habitat, for human beings, animals, plants and soil organisms (in one word: the biospherical function),as part of natural systems, especially as part of water and nutrient cycles,as a medium for decomposition, balance and restoration as a result of its filtering, buffering and substance-converting properties, especially for groundwater protection,

In the last two decades it has been shown that many pollutant compounds being found in soils or aqueous ecosystems may potentially be transformed by microorganisms. Transformation may be either completely into carbon dioxide and water, or at least into non-toxic metabolites (Klein 2000). These promising research results inspired many remediation companies to set up particular bioremediation approaches for the clean-up of such contaminated areas. The bio-enthusiasm of the early years, however, is now followed by a more realistic and sometimes even sceptical view of bioremediation. The major reason for this turn-around is that it has now become clear that results being obtained in the laboratory with artificially contaminated soils do not necessarily indicate what may happen actually in the field with soil from contaminated sites. With hydrophobic pollutants like PAH (Bossert et al. 1984; Erickson et al. 1993; Schaefer et al. 1995; Weissenfels et al. 1992) or some sorts of mineral oil (Angehrn et al. 1997; Bossert et al. 1984; Riis et al. 1998) in particular, it has been observed that even the degradation of compounds being completely mineralisable in the lab-culture may be incomplete in practical field bioremediation. Considerable residual concentrations of analytically detectable pollutants in the soil are subsequently left behind. An example for such a typical “hockey-stick-kinetic” (a term coined by M. Alexander, see Chapter 14) is shown in fig. 13.1.

The pollutants present in soil and sediments usually are not recent introductions, rather they were typically added one, two and frequently more than three decades ago. This fact was one of the concerns as we devised our initial studies. The studies were also prompted by a number of long- and medium-term field studies that showed that persistent pesticides (such as DDT, dieldrin, heptachlor and kepone) and a number of other compounds initially disappear from soil at reasonable rates, but the rate subsequently slows appreciably. Indeed, the subsequent disappearance is often so slow that the rate of decline in concentration sometimes cannot be estimated. Because the compounds were initially subject to biodegradation by microorganisms (as well as loss by volatilisation or other abiotic processes in some instances), clearly something was occurring in soil that was slowly but progressively making the compounds less and less bioavailable, at least to microorganisms (Alexander 1995).

Bioremediation of polluted soil mostly depends on indigenous micro-organisms. The principle of engineered bioremediation is to stimulate those micro-organisms, which are able to degrade xenobiotic substrates (Aelion et al. 1987). The most often reported genera in degradation of anthropogenic pollutants include Ralstonia,Burkholderia, Comamonas, Arthrobacter, Mycobacterium,Nocardia, fluorescent Pseudomonas,Rhodococcus, and Sphingomonas (Haggblom and Valo 1995; Neilson 1995; Commandeur et al. 1995). In response to the introduction of electron acceptors (like oxygen) or nutrients to soil, specific degrader organisms will multiply. However, in spite of a potentially metabolically active biomass, in-situ bioremediation of soils polluted with hydrophobic organic pollutants (HOC) frequently results in slow pollutant degradation rates, high residual concentrations and, as a consequence, in limited clean-up efficiencies (Zhang et al. 1998; Luthy et al. 1994). The unequal spatial distribution of micro-organisms and pollutants in combination with physically retarded substrate diffusion are nowadays generally accepted as key limiting factors for efficient biodegradation of hydrophobic contaminants in soil.

All anthropogenic organic chemicals form non-extractable residues to some extent when entering soils. This well-known phenomenon has been studied for many years especially in the field of soil agrochemistry. Similar processes have been observed during the bioremediation of oil-contaminated soils (see table 16.3) and the residue formation of toxic and carcinogenic polycyclic aromatic hydrocarbons is of particular concern. Apart from mineralisation, the formation of non-extractable residues is certainly the major sink of anthropogenic pollutants in soils. However, macromolecular non-extractable xenobiotic residues are difficult to examine by conventional analytical techniques. Therefore, formation, structure and significance of these macromolecular residues in the environment are still unknown today.

The formation of bound residues has intensively been studied in agrochemistry with the main focus on the fate of pesticides in soil. It is assumed that nearly all xenobiotics entering the soil environment build up these residues, and for many classes this has already been proven (Calderbank 1989).

The depletion of contaminants in soil is not only based on degradation or mineralisation, but also on the fact that a fixation or immobilisation of the xenobiotic substances as bound residues takes place within the soil matrix. This binding of organic contaminants (called a humification process if bound to soil humus) can reduce the bioavailable and analytically detectable part of the xenobiotics. The binding was investigated in detail by the use of 14C-labeled substances for PAH and TNT in the last decades and this immobilisation process has been proposed as a remediation measure. The intentional humification process may be achieved by adjusting bioremediation process parameters such as the supplementation of soil with organic substances (compost etc.) or by changing the incubation conditions (anaerobic and aerobic phases).

