Impact of composting strategies on the treatment of soils contaminated with organic pollutants

Environmental Pollution 112 (2001) 269±283

www.elsevier.com/locate/envpol

Review

Impact of composting strategies on the treatment of soils contaminated with organic pollutants K.T. Semple a,*, B.J. Reid a, T.R. Fermor b a

Department of Environmental Science, Institute of Environmental and Natural Sciences, Lancaster University, Lancaster LA1 4YQ, UK b Department of Plant Pathology and Microbiology, Horticulture Research International, Wellesbourne CV35 9EF, Warwickshire, UK Received 18 June 1999; accepted 16 February 2000

``Capsule'': A review of the e€ectiveness of composting strategies in biomediation of contaminated soils is presented. Abstract Chemical pollution of the environment has become a major source of concern. Studies on degradation of organic compounds have shown that some microorganisms are extremely versatile at catabolizing recalcitrant molecules. By harnessing this catabolic potential, it is possible to bioremediate some chemically contaminated environmental systems. Composting matrices and composts are rich sources of xenobiotic-degrading microorganisms including bacteria, actinomycetes and lignolytic fungi, which can degrade pollutants to innocuous compounds such as carbon dioxide and water. These microorganisms can also biotransform pollutants into less toxic substances and/or lock up pollutants within the organic matrix, thereby reducing pollutant bioavailability. The success or failure of a composting/compost remediation strategy depends however on a number of factors, the most important of which are pollutant bioavailability and biodegradability. This review discusses the interactions of pollutants with soils; look critically at the clean up of soils contaminated with a variety of pollutants using various composting strategies and assess the feasibility of using composting technologies to bioremediate contaminated soil. # 2001 Elsevier Science Ltd. All rights reserved. Keywords: Contamination; Composting; Compost; Bioremediation; Bioavailability

Contents 1. Introduction...........................................................................................................................................................270 2. Composting of pollutants and polluted soils .........................................................................................................271 3. Interactions of organic pollutants with soil ...........................................................................................................272 4. Application of composting bioremediation technologies .......................................................................................274 4.1. Explosives ......................................................................................................................................................274 4.2. Chlorophenols................................................................................................................................................275 4.3. Aromatic hydrocarbons .................................................................................................................................276 4.4. Petroleum hydrocarbons................................................................................................................................278 4.5. Pesticides ........................................................................................................................................................278 5. Treatment of pollutants and polluted matrices with compost ...............................................................................278 5.1. Chlorophenols................................................................................................................................................279 5.2. Volatile organic compounds ..........................................................................................................................279 5.3. Aromatic hydrocarbons .................................................................................................................................279 6. Conclusions............................................................................................................................................................280 References ..................................................................................................................................................................281 * Corresponding author. Tel.: +44-1524-594534; fax: +44-1524593985. E-mail address: [email protected] (K.T. Semple). 0269-7491/01/$ - see front matter # 2001 Elsevier Science Ltd. All rights reserved. PII: S0269-7491(00)00099-3

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1. Introduction The past 200 years has seen a rapid increase in populations world-wide resulting in the need for even greater amounts of fuel and development of industrial chemicals, fertilizers, pesticides and pharmaceuticals to sustain and improve quality of life (Chakrabarty et al., 1988). Although many of these chemicals are utilized or destroyed, a high percentage are released into the air, water and soil, representing a potential environmental hazard (Alexander, 1995). As a result a legacy of contaminated sites requiring attention exists. Environmental pollution has become unacceptable for technological societies as awareness of its e€ects on the environment has increased. Unfortunately, it is not possible to replace all the industrial processes generating polluting wastes with clean alternatives. Therefore, treatment both at source and after release, whether accidental or not, must be considered as alternatives in many cases (Betts, 1991). Current legislation and recent waste management strategies have placed signi®cant emphases on waste minimization, recycling and remediation rather than disposal, which is now perceived as being the least desirable option (Colleran, 1997). The persistence of organoxenobiotics in the environment is a matter of signi®cant public, scien-

ti®c and regulatory concern because of the potential toxicity, mutagenicity, carcinogenicity and ability to bioconcentrate up the trophic ladder. These concerns continue to drive the need for the development and application of remediation techniques (Colleran, 1997). In the past, chemical pollution in soil has been treated using physical and chemical processes that have proven to be expensive (Table 1). Physical and chemical remediation techniques, including removal to land®ll, soil washing, solvent extraction and incineration, will not be discussed in this review. However, Table 1 summarizes and compares the physical and chemical remediation strategies with biological techniques, in terms of how the various treatments a€ect the soil chemically, physically and biologically. Bioremediation is the use of biological treatments, for the clean-up of hazardous chemicals in the environment. This review focuses exclusively on microbiological bioremediation technologies, although in many cases phytoremediation (the use of plants for remediation) has proven successful. At present, employing the biochemical abilities of microorganisms is the most popular strategy for the biological treatment of contaminated soils and waters (Head, 1998). Microorganisms, more so than any other class of organisms, have a unique ability to interact both chemically and physically with a huge

Table 1 E€ects of remediation methods on soil characteristics and the estimated costs of treatment (adapted from Houghton, 1996)a Treatment

E€ects on soil chemistry

E€ects on physical structure

E€ects on microorganisms

Approximate remediation cost (£/tonne)

Removal to land®ll

?

?

?

Up to 100

Solidi®cation Cement and Pozzolan based Lime based Vitri®cation

N N N

N N N

N N N

25±175 25±50 50±525

Physical processes Soil washing Physico-chemical washing Vapour extraction

Y Y Y

N N Y

N N Y

25±150 50±175 75

Chemical processes Solvent extraction Chemical dehalogenation In situ ¯ushing Surface amendments

Y Y Y Y

N N Y Y

? ? ? Y

50±600 175±450 25±80 10±25

Thermal treatment Thermal desorption Incineration

Y N

N N

N N

Biological treatments Windrow turning Land farming Bioventing Bioslurry Biopiles In situ bioremediation

Y Y Y Y Y Y

N N Y N N Y

Y Y Y Y Y Y

a

25±225 50±1200 10±50 10±90 15±75 50±85 15±35 175

N indicates that the above factors will not generally survive in a particular treatment method and Y indicates that they will generally survive. ? indicates that the e€ects are unclear.

