Bioremediation of Polycyclic AromaticHydrocarbon (PAH)-Contaminated Waste Using
Composting Approaches
B. ANTIZAR-LADISLAO, J. M. LOPEZ-REAL, and A. J. BECKImperial College London, Wye Campus, Wye, Ashford, Kent, United Kingdom
Environmental pollution and in particular contaminated wastedue to polycyclic aromatic hydrocarbons (PAHs) are of currentenvironmental concern. In an effort to find solutions, bioremedi-ation techniques have shown promising results in the treatment ofcontaminated wastes. Composting approaches as a bioremediationtechnology to treat contaminated waste were first reported in the1980s. This article provides a comprehensive review of researchto date on the use of composting approaches for bioremediationof PAH-contaminated waste. It critically evaluates the existing re-search in an effort to assess the relative effectiveness of differentcomposting approaches, determine optimal composting operationconditions, and identify the limitations and advantages of usingcomposting approaches relative to other solutions, and recommendsareas of further research effort.
Polycyclic aromatic hydrocarbons (PAHs) are a class of organic chemicalsconsisting of two or more benzene rings fused in a linear, angular, or clusterarrangement. Their occurrence in the environment is partly the result ofnatural processes including forest fires and volcanic eruptions, and partlydue to anthropogenic activities including the incomplete combustion of fossilfuels, accidental discharge during transport, use and disposal of petroleum
Address correspondence to Blanca Antizar-Ladislao, Imperial College London, WyeCampus, Wye, Ashford, Kent, TN25 5AH, United Kingdom. E-mail: [email protected]
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250 B. Antizar-Ladislao et al.
products, and incineration of refuse and wastes (Dyke et al., 2003; Fernandezet al., 2002; Beck et al., 1996). Thus, in areas of high population density andindustrial activity, PAH releases to soils and the wider environment haveled to higher concentrations of these contaminants than would be expectedfrom natural processes alone. The problem is exacerbated because PAHshave accumulated in soils, sediments, and animals after their release to theenvironment because of their hydrophobicity. This gives cause for concernbecause they have been implicated in many adverse effects on wildlife andhuman health—most notably carcinogenesis (Brandt et al., 2002; Brenneret al., 2002; Yu, 2002; VanBrummelen et al., 1996). As a result, there is a needto identify and cleanup sites that have become heavily contaminated so thatthey do not pose unnecessary risks to health. A further incentive for cleanupis the presence of many highly contaminated sites on the premises of formermanufactured gas plants (MGP), which were typically found on land close tothe center of towns and cities, land that is often now a valuable capital assetif it can be cleaned up to a salable condition (Haeseler et al., 1999). Thus, thepotential of using physical, chemical, or biological technologies (or hybridcombinations of these) to remediate PAH-contaminated sites has receivedmuch attention in recent years (Maini et al., 2000; Zeng et al., 2000; Thomasand Lester, 1993). Since the 1970s, research on the biological degradationof PAHs has demonstrated that bacteria, fungi, and algae possess catabolicabilities that may be used for the bioremediation of PAH-contaminated wasteand water (Eriksson et al., 2003; Dean-Ross et al., 2002; Bhatt et al., 2002; Leiet al., 2002). Bioremediation technologies such as phytoremediation (Jonerand Leyval, 2003; Mougin, 2002), land farming (Picado et al., 2001), andcomposting (Sasek et al., 2003b; Canet et al., 2001; Potter et al., 1999) havebeen used for biodegradation of PAH-contaminated wastes.
Composting of soils contaminated with hazardous materials is still anemerging ex situ biotreatment technology. However, interest is increasing andit has been shown to be effective in biodegrading PAHs (Sasek et al., 2003b;Canet et al., 2001; Potter et al., 1999), chlorophenols (Laine and Jørgensen,1997), polychlorinated biphenyls (Michel et al., 2001), explosives (Gundersonet al., 1997), and petroleum hydrocarbons (Jørgensen et al., 2000; Namkoonget al., 2002) at the laboratory and/or field scales. There are several reportson soil composting–bioremediation using a variety of composting systemssuch as open-air systems, which include mechanically turned windrow andstatic aerated piles, and enclosed systems, which include modular containersand tunnels or buildings (Potter et al., 1999; Jørgensen et al., 2000; Saseket al., 2003b). The use of composted materials to ameliorate contaminatedwaste has recently been reviewed by Semple et al. (2001), Buyuksonmezet al. (1999) have reviewed the occurrence, degradation and fate of pesticidesduring composting, and Ro et al. (1998) have reviewed the use of compostingin the bioremediation of explosives-contaminated waste. However, as yet
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Bioremediation of Polycyclic Aromatic Hydrocarbon 251
there is no single source that reviews the success of composting in treatmentof PAH-contaminated wastes—hence, the purpose of this article.
Our primary objective is to provide a comprehensive review of researchto date on the use of composting approaches for bioremediation of PAH-contaminated waste. In doing so we critically evaluate the existing researchin an effort to determine optimal composting operation conditions, assessthe relative effectiveness of different composting approaches, identify thelimitations and advantages of using composting approaches relative to othersolutions, and recommend areas of further research effort.