Currently, soils from military sites, former ammunition factories and similar sites are simply dumped or remediated by thermal treatment. Low cost alternative microbiological remediation techniques based on humification processes have now been developed for sustainable soil treatment.

Hydrophobic compounds, such as polycyclic aromatic hydrocarbons (PAHs), mineral oils, or halogenated chemicals, represent pollutants of high ecotoxicological relevance at many contaminated sites. Most of these pollutants are biodegradable (Cerniglia 1992; van Hylckama Vlieg and Janssen 2000; Wischnak and Müller 2000) but their rate of biodegradation is limited by low bioavailability. In general, microbial uptake and degradation of pollutants predominantly occurs in the aqueous phase. Pollutants present as crystals (Stucki and Alexander 1987; Tiehm 1994) or in nonaqueous phase liquids (Mukherji and Weber Jr. 1998), as well as compounds sorbed by organic or inorganic matter (Guerin and Boyd 1997; Harms and Zehnder 1995), first have to be transferred into the aqueous phase before biodegradation is possible (Mahro 2000). Con-sequently, recently published models taking into account mass transfer and micro-biological parameters correlate well with measured biodegradation kinetics of hydrophobic model compounds (Ghoshal and Luthy 1998; Mulder et al. 1998a).

The use of micro organisms with specific properties in the environment is a practice which has been known for about a century. Examples can be found in agriculture, where Rhizobia have been used to enhance nitrogen fixation at the roots of leguminoses or where Bacillus thuringiensis has been used to fight insect pests (Elsas et al. 1991). The inoculation of contaminated soils with bacteria, that have the ability to degrade environmental contaminants, a process called bioaugmentation, is a relatively new technique (Trevors et al. 1994). At present it is still unclear which factors determine the success of such a treatment. This situation prompted us to investigate which prerequisites are required to make bioaugmentation a successful tool in bioremediation of contaminated soils.

Biological treatment of contaminated soil is the most common method of soil treatment during the redevelopment of old sites (ITVA 1997). Biopiles in particular are frequently used due to the simple handling involved and process realisation. The significance of the bioremediation of old sites depends on the reutilisation of the soil material to be treated. The objective should be an increase in the amount of decontaminated soil for ecological and economic reutilisation e.g. as a culture medium in landscape architecture.

Biological treatment is a useful tool for mineral oil contaminations as they are easily biodegradable. Numerous publications about oil degradation are available and many remediations have been carried out. However, to date, there is little information about carbon mass balances and carbon turnover of oil in soil.

In the last decade, attention has been paid to the development of new strategies for the biodegradation of pollutants such as aliphatic and aromatic hydrocarbons. These compounds are usually found in the wastewaters of petrochemical plants or in contaminated soils. The bioremediation of such sites has been investigated using pure and mixed microbial cultures under mesophilic conditions (Liu 1985; Cuno 1996). Due to the limited bioavailability of these compounds at ambient temperatures, their biodegradation is rather limited (Kästner et al. 1993; Wilson and Jones 1992). A promising approach to improve the bioavailability of these hydrophobic compounds without using additional chemicals is to develop biological processes at elevated temperatures. Although the treatment of large amounts of polluted water or soils at temperatures above 60 °C requires additional energy, there are, however, many advantages that make such an approach very attractive.

The mineralisation of soil organic matter depends on the activity of fungi and bacteria. The different organisms degrade and transform the organic matter by various processes resulting in an availability of nutrients as part natural nutrient cycles. The activity of microorganisms, and thus the intensity of mineralisation, is strongly related to the physicochemical environment (pH, temperature, water content a.o.) and the existence of appropriate substrate and energy sources. As the natural degradation of organic matter includes various hydrocarbons, microorganisms are used to purify soils and wastewater on a technical scale. Investigations of these techniques mostly focus on the degree of mineralisation caused by the activity of heterotrophic microorganisms neglecting those autotrophic organisms (e.g. nitrifying bacteria) that are not directly involved in mineralisation. The strongly increased activity of heterotrophic microorganisms in oil contaminated soil may result in a considerable change of environmental conditions for the anaerobic autotrophic nitrifying bacteria, especially with regard to oxygen supply and the availability of NH₃ as electron donator for the first step of nitrification.

Petroleum releases to the environment can cause safety hazards, ecological harm and adverse human health effects, and therefore the treatment of petroleum contaminated sites is necessary (Weismann et al. 1998). Today, the treatment of soils contaminated by petroleum products is one of the most frequently occurring cases in soil remediation, and many different techniques are available to purify petroleum contaminated soils (e.g. Koning et al. 2000a). Within the biological remediation processes in particular, the biopile process has been established as one of the most effective and competitive technologies for the treatment of petroleum contaminated soils. Nevertheless, in order to increase the economic efficiency of the process further developments are crucial (Schulz-Berendt 2000).