K.T. Semple et al. / Environmental Pollution 112 (2001) 269±283

range of both man-made and naturally occurring compounds leading to a structural change to, or the complete degradation of, the target molecule (Head, 1998). The relatively recent development of bioremediation has added to existing clean-up strategies currently available for the restoration and rehabilitation of contaminated sites and can be conducted either in situ or ex situ. This biological strategy is dependent on the catabolic activities of the indigenous micro¯ora, optimizing the conditions in situ for growth and biodegradation. Ex situ treatments involve the physical removal of the contaminated matrix to controlled and contained reactors, compost heaps or lagoons. Many techniques of dispersal, collection, removal, land ®ll disposal and incineration simply dilute or sequester the contaminants or transfer them to another environmental medium. In contrast, bioremediation can be regarded as a more e€ective and environmentally friendly strategy since it results in the partial or complete biotransformation of organoxenobiotics to microbial biomass and stable, innocuous end-products (Colleran, 1997). The acceptance of bioremediation as a viable clean-up strategy, however, in many cases also depends on cost i.e. the method needs to be no more expensive than existing chemical and physical

271

treatments. The thorny issue of the cost of a remediation strategy is highlighted in Table 1. It indicates that the bioremediation strategies listed are competitive in terms of cost as well as in terms of the impact on the contaminated matrices. It is only over the last 3±5 years that the use of composting strategies in biodegradation/bioremediation of organic pollutants has been seriously adopted; as a result there is a lack of general information as well as a limited number of pollutant/pollutant mixtures treated. Pollutants investigated include petroleum hydrocarbons, monoaromatics (benzene and toluene) explosives [2,4,6-trinitrotoluene (TNT)], chlorophenols [pentachlorophenol (PCP)], pesticides [2,4-dichlorophenoxyacetic acid (2,4D) and diazinon] and polycyclic aromatic hydrocarbons (PAHs) (anthracene, phenanthrene benz[a]anthracene and benzo[a]pyrene). PAHs are perhaps the most studied of these contaminants. The chemical structures of these pollutants are shown in Fig. 1. The aims of this review are to discuss the interactions of pollutants with soils, looking critically at the e€ectiveness of remediation of contaminated soils using composting technologies and to address the issue of augmentation of contaminated soils with composted materials. Additionally, other important issues such as bioavailability, metabolite production and toxicity and their bearing on remediation strategies will be discussed. By addressing these issues this paper aims to critically assess whether composting is a viable technology for use in bioremediation strategies. 2. Composting of pollutants and polluted soils

Fig. 1. Soil pollutants that have been treated using composting/compost remediation strategies.

Composting is an aerobic process that relies on the actions of microorganisms to degrade organic materials, resulting in the thermogenesis and production of organic and inorganic compounds. The metabolically generated heat is trapped within the compost matrix, which leads to elevations in temperature, a characteristic of composting (Williams et al., 1992). Further, Fogarty and Tuovinen (1991) divided the composting process into four major microbiological stages in relation to temperature: mesophilic, thermophilic, cooling and maturation. With these changes in temperature, there are related changes in the structure of the microbial community. With increases in the respiratory activity, there is an increase in temperature resulting in a decrease in mesophilic microbes and an increase in thermophiles and it is at these higher temperatures (45±65 C) that most of the microbial decomposition and biomass formation takes place (Fogarty and Tuovinen, 1991). In the third phase, there is a cooling e€ect due to the decrease in microbial activity as most of the utilizable organic carbon has been removed, resulting in an increase in mesophilic microorganisms (Fogarty and Tuovinen, 1991).

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Fig. 2. Summary of the environmental fates of organic pollutants in soil.

There are a variety of composting systems including in-ground trenches, rotating drums, circular tanks, open bins, silos, windrows and open piles. Initially, many of these systems were developed for the stabilization of sewage sludge waste relying more on aerobic microbial activity rather than anaerobiosis because the latter leads to the formation of H2S and SO2 (Miller et al., 1991). Also, aerobic composting gives a higher degree of decomposition for most compounds. Thus, most composting systems utilize bulking agents (such as bark chips, straw and chopped sugar beet), which increases the porosity and, therefore, aerobicity of the medium under treatment and decreases the moisture levels. Where composting contaminated land with organic bulking agents, the thermophilic state is usually not obtained and, therefore, the temperature does not exceed 45 C. It is important to di€erentiate at the onset the dissimilarity between compost and composting. Composting is the process by which compost is produced, i.e. the maturation of, for example straw and manure. Compost is the resultant product of composting, with the exception of horticultural potting composts. Thus, a composting bioremediation strategy relies on the addition of compost's primary ingredients to contaminated soil, wherein the compost matures in the presence of the contaminated soil. In contrast, compost can be added to contaminated soil after its maturation. These distinct approaches are discussed separately. 3. Interactions of organic pollutants with soil Before addressing the role of composting strategies for the bioremediation of contaminated soils, it would be useful to discuss the limiting factors governing bioremediation, namely soil±pollutant interactions. Possible fates of pollutants and their breakdown metabolites

entering soil environments include volatilization to air, biodegradation, transfer to organisms, binding to soil and leaching into groundwater (Cerniglia, 1992; Jones et al., 1996). These processes are summarized in Fig. 2. The fate and behaviour of organic pollutants in soil is governed by many di€erent factors including soil characteristics, chemical properties and environmental factors such as temperature and precipitation. The persistence of certain organic pollutants in soil has been proposed to be related to compound hydrophobicity (Cerniglia, 1992). Pollutants generally dissipate from soils in a biphasic manner, i.e. a preliminary short period of rapid loss is followed by a subsequent longer period of slower loss (Jones et al., 1996). Pollutant volatility, hydrophobicity and anity for organic matter govern the relative importance of each phase. In addition to removal/loss processes, intra-soil processing of pollutants also occurs. These processes reduce pollutant bioavailability and promote the formation of non-bioavailable residues with time. It has not been established in the literature that such non-bioavailable residues (at time of assessment) are permanently immobilised, future release must not, therefore, be ruled out as a possibility. It has been observed that as the length of time an organic chemical remains in contact with soil increases, the ability for that chemical to be degraded by microorganisms decreases (Hatzinger and Alexander, 1995). This decrease in compound availability, with time, has been termed `ageing'. Bioavailability is perhaps the most important single factor governing the success of many bioremediation strategies; the processes that control it (i.e. ageing) must, therefore, be discussed. The nature and extent of ageing is dependent upon ®rstly, the pollutant's intrinsic properties, fundamental parameters including aqueous solubility, vapour pressure and octanol:water partition coecient (Kow) (Brusseau et al., 1991; Cerniglia, 1992; Jones et al., 1996). Secondly, ageing relates to soil properties including soil organic matter (SOM), both in amount (Hatzinger and Alexander, 1995) and in nature (Piatt and Brusseau, 1998); soil inorganic constituents (Ball and Roberts, 1991a, b; Mader et al., 1997) with particular reference to pore size and structure (Nam and Alexander, 1998). Organic pollutants interaction with soil also occurs through a number of attractive forces, such as dipole± dipole, dipole-induced dipole and hydrogen bonding (Pignatello and Xing, 1996). Sorption is widely accepted as the controlling factor in the ageing process and can occur in a number of ways, including: sorption onto inorganic soil constituents (Ball and Roberts, 1991a; Fu et al., 1994; Burgos et al., 1996) and sorption onto SOM (Brusseau et al., 1991; Fu et al., 1994; Xing and Pignatello, 1997; Schlebaum et al., 1998). Pollutant entrapment occurring from di€usion of compounds into spatially remote areas, such as soil macro and micro pores (Steinberg et al., 1987; Ball and Roberts, 1991b; Beck