II. BIOREMEDIATION OF POLYCYCLIC AROMATICHYDROCARBONS
A. Regulatory Considerations
Regulatory objectives and priorities relating to the clean up of PAH-contaminated soils vary by country. In the United Kingdom there is an ac-tion level of 500 mg PAH kg−1 air-dried soil for land used for recreationand 10,000 mg PAH kg−1 air-dried soil for land with a hard covering. TheUnited Kingdom takes the view that land should be cleaned to make it fit forits intended use. By contrast, the United States and the Netherlands adopta multifunctional approach whereby land must be fit for any use. Conse-quently, soil quality limits are much lower than in the United Kingdom; forexample, the Netherlands has an action level of 40 mg PAH kg−1 air-driedsoil (Wilson and Jones, 1992). Table 1 shows trigger concentrations of PAHsgiven by the Interdepartmental Committee on the Redevelopment of Con-taminated Land (ICRCL, 1987) in the United Kingdom, and Table 2 showsthe optimum and action concentrations for soil, sediments, and groundwa-ter adopted by the Dutch List. In addition, the United States, among othercountries, has international obligations under the United Nations (UN) ECEPersistent Organic Pollutants (POPs) Protocol (UN ECE POPs, 2003) to reportreleases of four PAHs. As compared with the existing procedure in the UnitedStates, Europe has more uniform criteria, established as the Dutch List.
TABLE 1. Interdepartmental Committee on the Redevelopment of Contaminated Land(ICRCL) 59/83 Trigger Concentrations
Trigger values(mg kg−1 air-dried soil)
Contaminant Planned use Threshold Action
Polyaromatic Domestic gardens, allotments, play areas 5 500hydrocarbonsa,b Landscaped areas, buildings, hard cover 1000 10,000
aUsed as a marker for coal tar. See CIRIA (1988, Annex 1).bSee CIRIA (1988) for details of analytical methods.
Note. From the Ministry of Housing, Spatial Planning and Environment, Directorate-Generalfor Environmental Protection, Department of Soil Protection (625), the Hague, the Nether-lands.aPAH (total of 10) is the total of anthracene, benzo[a]anthracene, benzo[a]fluoranthrene,benzo[g,h,i]perylene, benzo[k]fluoranthrene, chrysene, fluoranthrene, indeno[1,2,3-cd]pyrene, naphthalene, phenanthrene.bIf contamination is due to only one compound, the value used is the intervention valueof that compound. Where there are two or more compounds the value for the total ofthese compounds applies.
Regulatory objectives and priorities relating to composting of biodegrad-able waste also vary by country. In the United Kingdom, composting is clas-sified as a waste recovery operation under the Waste Framework Directive.In addition, composting of waste is a vital component of meeting the WasteStrategy 2000 (DETR, 2000) targets for recycling and composting set at 25%by 2005, 30% by 2010, and 33% by 2015. In Europe, the European Com-munity (EC) Landfill Directive 1999 (EC, 1999) sets a target for reductionof biodegradable waste to landfill of 25% by 2010, 50% by 2013, and 65%by 2020. In recent years an increasing tendency to compost green (garden)waste has become apparent (Slater and Frederickson, 2001). In relation tothe longer term requirements of the EC Landfill Directive, composting ofkitchen waste, green waste, and in general biodegradable organic waste,which might include contaminated soil, will probably have an importantrole to play (Burnley, 2001; Phillips et al., 2001). Recent European Union(EU) regulation (EU Animal By-Products Regulation, 2003) requires that thecomposting of catering waste containing meat must take place in a “closedcomposting reactor” operated at least 70◦C for 1 h. This means that cateringwaste containing meat cannot be treated in an open windrow, except as asecond stage after it has first been treated in a closed reactor.
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Bioremediation of Polycyclic Aromatic Hydrocarbon 253
FIGURE 1. Regulatory considerations for bioremediation of PAH-contaminated soil usingcomposting approaches.
Figure 1 summarizes the regulations to consider at the different stagesof the application of composting-bioremediation to PAH-contaminated waste.“A” indicates regulations setting up an action concentration of PAHs presentin soil, that is, the New Dutch List. “B” indicates regulations dealing withorganic wastes to compost, that is, the EC Landfill Directive. “C” indicatesregulations to the specific operational parameters required for compostingof catering waste or waste where animal by-products are present, that is, theEU Animal By-Products Regulation.
B. PAHs in the Environment
PAHs are present in contaminated soils and sediments at varying concen-trations depending on the nature of the industrial sites and sources of con-tamination to the environment (Mueller et al., 1996). Typical concentrationsof PAHs in contaminated soils from creosote production, wood preserving,MGP, and petrochemical sites are given in Table 3. Total concentration ofPAHs at wood preserving sites (WPs) can be high when compared to the
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254 B. Antizar-Ladislao et al.