In the early Seventies the protection of soil was included in the environmental program of the government of the Federal Republic of Germany (Umweltprogramm der Bundesregierung 1971) as one of the major goals for the environmental policy. Nevertheless, the enormous relevance of the contaminated sites problem was not perceived in its total extent until the registration of contaminated sites began in the early Eighties. By 1993 more than 139,000 presumably contaminated sites were registered in Germany, including about 86,000 abandoned waste disposal sites and about 53,000 closed industrial sites (Franzius 1995).

The physical treatment with respect to soil washing is an established technique for cleaning contaminated soils. Its basic principle is the removal of the fine particle fraction from the soil material, in which the contaminants are predominantly concentrated. The goal of the separation process of the fine fraction is to achieve a lowly contaminated coarse fraction that is economically reusable at a yield which should be as high as possible. Industrial soil washing plants basically consist of a wet liberation step and a classification unit. The wet liberation step, which is carried out by application of mechanical energy through impact, is used to transfer the contaminants from the coarse into the fine particle fraction, and to concentrate the contaminant in the fine part of the solids input. The highly contaminated fines are removed in the classification step by means of screens and hydrocyclones, for example.

In Germany we can now look back on more than a decade of experience in soil washing based on the state of the art of mineral processing techniques (Neeße and Grohs 1990a; Neeße and Grohs 1990b; Neeße and Grohs 1991a; Neeße and Grohs 199 lb; Wilichowski and Werther 1996). In recent years washing technology has changed remarkably. The cost of soil cleaning 10 years ago is now no longer acceptable. As a result of competition with biotechnology, in situ-technologies and natural attenuation, the tonnages sent for soil washing have been reduced dramatically. This is a challenge that can be met only by a new generation of soil washing plants, which is the subject of this paper.

Heavy metal contamination of soils is still an unsolved problem although metals are associated with human life and have been used for thousands of years. Mining activities in particular have led to the spreading of large amounts of heavy metals in the environment. In many industrial processes (e.g., in the plating industry, accumulator production, chlorine-alkali-electrolysis, pesticide production), metals and their compounds are used, produced and subjected to different finishing proc­esses, and then applied in various fields of human life. When introduced into soils, heavy metal compounds are hazardous pollutants because they are not biodegrad­able, toxic at relatively low concentrations, and they may be mobilised under changing physical-chemical conditions like redox potential or pH. Soils have a limited capacity to accumulate pollutants, and if this retention capacity is ex­ceeded, the environment, (e.g. ground and surface water, plants and livestock) is likely to be at further risk.

Thermal treatment of contaminated soil has gained great interest. If biological methods have no sufficient cleaning effect, the soil material must be deposited or incinerated. Thermal treatment of contaminated soil with supercritical water (T_c = 374 °C, P_c = 22.1 MPa), in contrast to incineration, leads to clean soil material without creating nitrogen oxides and sulfur dioxide. Depending on the type of contamination, reaction products are CO₂, H₂O, inorganic acids and highly volatile hydrocarbons. The process is carried out at elevated pressures (25 MPa) and, compared to incineration, at moderate temperatures (375–600 °C). Residence times are short (< 120 s) and reaction products can be controlled by conditions of state relating to the oxidation (Brunner 1994).

The last decade was characterised by intensive efforts to gather information on contaminated sites (such as location, chemical species and scale). For example, in Germany 240,000 to 300,000 relevant sites were identified (Franzius et al. 1995; Umweltbundesamt 2000). The increasing knowledge of the potential risks to mankind and the natural environment has stimulated profound investigations to understand the mechanisms and the procedure related with soil remediation and to develop new processes for decontamination. The direct application of the research results was funded both in the USA and in Europe (U.S. Environmental Protection Agency 1996; Rulkens 2000; Stegmann 2000).

In recent years the prevention and management of waste have become an important issue of environmental protection. In the production of raw materials, industrial processing, and soil washing processes large amounts of fine-grained solids and sludges (dp < 100 µm) are produced. Due to the lack of reprocessing plants working economically these materials are currently being dumped. It is estimated that in Germany 180,000 t of fine-grained residues are generated annually, solely by soil washing processes.

Public discussion on serious environmental problems caused by “the chlorine chemistry” arose with Rachel Carson’s book “Silent Spring”, in which the disastrous ecological aftermath of massive use of persistent polychlorinated hydrocarbons as insecticides was pointed out. As a consequence, production and application of these compounds was banned. However, methods had to be developed not only to dispose of materials that contain these substances, but also to remedy the numerous contaminated sites.