K.T. Semple et al. / Environmental Pollution 112 (2001) 269±283

et al., 1995; Burgos et al., 1996; Pignatello and Xing, 1996) and within SOM (Brusseau et al., 1991; Fu et al., 1994; Xing and Pignatello, 1997; Schlebaum et al., 1998), has also been proposed to be a major factor. The in¯uence of SOM has been proposed by many researchers to be the single most signi®cant factor dominating organic pollutant interactions within soil (Brusseau et al., 1991; Cornelissen et al., 1998). Brusseau et al. (1991) provided strong evidence that intra-organic matter di€usion (IOMD) was responsible of the non-equilibrium sorption exhibited by a wide range of hydrophobic organic chemicals. Higher pollutant concentrations have also been shown to enhance sorption to soil (Divincenzo and Sparks, 1997), thus enhancing ageing. Paradoxically, a minimum threshold concentration is required for catabolic induction in degrading microorganisms (Boethling and Alexander, 1979). So while a higher concentration of pollutant is a prerequisite for biodegradation, its availability for degradation is compromised. Low pollutant concentrations have been shown to prevent endpoints being attained in a number of attempted bioremediation strategies (Head, 1998). In contrast to this point, biodegradation of pentachlorophenol has been shown to be promoted by the reversible sorption of the chemical to bulking agents used in composting strategy (Apajalahti and Salkinoja-Salonen, 1984). The authors proposed this to be attributable to detoxi®cation of the surroundings and thus enhanced microbial activity. Reviews on the processes inherent to ageing are provided

273

by Alexander (1995), Loehr and Webster (1997), Pignatello and Xing (1996), Luthy et al. (1997) and Reid et al. (2000). The ultimate result of the ageing processes is the movement of chemicals from accessible soil compartments into less accessible or inaccessible compartments, the result of which is a reduction in bioavailability. It is generally recognised that three soil-associated chemical pools exist after ageing: (1) a fraction which can be rapidly desorbed; (2) a fraction which is more slowly desorbed (Fu et al., 1994; Pignatello and Xing, 1996); and (3) a fraction which has been termed `bound residue' or `non-extractable'. It must be stressed that non-bioavailable residues and bound residue are distinct terms. At a recent workshop, a modi®cation to the existing IUPAC de®nition (initially proposed by Roberts, 1984) of what constitutes a bound residues was proposed. The revised de®nition proposed that: ``bound residues represent compounds in soils, plants or animals which persist in the matrix in the form of the parent substance or its metabolite(s) after extraction. The extraction method must not substantially change the compounds themselves or the structure of the matrix'' (Fuhr et al., 1996). It is important however, to bear in mind that `extractability' of a compound is operationally de®ned by the nature of the extractant and the experimental conditions under which an extraction is carried out (Gevao et al., 2000).

Fig. 3. Summary of the fates of organic pollutants in soil±composting/compost systems.

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Although the amount of pollutant which is potentially available for degradation is of fundamental importance for bioremediation, a more signi®cant factor is the rate at which compounds are transferred to the aqueous phase (Bosma and Harms, 1996; Bosma et al., 1997). Desorption kinetics are widely accepted to be the rate-limiting factor in biodegradation (Stucki and Alexander, 1987; Bosma and Harms, 1996; Pignatello and Xing, 1996; Bosma et al., 1997; Carmichael et al., 1997; Yeom and Ghosh, 1998), even to the extent that they take precedence over intrinsic microbial activity (Bosma et al., 1997). Cornelissen et al. (1998) proposed that by assessing the size of the rapidly desorbable fraction of soil-associated organic compound (using Tenax resin), a measure of bioremediation potential could be obtained. Cornelissen et al. (1998) also indicated that the presence of SOM signi®cantly increased the size of the slowly desorbing pool of PCBs and chlorobenzenes. It is evident that the fate and behaviour of organic compounds within the soil environment are dependent on a complex array of processes, which ultimately govern bioavailability and thereby dictate the feasibility of bioremediation strategies. A summary of some of these processes is provided in Fig. 3. At one extreme, where pollutants are completely available and biodegradable, a bioremediation strategy should prove favourable. At the other extreme, where compounds are non-available and/or recalcitrant in nature, such a strategy may fail. However, at this second extreme, such non-availability and thus decreased toxicity raises the question ``has the remediation endpoint not already been achieved?'' (Bosma et al. 1997). 4. Application of composting bioremediation technologies Composting is the process by which most composts are produced. Thus, a composting bioremediation strategy relies on mixing the primary ingredients of composting with the contaminated soil, wherein as the compost matures, the pollutants are degraded by the active micro¯ora within the mixture. Composting is a relatively new clean-up strategy and because of this, there are a limited number of studies to comment upon. However, the studies described in this and subsequent sections have been dealt with on a pollutant-class basis. 4.1. Explosives Explosives, such as TNT and hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX), have been the subject of intense study as they have accumulated in soils and sediments as a result of manufacture, loading and demilitarization of weapons (Isbister et al., 1984). Several research groups have investigated composting of explosives; e.g. in one of the ®rst studies to address this issue, Kaplan and Kaplan (1982) described the biotransformation of 14C-

labelled TNT by thermophilic microorganisms in a composting environment. In this study, no TNT mineralization was detected, although the formation of 14C-labelled reduction products, including various aminonitrotoluenes, was described. From this study, a signi®cant portion were bound to the humic fraction and the percentage of humicbound 14C-label was found to increase with contact time with the compost. Further, Isbister et al. (1984) investigated composting to degrade or immobilize 14C-labelled TNT and 14C-labelled RDX in contaminated soils. The authors found that the concentrations of both chemicals were quickly reduced over a 6-week composting incubation. RDX was degraded to CO2 and compost leachates were not found to be mutagenic. There was no evidence of mineralization of the aromatic ring of TNT, however 14 C-label associated with TNT was found to be incorporated into humic materials within the composting matrix, suggesting polymerization of 14C-labelled TNT residues to insoluble macromolecules. This is in agreement with Pennington et al. (1995) where no 14C-labelled TNT mineralization or volatilization was detected and the only transformation products were aminodinitrotoluenes. The dominant fate process was incorporation of the 14Clabel into the organic matter fraction, of which 32% of the 14 C-activity was associated with cellulose, 22% with humin and 18% with fulvic acid fractions. Williams et al. (1992) investigated the feasibility of using composting to bioremediate sediments contaminated with TNT and soils contaminated with nitrocellulose under thermophilic and mesophilic conditions. Under thermophilic conditions, extractable TNT and nitrocellulose residues were reduced from 11,840 to 3 mg kgÿ1 and 13,090 to 16 mg kgÿ1, respectively. Under mesophilic conditions, solvent extractable TNT residues were reduced from 11,190 mg kgÿ1 to 50 mg kgÿ1. The authors also calculated the half-lives for TNT, octahydro-1,3,5,7tetranitro-1,3,5,7-tetraazocine (HMX) and RDX in contaminated sediments and nitrocellulose-contaminated soils under thermophilic and mesophilic conditions ®nding that these values were 11.9 and 21.9 days for TNT, 17.3 and 30.1 days for RDX and 22.8 and 42 days for HMX, respectively. This study showed that concentrations of solvent-extractable explosives were signi®cantly reduced during the composting process. The authors suggested that the mechanisms for this were: (1) sorption to the compost matrix; (2) incorporation into environmentally stable molecules such as humic complexes; (3) degradation to CO2, H2O and other inorganic molecules (Williams et al., 1992). Caton et al., (1994) found the 14C-label initially introduced as 14C-TNT to be associated with a number of fractions, after 90 days of composting. They found, 1.2% of the 14C-radiolabel present in the labile fraction, 17.9% associated with the insoluble-particulate fraction, 56.8% associated with the insoluble but hydrolysable fraction and 4.7% associated with the non-extractable fraction. The authors