TABLE 3. Examples of PAH Concentrations in Contaminated Soils from Industrial Sites
PAH concentration (mg · kg−1 soil) in PAH-contaminatedsoils/sediment
Note. CP, creosote production site (Ellis et al., 1991); WP, wood preserving site (Ahtiainen et al., 2002);SFMn, Superfund site Minnesota, (U.S. EPA, 1995a); GW, gas works site (Bewley et al., 1989); PC, petro-chemical site (Juhasz and Naidu, 2000); MGP, manufacturing gas plant site (Birnstingl, 1997); SFSMs,Superfund site, Mississippi (U.S. EPA, 1995b); COGEMA, French MGP site (Bogan et al., 1999).
concentration of PAHs at MGP sites. In addition, MGP are well known tohave concentrated and diffuse areas of contamination (Birnstingl, 1997) .Forexample, in the sites mentioned in Table 3, �PAHWPs is 67 times higher than�PAHMGPs. The nature of different sites will also influence the concentra-tion of particular PAHs associated with contaminated soils. For example, ina soil from a WP site, Ahtiainen et al. (2002) reported a total PAH concen-tration of 23,600 mg kg−1, and benzo[a]pyrene comprised 0.3% of the totalPAH content of the soil, but indeno[1,2,3-cd]pyrene was not detected. At anMGP site, benzo[a]pyrene and indeno[1,2,3-cd]pyrene comprised 1.3% and2.8%, respectively, of the total PAH content of the soil, and fluoranthene anddibenzo[a,h]anthracene comprised 30% and 33%, respectively (Bewley et al.,1989). These differences in PAH concentrations in soils from different sourcesshould be considered when optimization of composting conditions need tobe achieved for maximum PAH removal.
C. Physicochemical Properties of PAHs
Biodegradation of different PAHs, regardless of the bioremediation technol-ogy applied, occurs to different extents and at variable rates depending ontheir physicochemical properties, waste characteristics, and environmental
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Bioremediation of Polycyclic Aromatic Hydrocarbon 255
conditions. Table 4 outlines some physicochemical properties of the 16 U.S.Environmental Protection Agency (EPA) regulated PAHs. Among the mostimportant physicochemical properties to be considered during biodegrada-tion processes appears to be water solubility. PAH water solubility decreases(and hydrophobicity increases) with an increase in number of fused ben-zene rings, and with angularity (Harvey, 1997). Thus, high-molecular-weightPAHs are more slowly desorbed and dissolved than low-molecular-weightPAHs and therefore are less available for microbial degradation (Cerniglia,1992). In addition, volatility is an important physicochemical property, espe-cially during elevated temperatures that occur in composting (>50◦C), whichfacilitates volatilization of PAHs. Volatilization generally decreases with an in-creasing number of fused rings. The number of aromatic rings, structure ofthe PAHs, and resonance energy might also affect the degradation rate ofPAHs (Kastner and Mahro, 1996).
D. Biodegradability of PAHs
The recalcitrance of PAHs to microbial degradation generally increases withtheir molecular weight and their octanol–water partitioning coefficient (logKow). In addition, many high-molecular-weight PAHs are only degraded withdifficulty or not at all, due to their low water solubility, high resonance en-ergy, and toxicity (Cerniglia, 1992). The microbial degradation of PAHs hasbeen reviewed (Juhasz and Naidu, 2000; Cerniglia, 1992). There are numer-ous genera of microorganisms, some of them present during compostingprocesses, that have been observed to degrade PAHs (in der Wiesche et al.,2003; Bhatt et al., 2002; Dean-Ross et al., 2002; Canet et al., 1999), but mech-anisms of metabolism vary (Cerniglia, 1992, 1997) (Table 5). The differentpathways used by bacteria, mammals, and fungi in the metabolism of PAHsare summarized in Figure 2.
Bacteria initially oxidize PAHs by incorporating both atoms of molec-ular oxygen into the PAH nucleus to form cis-dihydrodiols (dioxygenase).The initial ring oxidation is usually the rate-limiting step. cis-Dihydrodiolsare then rearomatized through a cis-dihydrodiol dehydrogenase to yield adihydroxylated derivative. Further oxidation of the cis-dihydrodiols leads tothe formation of catechols, which are substrates for other dioxygenases thatbring about enzymatic cleavage of the aromatic ring. Catechol can be oxi-dized via two pathways. The ortho pathway involves cleavage of the bondbetween carbon atoms of the two hydroxyl groups. The meta pathway in-volves cleavage of the bond between a carbon atom with a hydroxyl groupand the adjacent carbon atom with a hydroxyl group. Ring cleavage resultsin the production of sucinic, fumaric, pyruvic, and acetic acids and alde-hydes, all of which are utilized by the microorganisms for the synthesis ofcellular constituents and energy. A by-product of these reactions is the pro-duction of carbon dioxide and water. PAHs such as phenanthrene, pyrene,
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TAB
LE4
.St
ruct
ure
and
Phys
ico-C
hem
ical
Pro
per
ties
of16
U.S
.EPA
Reg
ula
ted
PAH
s
Mole
cula
rm
.p.
b.p
.lo
gSo
lubili
tyVap
or
pre
ssure
Hen
ryco
nst
ant
Res
onan
cePA
Hw
eigh
tFo
rmula
Stru
cture
(◦ C)
(◦ C)
Kow
(mm
olL−1
)(P
aat
25◦ C
)(a
tmm
3m
ol−1
)en
ergy
(eV
)
Nap
hth
alen
e12
8C
10H
881
218
3.00
–4.0
02.
4×
10−1
10.9
4.5
×10
−31.
325
Ace
nap
hth
ylen
e15
2C
12H
895
270
3.70
—5.
96×
10−1
—1.
325
Ace
nap
hth
ene
154
C12
H10
96.2
279
3.92
–5.0
72.
9×
10−2
5.96
×10
−12.
4×
10−4
—
Fluore
ne
166
C13
H10
115–
116
294
4.18
1.2
×10
−28.