During the 1990s, natural attenuation grew from a laboratory research phenomenon to a commonly used approach for the cleanup of contaminated groundwater (MacDonald 2000). The concept of natural attenuation/intrinsic remediation, relies on natural, subsurface processes rather than traditional, engineered procedures. Applications of natural attenuation concepts typically involve biological, chemical and geotechnical approaches. Common objectives are the characterisation of the site with regard to the efficiency of the expected retardation/degradation mechanisms, proof of applicability of the natural attenuation concept (i.e. time frame) and elucidation of questions about the persistence of pollution sources.

In recent years natural attenuation has gained acceptance in several countries, in particular in the USA, as an approach to groundwater remediation for plumes of petroleum hydrocarbons and, in some cases, chlorinated aliphatic solvents. These are the most frequently found pollutants in groundwater, and natural attenuation is used as a remedy in an increasing number of cases.

Natural Attenuation (NA) has just recently been re-assessed by the US National Research Council (NRC) and is considered as an established remedy with a high likelihood of success (at more than 75% of contaminated sites) for only a few types of organic contaminants such as BTEX, low-molecular-weight alcohols, ketones and esters, and methylene chloride. For inorganics such as metals, metalloids, oxyanions and radionuclides, the likelihood of success is rated as moderate or low (at more than 50% or at less than 25% of contaminated sites, respectively, table 37.1). It is assumed that any given site will have the right conditions for natural attenuation of the particular contaminant. The ratings are based on field evidence and the current understanding of the attenuation processes (MacDonald 2000). It seems that natural attenuation as a remedy for inorganics, in particular, is not given much potential.

The detection limit of thin layer chromatography is at 0.1 µg absolute for squalane and lubricating oil respectively, and at 0.4 µg absolute for diesel fuel. Under the conditions indicated below, this equates to detection levels of 170 mg kg−1 squalane and lubricating oil, and 670 mg kg−1 diesel fuel. Smaller contents can be determined by spreading greater volumes. At a spread volume of 20 µl 8 mg kg−1 lubricating oil and 30 mg kg−1 diesel fuel can be detected.

PAH determination by means of HPTLC is a screening method for the determination of the PAH in soil materials. PAH-contents of more than 1.6 mg kg−1 can be determined under the conditions indicated under 39.5 and 39.6. Lower concentrations can be determined by concentration of the eluate and spreading of greater volumes.

Due partly to careless handling of mineral oils and mineral oil products during transport, storage and disposal, soil is often contaminated and ground water threatened. To be able to assess the extent of the pollution, to identify the contaminants and to evaluate the efficiency of a certain treatment process, extensive chemical analyses are necessary.

For the set-up of a remediation concept, suitable field simulation pre-tests on a laboratory, as well as on a technical scale, should principally be carried out. The application of test systems and laboratory reactors is also necessary for basic examination and a general understanding of the degradation processes within soils. When conceptualising biological remediation measures the respective laboratory scale pre-tests should be designed in a way, that answers the questions posed concerning the possibility to comply with the respective microbial, physiochemical and process-geological conditions. The most important, but also most difficult, task is the evaluation of the possibility for remediation of soils. Laboratory results about microbial degradation of relevant contaminants in problem sites have to be transferable into practice and consequently provide data about degradation. This includes degradation rates, final concentrations that can be yielded, formed metabolites and end products respectively. They should be selected between different, complex test systems according to the initial question.

Electrolysis was carried out at 40°C in a batch cell with a volume of 2×100 ml. The cathodic and anodic compartment were each filled with 100 ml methanol containing 0.1 M tetraethylammonium bromide as the supporting electrolyte. After the addition of 106 mg (0.26 mmol) of octachloronaphthalene (Jakobsson et al. 1992) and 100 mg (0.35 mmol) of the mediator 2 (Uchino et al. 1975) to the catholyte, nitrogen was bubbled through the solution to avoid any oxidation of the reduced mediator by oxygen. In order to achieve sufficient turn-over rates the potential was adjusted to −1.40 V vs. Ag/AgBr. The progress of the electrolysis was monitored by GC. For work-up the catholyte was poured into 100 ml water, acidified with half conc. hydrochloric acid and extracted twice with 150 ml hexane each time. The organic layers were dried with magnesium sulfate. The contents of naphthalene derivatives were determined per GC. The solvent was evaporated in vacuo and the products were weighed and analysed by GC. After the passage of 11700 A s at a current of 50-900 mA, a total chemical yield of 91% (current efficiency 3%) was obtained, which consisted of 86% naphthalene, 10% 2,2’-binaphthyl, 3% 1,2′-binaphthyl and < 1% 1,1′-binaphthyl (Nünnecke 2000).