K.T. Semple et al. / Environmental Pollution 112 (2001) 269±283

described that the majority of the 14C-activity was accountable in the non-extractable but hydrolysable fraction, but were unable to state whether it was the parent compound or breakdown metabolites. Breitung et al. (1996) further investigated the issue of TNT-contaminated soil bioremediation using two different composting regimes. This study aimed to elucidate biodegradation and immobilization processes that led to the detoxi®cation of the contaminated soils. The ®rst composting system was aerated from the start of the incubation, while the second was aerated for 95 days after a 65-day anaerobic phase. In the solely aerated system, there was a rapid decline in extractable TNT (approx. 92% reduction); however, TNT was still detectable after 28 days. In the anaerobic/aerobic system, TNT was almost completely transformed to aminodinitrotoluenes in the anaerobic phase and completely removed from the extractable phase following aeration. Successful bioremediation was further con®rmed in testing the compost water against lux-marked bacteria which showed decrease in toxicity with increased composting time. More recently, Bruns-Nagel et al. (1998) composted TNT-contaminated soil with chopped sugar beet and straw anaerobically for 19 days then aerobically for a further 58 days. Under the initial anoxic conditions, approximately 90% of the TNT was transformed to mono- and diaminonitrotoluene intermediates, which were further transformed under aerobic conditions to putatively less toxic acetylated and formylated metabolites. The authors concluded that they had developed an ecient system for bioremediating TNT-contaminated soils. These studies indicated the formation of TNT transformation products. However, as with other studies discussed, they did not report mineralization of TNT or its metabolites in the composting system described.

275

4.2. Chlorophenols Chlorinated organic compounds represent a signi®cant environmental problem (HaÈggblom, 1992). Included within this class of putative environmental pollutants are the chlorophenols, which have been used extensively in agriculture, industry and public health because of their wide-spectrum biocidal properties (Apajalahti and Salkinoja-Salonen, 1986). In 1985, the world-wide industrial production of pentachlorophenol (PCP) was over 100,000 tonnes, with approximately 80% being used for wood preservation (Wild et al., 1992). As a result, chlorophenol contamination of the environment is far reaching (Valo et al., 1984; HaÈggblom and Valo, 1995). PCP was one of the most commonly used compounds of its class, and is considered a priority toxic pollutant by the US Environmental Protection Agency (Sittig, 1981). The recalcitrance of chlorophenols, including PCP, can be attributed to its chemical structure (Apajalahti and Salkinoja-Salonen, 1984). Owing to the presence of ortho-chlorine atoms relative to the hydroxyl functional group the formation of catechol analogues is prevented. Thus, the principal route of oxidative aromatic ring cleavage is blocked and, as a result, biodegradation inhibited. However, PCP can be degraded through the actions of microorganisms, but there has only been limited success in soil systems using microbial inocula due to the associated toxicity (Apajalahti and SalkinojaSalonen, 1984). In a study carried out by Salkinoja-Salonen et al. (1986), the half-life of PCP was accelerated from 10 to 3 months with the addition of bark chips to a contaminated e‚uent. The addition of a PCP-degrading inoculum further shortened the half-life to less than a week. The promotive role of bark chips was shown to be due to the sorption of PCP by the bark chips, which

Table 2 Removal of chlorophenols from contaminated soils using di€erent composting regimes (adapted from Laine and Jùrgensen (1997) Time (weeks)

Total chlorophenols Soil+bark chips mg kg

0 1 3 5 7 9

43 19 16 10 9 7

ÿ1

soil

Soil+straw compost % Remaining 100 44.2 37.2 23.3 20.9 16.3

mg kg

ÿ1

soil

Soil+remediated soil % Remaining

mg kgÿ1 soil

% Remaining

45 23 21 13 10 10

100 51.1 46.7 28.9 22.2 22.2

43 21 18 11 9 7

100 48.8 41.9 25.6 20.9 16.3

683 233 42 44 42 38

100 34.1 6.1 6.4 6.1 5.6

1108 585 103 53 67 49

100 52.8 9.3 4.8 6.0 4.4

Introduction of highly chlorophenol-contaminated soil 0 4 8 12 16 42

771 203 35 33 34 29

100 26.3 4.5 4.3 4.4 3.8

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detoxi®ed the surroundings for the PCP-degrading bacteria (Apajalahti and Salkinoja-Salonen, 1984). Valo and Salkinoja-Salonen (1986) carried out a ®eld-scale composting study in 50-m3 windrows (using bulking agent, organic matter source). The soil concentrations of chlorophenols were reduced from 212 to 30 mg kgÿ1 in the ®rst summer (4 months) and further reduced to 15 mg kgÿ1 in the second summer. All the chlorophenol congeners were degraded; however, dimeric impurities, such as dioxins (found in the chlorophenol technical mixture), were resistant to the degradative actions of the indigenous micro¯ora. More recently, Laine and Jùrgensen (1997) investigated bench-scale composting of chlorophenol-contaminated soils using di€erent inoculants: mushroom straw compost, remediated soil and indigenous soil micro¯ora. Over a period of 30 days, approximately 50± 60% of [UL-14C]PCP was mineralized in all the composting systems. From this, pilot-scale composting of chlorophenol-contaminated soils (approx. 44 mg kgÿ1) in windrow systems was investigated using the di€erent inoculants. This study showed that 80% of the chlorophenols were removed reaching an acceptable concentration of less than 10 mg kgÿ1 over a period of 2 months (Table 2). At this point, the composting windrows were further spiked with highly contaminated soils (683±1108 mg kgÿ1) and, after a further 3 months of composting, more than 90% of the chlorophenols had been removed, with no di€erences being found between the piles with or without augmentation of compost or remediated soil (Table 2). Laine and Jùrgensen (1997) found that mixing, along with nutrient addition, of the composting piles improved degradation by the indigenous micro¯ora of the contaminated soils. Additionally, when the chlorophenol-contaminated soils were added a second time, the degradation rate was very fast suggesting that the initial 2-month composting period enhanced the catabolic activity within the composting piles. Laine and Jùrgensen (1997) concluded that the degradation of chlorophenols occurred faster in the laboratory systems than in the ®eld, as conditions were more favourable. Chlorophenol removal in the laboratory was approximately 2% dayÿ1 but varied between 0.3 and 1.3% dayÿ1 in the ®eld depending upon the initial chlorophenol concentration. The determining factor was temperature as in the laboratory an ambient temperature of 20 C was maintained. However, suitable temperatures in Finland are dicult to maintain and will, therefore, have a negative impact on the rate of degradation in the ®eld. Additionally, polychlorinated dibenzo-dioxins/furan (PCDD/F) contaminants present within the original technical grade chlorophenol mixture were not degraded during the composting process, suggesting that composting is unsuitable for treating PCDD/F-contaminated soils. It has been proposed that