86×
10−2
7.4
×10
−5—
Anth
race
ne
178
C14
H10
264
547
4.46
–4.7
60.
072.
0×
10−4
——
Phen
anth
rene
178
C14
H10
100.
533
84.
457.
2×
10−3
1.8
×10
−22.
7×
10−4
1.92
4
Fluora
nth
ene
202
C16
H10
108.
838
34.
901.
3×
10−3
2.54
×10
−11.
9×
10−3
2.18
4
Pyr
ene
202
C16
H10
150.
439
34.
907.
2×
10−4
8.86
×10
−41.
3×
10−5
2.09
9
Ben
zo[a
]anth
race
ne
228
C18
H20
160.
742
55.
61–5
.70
—7.
3×
10−6
1.2
×10
−62.
313
256
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Chry
sene
228
C18
H20
253.
843
15.
611.
3×
10−5
5.7
×10
−76.
7×
10−7
2.47
7
Ben
zo[b
]fluora
nth
ene
252
C20
H12
——
6.57
——
——
Ben
zo[k
]fluora
nth
ene
252
C20
H12
——
6.84
——
——
Ben
zo[a
]pyr
ene
252
C20
H12
178.
149
66.
041.
5×
10−5
8.4
×10
−72.
7×
10−7
2.57
9
Dib
enzo
[a,h
]Anth
race
ne
278
C22
H14
266.
653
55.
80–6
.50
1.8
×10
−63.
7×
10−1
02
×10
−92.
958
Inden
o[1
,2,3
-cd]p
yren
e27
6C
22H
1216
353
67.
66—
——
—
Ben
zo[g
,h,i]
per
ylen
e27
6C
22H
1227
8.3
542
7.23
2×
10−5
6×
10−8
2×
10−7
3.09
8
Not
e.Fr
om
ISPA
C(2
003)
.
257
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258 B. Antizar-Ladislao et al.
TABLE 5. Microorganisms Degrading PAHs
PAH Organism
Naphthalene Bacteria: Acinetobacter calcoaceticus, Alcaligenes denitricans,Brevundimonas vesicularis, Burkholderia cepacia, Comamonastestosteroni, Corynebacterium renale, Cycloclasticus sp., Neptunomonasnaphthovorans, Moraxella sp., Mycobacterium sp., Pseudomonas sp.,P. Fuorescens, P. marginalis, P. putida, P. saccharophila, P. stutzeri,Sphingomonas paucimobilis, Streptomyces sp., Rhodococcus sp.
Fungi: Trametes versicolor, P. janthinellum.Indeno[1,2,3-c,d]pyreneBenzo[g,h,i]perylene
Note. Adapted from Cerniglia (1992, 1997) and Kanaly and Harayama (2000).
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260 B. Antizar-Ladislao et al.
FIGURE 2. Different pathways used by bacteria, mammals, and fungi in the metabolism ofPAHs.
and benzo[a]pyrene have complex fused ring structures so bacteria metabo-lize them at multiple sites to form isomeric cis-dihydrodiols (van Herwijnenet al., 2003; Juhasz and Naidu, 2000; Kanaly and Haramaya, 2000).
Mammals incorporate one atom of molecular oxygen into the PAH toform arene oxides that can either undergo enzymatic hydration by epoxidehydrolase to form trans-dihydrodiols or else rearrange nonenzymatically toform phenols (Lei et al., 2002). Exposure to PAH-contaminated waste in-creases the risk of cancer in mammals. The carcenogenicity of PAHs is asso-ciated with the complexity of the molecule, and with metabolic activation toreactive diol epoxide intermediates and their subsequent covalent binding tocritical targets in DNA (Bostrom et al., 2002).
A diverse group of ligninolytic and nonligninolytic fungi have the abilityto oxidize PAHs. In contrast to bacteria, nonligninolytic fungi and prokaryoticalgae (cyanobacteria) metabolize PAHs in pathways that are generally similarto those used by mammalian enzyme systems (Cerniglia, 1992). Enzymesfrom both fungal and mammalian systems oxidize PAHs to arene oxides bythe cytochrome P-450 enzyme system. Then the oxides can isomerize toyield phenols or undergo enzymatic hydration to yield trans-dihydrodiols.Lignin degradation is carried out by mechanisms related to the productionof highly reactive intermediates by enzymes, such as lignin peroxidase andmanganese-dependent peroxidase. Most degradative mechanisms reported
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Bioremediation of Polycyclic Aromatic Hydrocarbon 261
are cometabolic, where an alternate carbon source is utilized for energyand growth, while the PAH is transformed as a consequence of this growth(Cerniglia, 1997; Mueller et al., 1996).
While there are many organisms capable of degrading the low-molecular-weight (LMW: two- and three-ring) PAHs, relatively few generahave been observed to degrade high molecular weight (HMW: four-,five-, and six-ring) PAHs. There are still very few reports of bacteria out-side the nocardioform actinomycetes group capable of growing on high-molecular-weight PAHs (Kastner et al., 1994); however, a variety of non-actinomycete bacteria (Pseudomonas putida, P. aeruginosa, Flavobacteriumsp., Burkholderia cepacia) have also been reported to metabolize fluoran-thene, pyrene, chrysene, and benzo[a]anthracene. Currently, there is only lim-ited information regarding the bacterial biodegradation of five- and six-ringPAHs in environmental samples and pure or mixed cultures. Benzo[a]pyrenebiodegradation by pure and mixed cultures of bacteria has been shown tooccur, although bacteria capable of utilizing benzo[a]pyrene as a sole sourceof carbon and energy have not been demonstrated yet.