PCP dimerization, as a result of microbial activity, may result in the formation of chlorinated phenoxyphenols (Fig. 3) or polychlorinated dibenzo-p-dioxins or dibenzofurans (Laine and Jùrgensen, 1997; Laine et al., 1997b). Laine et al. (1997b) also investigated the toxicity of chlorophenol-contaminated soils using Vibrio ®sheri strain NRRL B 1117. As composting of the contaminated soil proceeded at the ®eld-scale, there was a commensurate decrease in the measured toxicity as well as a decrease in chlorophenol concentration. There was no polymerization of the chlorophenols during the composting process and that the existing polymerized fraction was neither degraded nor remobilized. Microbial functionality within this full-scale composting bioremediation system was studied using conventional microbial enumeration, soil respiration and Biolog1 substrate utilization (Laine et al., 1997a). The best indicator of chlorophenol degradation eciency was enumeration of microbes capable of growing on agar with 2 mM PCP as the sole carbon source. It was found that the chlorophenol degraders did not contribute to the substrate utilization measured using Biolog1 or that of the respirometric measurements, but it was fast-growing microbes that were responsible. It was suggested that these microbes originated from the bulking agent, i.e. bark chips and straw compost. 4.3. Aromatic hydrocarbons Traditionally, bioremediation feasibility studies have been carried out on single pollutants, and this is true of composting research also. This is important as it allows the elucidation of the fate processes for speci®c compounds. This has been highlighted in composting strategies which have been used to treat volatile compounds such as benzene and toluene through a process of bio®ltration. BTEX (benzene, toluene, ethyl benzene and the three xylene isomers) compounds are commonly found in petroleum-contaminated sites and are of major concern because of their toxicity and carcinogenicity (EPA, 1995). For example, Matteau and Ramsay (1997) ®rst reported the feasibility of using thermophilic and mesophilic phases during the composting of leaves and alfalfa to biodegrade toluene. Under thermophilic conditions, toluene was degraded at a rate of 110 g mÿ3 hÿ1, whereas under mesophilic conditions the aromatic compound was removed at the reduced rate of 98 g mÿ3 hÿ1. Additionally, benzene was degraded under mesophilic conditions at a similar rate to that of toluene. PAHs also represent a signi®cant environmental risk and human health threat (Cerniglia, 1992), and soils are a major sink for these pollutants (Wild and Jones, 1995). For example, Adenuga et al. (1992) showed that pyrene could be degraded in the composting of soil/sludge mixtures although the rate and extent were not mentioned in this study. The fate of benzo[a]pyrene in soil was investi-

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277

Table 3 Removal (%) and bound residuea formation (%) of benzo[a]pyrene under composting conditions in the presence and absence of Phancrochaete chrysosporium (adapted from McFarland and Qiu, 1995) Incubation time (days)

Uninoculated

Inoculated

Removal

1 7 14 21 28 35 84 91 95

Bound residue formation

Removal

Bound residue formation

(%)

S.D. (%)

(%)

S.D. (%)

(%)

S.D. (%)

(%)

S.D. (%)

21.9 32.3 42.3 49.3 43.5 44.7 61.8 74.3 65.6

0.1 1.9 4.5 4.6 0.7 3.5 18.1 9.1 1.2

2.05 6.07 12.56 13.59 17.20 25.14 38.97 50.72 40.42

0.76 .086 0.69 1.91 1.54 5.4 16.73 9.54 4.0

32.5 44.1 44.6 51.6 60.7 49.5 60.8 58.9 62.8

0.5 13.0 0.5 6.2 3.5 0.8 8.4 3.5 5.9

3.76 13.37 15.67 20.05 36.69 24.24 30.33 35.51 37.58

0.28 9.27 2.51 9.09 3.69 1.17 5.61 5.3 8.22

a Bound residue determined as the non-extractable residue following soxhlet extraction (1:1 vol, methylene chloride:acetone) of the sample in accordance with USEPA method 3540.

gated under a composting regime in the presence and absence of Phanerochaete chrysporium (McFarland and Qiu, 1995). This study showed that although the benzo[a]pyrene appeared to be removed, there was no appreciable di€erence between the uninoculated and inoculated systems with 65.6 and 62.8% removal, respectively, after 95 days (Table 3), although initial rates of removal were faster in the inoculated incubations. Interestingly, analysis of gaseous traps indicated that there was no loss through volatilization or mineralization and that nearly 100% of the benzo[a]pyrene removed was attributable to bound residues as the parent compound (approx. 60%) or as chemical intermediates (Table 3). Further, this study highlighted that the presence of the fungus increased the rate of bound residue formation in the ®rst 30 days of the composting study, where the rate went from 0.73 mg kgÿ1 dayÿ1 in the absence of the fungus to 1.58 mg kgÿ1 dayÿ1 in the presence of the fungus. The authors conclude that the Table 4 Removal (%) of PAHs in creosote during composting (adapted from Civilini, 1994) PAH

Naphthalene Acenaphthene Fluorene Phenanthrene Anthracene Fluoranthene Pyrene Benz[a]anthracene Chrysene

Removal (%) of PAHs during composting

Total removal

5 days

10 days

15 days

93.33 17.40 67.44 75.90 57.14 45.47 55.14 34.18 27.39

5.23 69.93 28.01 18.87 30.32 37.94 32.64 22.15 18.80

0 10.23 3.18 3.21 10.23 9.25 7.80 25.30 41.16

98.56 97.56 98.63 97.98 97.69 92.66 95.58 81.63 87.35

bioaugmentation of a soil-composting system with P.chrysosporium was ine€ective in degrading benzo[a]pyrene during the 95-day incubation. However, in terms of `locking up' the PAH within the compost matrix, this technique proved very successful, although the long-term implications for the fate of benzo[a]pyrene are unknown. There are a few studies that consider the implications of pollutant mixtures. An example is that of the wood preservative creosote, of which approximately 4500 tonnes per annum is used in the USA (Civilini, 1994). Creosote is a complex mixture of PAHs (85%), phenolic compounds (10%), and N-, S- and O-heterocycles (5%). Of the 150±200 compounds present in creosote, only a few are present in concentrations of 1% (Mueller et al., 1989). Civilini (1994) described a composting process using municipal solid wastes and fertilizer, to clean up PAHs present in creosote-contaminated soil. At 45 C, composting was found to remove substantial amounts of the high molecular weight PAHs, after 15 days (Table 4). However, although the author accounts for volatilization, which was found to be less than 10% for all the PAHs with the exception of acenaphthene (54%), this study based the removal of the PAHs on total extractability of the PAHs and did not consider any fraction which is non-extractable. From the data described in Table 4, it can be seen that as the molecular size and weight increased, there was a commensurate decrease in recovery, suggesting that a fraction of the PAHs may have become sequestered within the compost matrix. Similarly, Joyce et al. (1998) investigated the fate of a mixture of three- and four-ring PAHs (¯uorene, anthracene, phenanthrene, pyrene, benz[a]anthracene) under composting conditions with solid municipal waste monitored over a 60-day period (30 days of active composting followed by 30 days of compost curing). The fate of the

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PAHs was also monitored in HgCl2-treated systems to compare the impact of biotic and abiotic processes. The results of this study showed that the loss processes occurred during the active phase of composting (®rst 30 days). Anthracene, phenanthrene and pyrene were removed e€ectively during the composting process by a combination of biotic and abiotic mechanisms, principally dominated by the biotic processes. Fluorene proved to be too volatile and so most of the compound (approx. 75%) was removed abiotically in the gas phase. Additionally, benz[a]anthracene proved to be resistant to biodegradation throughout the composting incubation with between approximately 40±50% being lost abiotically. 4.4. Petroleum hydrocarbons Oil pollution in the environment is now being taken seriously by the oil industries and as such, these companies are always looking for cost-e€ective methods of dealing with this pollution (Milne et al., 1998). Oil sludge is a waste product of the petroleum industry and Milne et al. (1998) considered composting with a variety of bulking agents as a method for dealing with this recalcitrant mixture. Three bulking agents were used in this study, namely chopped barley straw, heat-treated peat moss and Solv-II, a preparation of peat moss enriched with nutrients and oil-degrading microbes. The authors found that both the Solv-II and the peat moss were equally good at activating biodegradative activity in the composting processes. However, over a composting period of 800 h, there was a reduction of approximately 25% in total petroleum hydrocarbons (TPHs) in the composting systems containing the barley and the peat moss. But, in the composting systems containing the Solv-II bulking agent, there was a 55% reduction in TPHs along with high CO2 production, suggesting high microbial respiratory activity. This study suggested that composting, coupled with bioaugmentation, was a successful approach to take in the remediation of oily residues. During the Iraqi invasion of Kuwait, large amounts of oil-contaminated the desert (Al-Daher et al., 1998). A number of remediation strategies were employed to alleviate this environmental problem, including the composting of desert soil in windrows. Al-Daher et al. (1998) looked at the degradation of two components, total extractable matter (TEM) with dichloromethanesoxhlet extraction and total PAHs. Composting systems comprising of various mixtures of dried sewage sludge, mature composts and petroleum-degrading bacteria (only added after 3 months of composting) resulted in approximately 49±59% degradation as measured by TEM after 8 months. Total PAH degradation for lightly contaminated composting piles was reduced by 55% within 8 months, whereas the overall extent of PAH degradation for heavily contaminated piles was approx. 60% over the same composting time period. However, PAHs with ®ve

or more rings were resistant to degradation during this process. The addition of the petroleum-degrading microbes, in this case, did not have a signi®cant a€ect on the degradation of TEM. 4.5. Pesticides Another class of pollutants in the environment is the organochlorine pesticides, including 2,4-D. In the USA alone, a total of 30,000 tons of pesticides are used annually on lawns, of which 2,4-D is the most commonly used (Michel et al., 1995). A study was carried out to look at the fate of [UL-14C]2,4-D when composted with leaf and grass trimmings (Michel et al., 1995). After 50 days of composting, 47% of the 2,4-D had been mineralized, 23% was complexed with high molecular weight humic acids and approximately 20% was found to be associated with the humin (non-extractable) fraction. Additionally, virtually none of the herbicide was lost through volatilization during composting and less than 1% of the 2,4-D was found in the water fraction of the compost after 50 days. The authors noted that the vast majority of the 2,4-D mineralization occurred at around 60 C due to thermophilic microbial activity. The degradation of diazanon, chlorpyrifos, isofenphos and pendimethalin was measured during the composting of garden wastes (Lemmon and Pylypiw, 1992). After 3 weeks of composting at the lab-scale, there were no extractable (hexane) residues of the added pesticides detected. The authors continued this study by looking at the fate of the four pesticides over a longer time period according to application during the summer months (early, mid- and late, or 1, 38 and 68 days, respectively). At all three test periods, degradation of the pesticides was rapid during the ®rst 2 weeks, with three of the four pesticides having almost disappeared, in terms of what was extractable into hexane. However, extractable (hexane) pendimethalin residues were still detectable after 28 days of composting with grass cuttings. All four pesticides were degraded rapidly in early and midsummer applications during composting, with grass cuttings, to values of less than 0.3 ppm. However, in the late summer application, there was continued detection of pesticide residues after 16 weeks of composting with grass cuttings, although the authors did not explain why this pesticide persistence occurred in this phase. 5. Treatment of pollutants and polluted matrices with compost In contrast to composting, compost (the resultant product of composting, with the exception of horticultural potting composts) can be added to polluted soil after its maturation for remediation purposes. Composts have an enormous potential for bioremediation as

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they are capable of sustaining diverse populations of microorganisms, such as bacteria including bacilli, pseudomonads, mesophilic and thermophilic actinomycetes and lignin-degrading fungi, all with the potential to degrade a variety of aromatic pollutants. Composts can act as a soil ameliorant capable of changing pH, moisture content, soil structure and acting as a nutrient source, thereby improving the contaminated soil environment for indigenous or introduced microbial degradative activity. To date, the use of composts has not been widely applied as a method for bioremediation. One major concern is the problem of mixing non-contaminated material, with contaminated soil resulting in a far greater quantity of contaminated material if the attempted bioremediation proves to be unsuccessful. To avoid this problem, fundamental research followed up by pilot scale testing must be carried out. This is essential for the future success of this putatively viable bioremediation technique. As in the previous section, studies reported in the literature have been discussed on a pollutant-class basis. 5.1. Chlorophenols Straw compost produced in the mushroom industry has been investigated as an inoculum in the bioremediation of chlorophenol-contaminated soil. Laine and Jùrgensen (1996) showed that after a 3-month catabolic induction stage the induced mushroom straw compost could mineralize up to 56% of added [UL-14C]PCP to 14 CO2 and that no dechlorinated intermediates were found, where as uninduced compost did not mineralize the chlorophenol. Induction of the compost was achieved by the percolation of a PCP-amended solution (5±10 mg lÿ1) through the compost. The induction of this catabolic activity for PCP was con®rmed by Semple and Fermor (1997) who investigated the degradative potential of mushroom compost taken at di€erent stages of its preparation/usage cycle, namely Phase 1 (immature compost), Phase 2 (mature compost) and end-of-crop composts. These composts were incubated in the absence or presence of PCP and then further incubated with [UL-14C]PCP until mineralization of the chlorinated substrate had stopped. It was found that, for all three composts, the rates and extents of mineralization of [UL-14C]PCP were greater for the composts previously exposed to PCP. Further, Laine and Jùrgenson (1997) produced composts (1 kg) over a period of 6 months from contaminated soil with bark chips; contaminated soil with bark chips and straw compost; contaminated soil with bark chips and remediated soil; and contaminated soil with bark chips, remediated soil and chlorophenol-contaminated wood chips. [UL-14C]PCP was subsequently added to the four composts, resulting in approximately 60% mineralization after 4 weeks of incubation in all cases.