Numerous genera of fungi with the ability to oxidize naphthalene havebeen identified. Most degradative mechanisms reported are cometabolic. Thewhite-rot fungus Phanerochaete chrysosporium has been reported to mineral-ize phenanthrene, fluorine, fluoranthene, anthracene, and pyrene in nutrient-limited cultures (Steffen et al., 2003; Bumpus, 1989). Other white-rot fungi,such as Trametes versicolor, Bjerkandera sp., and Pleurotus ostreatus, havebeen shown to metabolize PAHs such as phenanthrene, anthracene, pyrene,benzo[a]pyrene, and fluorine (Cerniglia, 1997). The basidiomycete Pleurotusostreatus can degrade a wide range of PAHs, including anthracene, phenan-threne, pyrene, fluorene, and benzo[a]pyrene (Bezalel et al., 1997).
III. COMPOSTING APPROACHES
Composting is an aerobic process where organic materials are biologically de-composed. Heat produced during this process leads to elevations in temper-ature characteristic of composting. Composting processes typically comprisefour major microbiological stages in relation to temperature: mesophilic, ther-mophilic, cooling, and maturation, during which the structure of the microbialcommunity also changes (Bolta et al., 2003; Liang et al., 2003; VanderGheynstand Lei, 2003).
Composting systems are generally divided into three categories:windrow, static pile, and in-vessel. In the windrow approach, the solid wastemixtures are composted in long rows and aerated by convective air move-ment and diffusion. The mixtures are mechanically turned periodically toexpose the organic matter to ambient oxygen. This approach to compostingleads to a characteristic fluctuating temperature regime, as core temperatures
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262 B. Antizar-Ladislao et al.
FIGURE 3. Variation of “compost” core temperature with the windrow system.
will fall on turning due to introduction of cooler masses of air and releaseof entrapped heated air, followed by a subsequent rise in temperature anda new temperature peak as maximum microbial activity resumes. Over timethe peak temperatures fall as the level of substrates declines. Recording oftemperature fluctuations results in the characteristic jagged tooth pattern ofwindrowing (Figure 3). Windrowing is a simple mechanical system with noeffective control of either oxygen or temperature levels within the organicmass undergoing decomposition. The technique, though popular commer-cially due to its ease of implementation and relative low cost of installa-tion, has been criticized microbiologically since considerable periods of timeelapse when either oxygen or temperature or both will become limiting onmicrobial diversity and decomposition.
In the static pile approach, piles of solid waste mixture, often with bulk-ing agents (wood chips, straw) as a matrix improver, are aerated using aforced aeration system, which is installed under the piles, to maintain a mini-mum oxygen level throughout the compost mass. The forced aeration systemmay be in either the positive or negative modes, which leads to maximumcore temperatures slightly above or below the central point, respectively.Aeration may be on a timed basis (Epstein, 1997), leading to high coretemperatures and a severely restricted microbial diversity or by tempera-ture feedback control (Finstein et al., 1986). This latter approach is based ona greater understanding of the microbial ecology involved; for this approachthe control of air flow is dictated by reference to an upper temperature limitabove which the fans blow continuously until heat loss through latent heatof vaporization of water (seen as steaming) reduces the core temperature toan acceptable operational level. In both the Epstein (1997) and the Finsteinet al. (1986) static systems there are substantial variations in temperaturethroughout the composting mass, and the cooler edges are routinely coveredwith a thick layer of finished compost acting as a thermal blanket to ensurethat these edges reach regulatory temperature levels. Although the Finsteinet al. (1986) system allows for the control of maximum core temperature,
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and in theory supplies excess of oxygen for decomposition, both static ap-proaches suffer from this gradation of temperature (and oxygen) throughoutthe organic waste mass. In static systems active composting will occur over a3- to 4-week period depending on the nature of the substrate being pro-cessed. This active phase is often followed by a 2- to 3-month period ofmaturation.
In-vessel composting takes place in a partially or completely enclosedcontainer in which environmental conditions can be controlled. Enclosedvessels more closely approximate a laboratory incubator, where the organicmass and its associated microflora should be exposed to a more even temper-ature profile. Control of temperature in in-vessel systems is usually achievedthrough recycling of exhaust gases with intermittent mixing of fresh air tomaintain an agreed temperature (Antizar-Ladislao et al., 2003; Seymour et al.,2001; Fraser and Lau, 2000). In theory, such systems should allow for excel-lent control of temperature within the vessel and considerably less variationof temperature through the composting mass. However, in-vessel systemshave serious limitations in general composting due to limited throughputsand high installation costs (Das and Keener, 1997). Because of this, theyare often used as a form of pretreatment “bioreactor” for up to 5 days priorto conventional composting (usually windrowing), where further decompo-sition, stabilization, and degassing take place (Sasek et al., 2003b; Epstein,1997; Haug, 1993).