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5.2. Volatile organic compounds Trichloroethylene (TCE) is one of the most frequently occurring subsurface contaminants and its levels are tightly regulated (Sukesan and Watwood, 1997). Both air stripping and vapour extraction are currently used to remove the TCE from the subsurface, resulting in high concentrations of gaseous TCE. It has been suggested that ®nished/mature compost may be used as a bio®ltration mechanism for the remediation of gas contaminated with TCE. For example, Sukesan and Watwood (1997) investigated the fate of gaseous TCE in two di€erent composts, namely leaf compost and wood chip-bark compost, as well as in hydrocarbon-induced and uninduced composts. The authors found that TCE removal was greater in the leaf compost than in the wood chipbark compost with 95 and 15% TCE removed, respectively. They also found that composts, which had been induced using methane or propane, were stimulated equally into removing TCE from the gas phase. Further improvement in TCE removal was achieved with the introduction of granular activated carbon. Sukesan and Watwood (1998) went on to characterize the growth of TCE-utilizing micro¯ora within the composts showing that enrichment with methane or propane resulted in increases in methanotrophic and propanotrophic populations, respectively. 5.3. Aromatic hydrocarbons The addition of ripe or mature compost to soil polluted with PAHs can induce the removal of these hydrocarbons from the soil. In an early study, Martens (1982) looked at the changes in concentration of fourto six-ring PAHs in two types of compost: fresh composts and mature composts, which had been allowed to ripen for 3±12 months in stacks. Martens (1982) found that there were lower concentrations of four- six-ring PAHs in the mature compost over those found in fresh compost. Additionally, when these composts were incubated with 14C-labelled anthracene, benz[a]anthracene, benzo[a]pyrene and dibenz[a,h]anthracene, it was found that there were signi®cantly higher levels of mineralization to 14CO2 in the mature composts. Maximal values for the four 14C-labelled PAHs in fresh and mature composts were 19 and 62% for anthracene, 8 and 58% for benz[a]anthracene, 0.5 and 19% for benzo[a]pyrene and 1.4 and 21% for dibenz[a,h]anthracene, respectively, over a 10-week incubation period. Much later, Mahro and KaÈstner (1993) investigated the fate of pyrene in soil and soil±compost mixtures over a period of 100 days, ®nding that the degradation of pyrene was enhanced signi®cantly with the addition of mature/ripe compost with >80% removed after 20 days in the presence and <5% removed in the absence of the compost. Further, KaÈstner et al. (1995) investi-

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gated the impact of mature compost addition on the fate of 14C-labelled anthracene in soil. To simulate more genuine conditions, the 14C-labelled PAH was dissolved in diesel and mixed into the soil. In soil±compost incubations, 23% of the 14C-labelled anthracene was mineralized to 14CO2 and 42% was irreversibly sequestered/ bound to the soil±compost matrix after 103 days. However, in soil-only incubations, approximately 88% of the PAH was recoverable by solvent extraction with the formation of bound residues being less signi®cant. KaÈstner and Mahro (1996) followed this work up by investigating the degradation of naphthalene, anthracene, ¯uoranthene and pyrene in soil and soil±compost incubations. The authors found that the presence of the compost enhanced the removal of the PAHs and that the presence of the organic matrix of the compost was essential for enhanced degradation. The authors suggested that the presence of the microorganisms capable of degrading natural humic substances were responsible for the co-metabolic degradation of the PAHs as no bacteria capable of mineralizing the PAHs could be detected in the compost. Wischmann and Steinhart (1997) investigated the removal of PAHs and the formation of PAH degradation products in soil±compost mixtures. In unamended soils, only PAHs up to three aromatic rings were degraded over 15 weeks; however, there was enhanced elimination of the parent compounds with the addition of compost with approximately 100% of ¯uoranthene and pyrene, >90% of benz[a]anthracene and chrysene and approximately 70% of benzo[a]pyrene removed from the soil mixed with the compost, after 180 days. BTEX compounds are toxic products of the petroleum industry (Semple et al., 1998). The degradation of benzene was assessed in spent mushroom compost after a 3-month enrichment period in the presence of a variety of BTEX compounds (Semple et al., 1998). It was found that as the incubation temperature was raised from 18 C to 37 C to 50 C, there was a commensurate increase in the mineralization of [UL-14C]benzene over 14 days. 6. Conclusions Composting and the use of composts for the bioremediation of contaminated soil have been highlighted in this review. Although both techniques have been successfully applied to ameliorate soil contaminated with a variety of organic pollutants; a number of processes, either singly or in combination, may prevent target endpoint concentrations being attained. Pollutant bioavailability is an important factor in determining the success or failure of a given bioremediation strategy. The absence of bioavailable pollutants may prevent biodegradation and, hence, bioremediation from

taking place. Bioavailability is primarily related to a compound's intrinsic physico-chemical properties, with aqueous solubility being of key importance (Cerniglia, 1992). For example, `lighter' more water-soluble PAHs can be removed more extensively than `heavier' hydrophobic PAHs (Civilini, 1994). Further, bound residue formation has been shown to be more extensive for hydrophobic pollutants, such as the heavier PAHs (McFarland and Qiu, 1995). The matrix with which pollutants are associated may also in¯uence pollutant bioavailability. The amount and nature of soil organic matter has been proposed by many workers as being one of the most signi®cant factors dominating organic compound interactions within soil (Brusseau et al., 1991; Hatzinger and Alexander, 1995; Cornelissen et al., 1998; Piatt and Brusseau, 1998). The nature of highly contaminated materials, such as spent oxides or tars, also in¯uences pollutant availability (Woolgar and Jones, 1999). Pollutant concentrations also in¯uence bioavailability, such that higher concentrations have been shown to enhance sorption to soil (Divincenzo and Sparks, 1997). Although `bioavailable pollutant' concentrations are of fundamental importance to bioremediation, a more signi®cant factor is the rate at which compounds transfer to the aqueous phase (Bosma and Harms, 1996; Bosma et al., 1997). Biodegradation pathways of organic pollutants may vary in accordance with the chemical structure of the pollutant and the particular degrading microbial species present. In some cases degradation products are more reactive to subsequent transformations and, therefore, may be mineralized (e.g. PCP; Laine and Jùrgensen, 1996), while other degradation products are more susceptible to binding with organic matter (e.g. PAHs; KaÈstner et al., 1995). Additionally, microbial catabolism can also be driven by pollutant concentration, since a minimum threshold pollutant concentration is required for catabolic induction in degrading microorganisms (Boethling and Alexander, 1979). Further, low pollutant concentrations have been shown to prevent end-points being attained in a number of attempted bioremediation strategies (Head, 1998). Conversely, pollutant concentrations that are too high have been shown to result in toxicity, which may inhibit biodegradation (Apajalahti and Salkinoja-Salonen, 1984). Composting and the use of composted materials have both been successfully applied to the bioremediation of PCP-contaminated soil (Valo and Salkinoja-Salonen, 1986; Laine and Jùrgensen, 1997). The principal vector of PCP loss under both bioremediation regimes was mineralization (Valo and Salkinoja-Salonen, 1986; Laine and Jùrgensen, 1997). Whereas Bruns-Nagel et al. (1998) reported 90% losses of TNT from contaminated soil during composting although no mineralization took place. Binding to organic matter (Pennington et al., 1995) and self-polymerisation (Isbister et al., 1984) have been proposed as the principal vectors of TNT removal from contaminated soil during composting. It has been