The capability of microorganisms to biodegrade specific contaminantsmay not differ significantly from the ambient soil environment to that of com-posting, but the transformation potential differs for several reasons (Williamsand Keehan, 1993). First, elevated temperatures of composting (>50◦C) canincrease the enzyme kinetics involved in biodegradation processes. Second,co-oxidation may be enhanced due to the range of alternative substratespresent. Third, modifications in the physical and chemical microenviron-ments within the composting mass can serve to increase the diversity ofthe microflora to which the contaminant is exposed. Last, high temperatureswill typically increase the solubility and mass transfer rates of the contami-nant, thereby making them more available to metabolism. However, some ofthese positive attributes just listed may be in conflict with the overall impactof temperature and microbial activity.
Over the past two decades there has been much discussion concerningthe appropriate levels of temperature for maximizing decomposition rates incomposting (Walter et al., 1992; Finstein et al., 1989). Reviews of the liter-ature during this period established unequivocally that lower temperaturesfavor more efficient composting (Bardos et al., 1989). In practice, however,composting processes are often subject to regulatory levels of temperature(e.g., 70◦C during 1 h; 55◦C during 72 h) in order to meet UK national levelsof pathogen reduction (EC, 2003; DETR, 2000). Emphasis is therefore placedon pathogen reduction before composting process efficiency. The impact of
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this approach to composting is important when considering organic wastesthat are likely to be nonpathogen- or low-pathogen-containing materials andwhose contaminant substrates are often complex, requiring multiple enzymesystems for degradation. Research into bioremediation of PAH contaminatedwastes has clearly shown that the most effective degraders of such substratesbelong in the fungal class (Sasek, 2003; Sasek et al., 2003a; Bhatt et al.,2002; Eggen and Sasek, 2002; Canet et al., 1999, 2001). More specifically, themajority of these belong to the Basidiomycota, which contain the fungal gen-era responsible for wood decomposition and hence possess the necessarycomplex array of ligninolytic enzymes and nonenzymatic mechanisms forthe degradation of lignin whose basic chemical structures are similar to thePAHs. Lignin degraders have been shown to be amongst the most potent de-graders of PAHs epitomized by Phanaerochaete chrysosopoium (Canet et al.,1999).
The relationship of these organisms to temperature in the context ofcomposting as a bioremediation technology therefore becomes critical. Mi-croorganisms exhibit a wide range of temperature adaptation and evolu-tion. Microorganisms are found in the coldest and hottest regions of Earth(Ferguson et al., 2003; Cavicchioli, 2002; Cowan et al., 2002). At the extremeends of this range are a very few highly specialized genera whose wall andcell membrane modifications allow them to survive and multiply in suchconditions. Figure 4 is a hypothetical diagram relating microbial diversity totemperature. At specific points on this temperature range not only speciesor genera but whole phyla will disappear as their maximum temperaturelimits are reached. As has been pointed out in this review, current compost-ing approaches and technologies tend to emphasize the higher end of therange (>70◦C) in order to meet regulatory requirements for pathogen control.Such temperatures, which will be easily reached in uncontrolled operations,such as windrowing (Epstein, 1997), severely inhibit the microbial diversityand hence enzymatic potential of the system. Those organisms of particularinterest in PAH degradation will be eliminated above approximately 45◦C.
FIGURE 4. Microbial diversity and temperature.
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Some speculation has been given to the temperature range at which mi-crobial diversity is maximized. What little research has been done suggeststhat the range between 40 and 45◦C allows for such maximum diversity tobe expressed for both, bacteria, actinomycetes, and fungi (Antizar-Ladislaoet al., 2003; Walter et al., 1992). It is of interest to note that this range repre-sents an overlap point at which both mesophilic and thermophilic microor-ganisms will be viable and active. A strategy of temperature control at thisrange would seem therefore to be appropriate for the bioremediation ofPAH-contaminated waste.
A. Composting Bioremediation of PAHs-Contaminated Wastes
Composting bioremediation approaches consist of the addition of compostingredients to a contaminated waste, where the compost matures in the pres-ence of the contaminated waste. Composting is a relatively new bioremedia-tion approach, so few investigations have been conducted and most of themhave been performed at the laboratory scale. In this review, investigationson composting bioremediation of PAH-contaminated waste are presented ina chronological approach, with emphasis on the progressive developmentof this technology on the treatment of PAHs either as single pollutants or amixture of pollutants.
Crawford et al. (1993) were one of the earliest research groups to reporton a composting bioremediation approach to remove PAHs from a contami-nated soil. They reported a controlled composting study in the United Statesat the beginning of the 1980s, although they do not give details on the tech-nology, scale, or conditions used. In their study, naphthalene, pyrene, andbenzo[a]anthracene, at initial concentration of 500 mg kg−1, showed varyingdegrees of degradation. Naphthalene was completely removed during thefirst 7 days of composting; benzo[a]anthracene showed 25% removal duringthe first 7 days and 42% total removal over 30 days of composting; pyreneshowed no decrease during the first 7 days of composting and a moder-ate removal (not percentage removal available) during the last 23 days ofexperiment.
Racke and Frink (1989) studied the fate of phenanthrene during sewagesludge composting at laboratory scale, using radiolabeled phenanthrene(1.3–1.6 mg kg−1 dry weight), which allowed them to study the fate ofphenanthrene. After 18–20 days of composting, between 10 and 11% ofthe phenanthrene was degraded and between 15 and 17% of unextractablephenanthrene metabolites remained in the compost. Unextractable phenan-threne residues formed were either bound to organic matter or incorporatedinto microorganisms.