K.T. Semple et al. / Environmental Pollution 112 (2001) 269±283

suggested that binding/reaction occurs after initial reduction of nitro-functional groups to form aminodinitrotoluenes containing more nucleophilic amino-functional groups. Finally, for some pollutants inherent recalcitrance prevents biodegradation completely; e.g. CDD/F contaminants present in technical grade chlorophenol mixtures have been shown to resist degradation during successful composting of chlorophenol-contaminated soil (Laine and Jùrgensen, 1997). In order for pollutants to become available to microbial attack, desorption may have to take place. In terms of composting, the rate of organic pollutant desorption from soil is enhanced as temperature is increased, socalled thermal desorption (Pignatello and Xing, 1996). It follows that the use of composting strategies for bioremediation, where temperatures can exceed 60 C, should prove advantageous by directly increasing rates of compound mass transfer to the aqueous phase. Another favourable factor inherent to compost-based remediation is that large amounts of organic matter are added to the system. Organic matter is considered a major factor in the `locking-up' of organic pollutants (Brusseau et al., 1991; Cornelissen et al., 1998). Thus, while available pollutants are transferred more readily for biodegradation (by thermal desorption) an additional fraction is made recalcitrant by strongly associating with residual organic matter. This may be particularly important when considering recalcitrant pollutants such as high molecular weight PAHs or PCDD/Fs. To support both of these processes the rich micro¯ora present in the composting/ compost materials, on the one hand promote biodegradation, while on the other favour pollutant association with organic matter through its turnover/humi®cation (Carmichael and Pfaender, 1997; Guthrie and Pfaender, 1998). The end result is a soil±compost mixture that contains reduced pollutant concentrations and/or a matrix that has less pollutants in a bioavailable form. As a result of these combined processes, a reduction in overall risk is achieved. It should be borne in mind that while composting/compost processes `lock up' pollutants, the long-term stability of such `stabilized' matrices is uncertain. Although the complete destruction of the pollutant molecule, through mineralization, may be desirable, it may not be achievable for every pollution scenario. Therefore, biotransformation to less toxic or even innocuous compounds may be a more likely scenario, which can be used. Further, pollutants which prove to be completely intractable, such as PCDD/Fs, may be locked up in the organic matrix of the compost±soil mixture, thereby reducing interaction with biota and minimizing any potential toxicity issues. In spite of these limitations, composting and the use of composted materials for the amelioration of contaminated soil have proven advantages. Not only have these methods reduced soil-associated pollutant concentrations, but they

281

have also improved soil quality through the addition of organic matter. Thus, in contrast to land®ll and destructive treatment methods, such as incineration, the use of composted materials and composting for amelioration promotes soil sustainability and re-use. While the numerous scienti®c investigations, described in this review, have shown that composting of pollutant-contaminated soils or the addition of mature/ripe composts can promote or enhance the degradation of soil-associated organic pollutants and thereby reduce the toxicity of contaminated soils, a note of caution should be sounded. This is particularly appropriate where comparisons are drawn between genuinely polluted soils and `spiked' soils. The chemical behaviour of the xenobiotic compound in question may be considerably di€erent in arti®cially contaminated soils, giving a false result, leading to over-optimistic assumptions. It is, therefore, paramount that biological and chemical processes inherent to composting/compost strategies are assessed on a `site by site' basis at laboratory and pilot scales prior to undertaking ®eld scale ventures. References Adenuga, A.O., Johnson Jr., J.H., Cannon, J.N., Wan, L., 1992. Bioremediation of PAH-contaminated soil via in-vessel composting. Water Science and Technology 26, 2331±2334. Al-Daher, R., Al-Awadhi, N., El-Nawawy, A., 1998. Bioremediation of damaged desert environment using windrow soil pile system in Kuwait. Environment International 24, 175±180. Alexander, M., 1995. How toxic are toxic chemicals in soil? Environmental Science and Technology 29, 2713±2717. Apajalahti, J.A., Salkinoja-Salonen, M.S., 1984. Absorption pf pentachlorophenol (PCP) by bark chips and its role in microbial PCP degradation. Microbial Ecology 10, 359±367. Apajalahti, J.A., Salkinoja-Salonen, M.S., 1986. Degradation of Polychlorinated Phenols by Rhodococcus chlorophenolicus. Applied Microbiology and Biotechnology 25, 62±67. Ball, W.P., Roberts, P.V., 1991a. Long-term sorption of halogenated organic chemicals by aquifer material 1. Equilibrium. Environmental Science and Technology 20, 1223±1235. Ball, W.P., Roberts, P.V., 1991b. Long-term sorption of halogenated organic chemicals by aquifer material 2. Interparticle di€usion. Environmental Science and Technology 20, 1237±1249. Beck, A.J., Wilson, S.C., Alcock, R.E., Jones, K.C., 1995. Kinetic restraints on the loss of organic-chemicals from contaminated soils Ð implications for soil-quality limits. Critical Reviews in Environmental Science and Technology 25, 1±43. Betts, W.D. (Ed.), 1991. Biodegradation: Natural and Synthetic Materials. Springer-Verlag, Germany. Boethling, R.S., Alexander, M., 1979. E€ect of concentration of organic chemicals on their biodegradation by natural microbial communities. Applied and Environmental Microbiology 37, 1211±1216. Bosma, T.N.P., Harms, H., 1996. Bioavailability of organic pollutants. EAWAG News 40E, 28±31. Bosma, T.N.P., Middeldrop, P.J.M., Schraa, G., Zehnder, A.J.B., 1997. Mass transfer limitations of biotransformation: quantifying bioavailability. Environmental Science and Technology 31, 248±252. Breitung, J., Bruns-Nagel, D., Steinbach, K., Kaminski, L., Gemsa, D., von LoÈw, E., 1996. Bioremediation of 2,4,6-trinitrotoluene-contaminated soils by two di€erent aerated compost systems. Applied Microbiology and Biotechnology 44, 795±800.

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