Adenuga et al. (1992) investigated the biodegradation of pyrene(13 mg kg−1) using in-vessel composting technology at the laboratory scale.
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They amended a spiked soil with composted sewage sludge (40 g mixture),in quantities ranging from 0 to 50% dry weight basis, and the moisture wasadjusted to 40%. The mixture was placed in amber flasks connected to anairflow system, in a water bath. Then the temperature was raised from 20◦Cby 5◦C day−1 to 60◦C over a period of 21 days, probably trying to stimu-late the mesophilic and thermophilic microbiological stages characteristic ofcomposting. The results from their preliminary studies showed that pyrenecould be degraded in the composting of soil/sludge mixture, although therate and extent were not fully described in their study.
Crawford et al. (1993) reported a pilot study to asses the feasibility oftreating two- to four-ring PAH-contaminated soils by composting leaves. Thestudy used small windrows (about 19 m3 each) of varied ratios of soils froma former industrial site contaminated with low levels of two- to four-ringPAHs (about 100 mg kg−1) and other semivolatile compounds (less than10 mg kg−1). During their study, it was observed that temperature, moisturecontents, and ratio of carbon to nitrogen were deficient for optimal compost-ing operation, which indicated the need for larger windrows and increasedprocess control. Complete removal of these PAHs due to mesophilic degra-dation, abiotic breakdown, volatilization, or a combination of them, occurredwithin 150 days, with most losses during the first 63 days. Crawford et al.found that the amendment ratio did not affect the extent of degradation ofthe PAHs, and appeared to only slightly decrease the rate of degradation ofsemivolatile compounds with increasing soil content.
After the completion of these preliminary studies, it was observed thatremoval of PAHs from contaminated waste was feasible using composting ap-proaches, although its optimization required adequate oxygen supply, suffi-cient nutrients, and suitable pH, temperature, and moisture for the microbialactivity. The ratio of PAH-contaminated waste in the composting mixtureneeded to be optimized as well to avoid toxicity effects. Attempts to studythe fate of the target PAHs were already observed, with special emphasis onthe entrapment of PAHs in the solid matrix or incorporation into the microbialbiomass.
Civilini (1994) described a laboratory-scale in-vessel composting process(2 kg mixture) operated at a constant temperature of 45◦C during 15 daysfollowing the Finstein approach (Finstein et al., 1996). Civilini (1994) usedmunicipal solid wastes and fertilizer to clean up PAHs present in creosote-contaminated soil, using an optimal ratio of starting material to creosote[80% compost, 5% fresh organic matter, 5% fresh organic mater mixed withfertilizer (NPK 20:50:5), 8% creosote-contaminated soil, 2% fertilizer (NPK20:50:5), 0.2% creosote] to avoid toxicity effects. Water moisture and air-flow were continuously controlled, and samples were taken at days 0, 5,10, and 15. Civilini investigated the fate of 2- to 4-ring PAHs (naphthalene,acenaphthene, fluorene, phenanthrene, anthracene, fluoranthene, pyrene,benzo[a]anthracene, and chrysene). He reported a PAH removal between
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81.63% (benzo[a]anthracene) and 98.63% (fluorene) after 15 days of treat-ment, although the initial PAH concentration was not given. The author ac-counted for volatilization in this study, where it was found to be less than10% of total losses for all the PAHs with the exception of acenaphthene,where it was found to be approximately 54%. The lack of thermophilicphase must be considered in this study, since it allows the developmentof important degraders of PAHs. Civilini (1994) observed a selection of mi-croorganisms during days 5 to 10, in which only sporogenic aerobic and/orfacultative gram-positive bacteria increased and all other groups decreased,maybe caused by the reduction of water-soluble PAHs. During days 10 to 15,a pathogenic problem was identified (Escherichia coli), which confirmed theunsatisfactory sanitization activity of the process at 45◦C (Haug, 1993).
So far, there was not any report on the bioremediation of PAH-contaminated wastes using composting approaches where the mixture wasinoculated. By the mid-1990s, research on biodegradation had shown thatspecial groups of microorganisms (i.e., white-rot fungi) have a remarkablepotential to degrade PAHs. This fungi naturally degrade lignin to obtain thecellulose that is inside wood fiber, but the nonspecific mechanisms, whichenable them to degrade lignin, also allow them to degrade a wide range ofpollutants.
McFarland and coworkers (Qiu and McFarland, 1991; McFarland et al.,1992; McFarland and Qiu, 1995) investigated the removal of PAH in a contam-inated soil using fungi. They investigated the fate of benzo[a]pyrene in a siltloam soil at laboratory scale under an in-vessel-composting regime (reactorvolume, 125 ml) in the presence and absence of Phanerochaete chrysospo-rium (McFarland and Qiu, 1995). The soil spiked with benzo[a]pyrene(150 mg kg−1) was amended with corncobs (primary growth substrate) usinga soil to amendment ratio of 2:1 (dry weight), and the reactor head space wasperiodically purged with humidified oxygen to keep aerobic conditions andwater moisture. PAHs were also monitored in HgCl2 (4%)-treated systems tocompare the impact of biotic and abiotic processes. Information on the tem-perature profile during composting was not given. Samples were taken after1, 7, 14, 21, 28, 35, 84, 91, and 95 days. This study showed that although thebenzo[a]pyrene appeared to be removed, there was not appreciable differ-ence between the uninoculated and inoculated systems with 65.6 and 62.8%removal, respectively, after 95 days, although initial rates of removal werefaster in the inoculated incubations (Table 6). During poison test conditions,removal of benzo[a]pyrene was observed, which suggested the possibility ofirreversible adsorption of benzo[a]pyrene to compost materials. A substantialconcentration of P. chrysosporium (>1 × 104 CFU g−1) was found in boththe inoculated and uninoculated compost systems at the end of the treat-ment period, which explained the similarity in removal in both systems. Thissuggests that amending soils with suitable fungal substrates may be sufficient
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TABLE 6. Removal of Benzo[a]pyrene (150 mg kg−1) in a Silt Loam Soil During 95-DayComposting Process
Note. Adapted from McFarland and Qiu (1995).a3 ml of 4% mercury chloride.bPhanerochaete chrysosporium BKMF-1767, from the Utah State University Biotechnology Center.Inoculated at 39◦C.
to encourage the growth of P. chrysosporium populations already present insoil. Other studies have found that autochthonous microflora, with no in-troduction of foreign microorganisms, offers the greatest potential for PAHdegradation in contaminated soils when an organic substrate is added (Eggenand Sasek, 2002; Canet et al., 2001; Eggen and Sveum, 1999).
McFarland and Qiu (1995) reported the loss of benzo[a]pyrene byfirst-order kinetics during composting with rate constants of 0.003 day−1,0.08 day−1, and 0.06 day−1 for poisoned control, fungal unamended, andfungal amended treatments, respectively. Analysis of gaseous traps indicatedthat there was no loss through volatilization or mineralization and that nearly100% of the benzo[a]pyrene removed was attributable to bound residues asthe parent compound (∼60%) or as chemical intermediates. Furthermore,this study highlighted that the presence of the fungi increased the rate ofbound residue formation in the first 30 days of the composting study, wherethe rate increased from 0.73 mg kg−1 day−1 in the absence of the fungi to1.58 mg kg−1 day−1 in the presence of the fungi. The authors concludedthat the bioaugmentation of a soil-composting system with P. chrysosporiumwas ineffective in degrading benzo[a]pyrene during 95 days of incubation.However, in terms of “locking up” the PAH within the compost matrix, thistechnique proved very successful, although the long-term implications forthe fate of benzo[a]pyrene are unknown. In addition, when bioaugmenta-tion is practiced the rates of nutrient uptake may increase due to a highermetabolism of the target contaminant, which might result in nutrient-limitingconditions resulting in a reduction in microbial activity during long-term soiltreatment.
Similarly, Joyce et al. (1998) investigated the fate of a mixture ofthree- and four-ring PAHs (fluorene, anthracene, phenanthrene, pyrene,
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benzo[a]anthracene) under laboratory-scale in-vessel composting conditionsof spiked simulated solid municipal waste, monitored over a 60-day period(30 days of active composting followed by 30 days of compost maturation).PAHs were also monitored in HgCl2 (2%)-treated systems to compare the im-pact of biotic and abiotic processes. The moisture of the compost was kept to50–60% during the first 30 days and to 30% during the last 30 days. The reac-tors were aerated and/or stirred periodically to maintain aerobic conditions;temperature was kept at 50◦C during the first 30 days and nutrients (0.8%ammonia nitrogen, 2.3% nitrate, 32.9% urea) were added only during the first30 days. The results of this study showed that the loss processes occurred dur-ing the active phase of composting (first 30 days), when all investigated PAHsshowed some abiotic losses in addition to their biological removal from thecompost. Anthracene, phenanthrene, and pyrene were removed during thecomposting process by a combination of biotic and abiotic mechanisms, butbiotic processes were predominant. However, fluorene was readily lost withabiotic processes accounting for approximately 75% of the removal of thisPAH. Benzo[a]anthracene was resistant to biodegradation throughout com-posting, but 40–50% was lost abiotically. These results suggest that the PAHspresent in the contaminated soil should be carefully screened, consideringthe potential for volatilization losses to prevent the process from becomingan air-stripping process.
Kirchmann and Ewnetu (1998) investigated the biodegradation ofpetroleum-based oil wastes in 280-L aerated composting bins using horsemanure as amendment. PAH concentrations in oil sludge and petroleumresidue was 16,800 and 78,500 mg kg−1 dry matter respectively. They stud-ied 4 treatments, using 1.8% oil sludge (5400 mg PAH kg−1 dry matter), 2.1%petroleum residues (6100 mg PAH kg−1 dry matter), 7.1% petroleum residues(1000 mg PAH kg−1 dry matter) and 7.0% paraffin oil (5300 mg PAH kg−1
dry matter). During all the treatments, the temperature increased from 15to 30◦C during the first 12–17 days and then decreased to 22–25◦C. In thetreatment using 7.1% petroleum residues, additions of horse manure wereapplied repeatedly, which resulted in repeated increases in the temperaturefollowing each addition. At the start of the composting period, the dom-inant PAH was naphthalene, which accounted for more than 50% of themeasured PAHs. After 135 days of treatment, Kirchmann and Ewnetu (1998)reported that the majority of the PAHs were removed with measurable quan-tities close to or below the detection limit of 0.1 mg kg−1 dry matter, butpyrene, chrysene, and dibenzo[a,h]anthracene were only partially removed(10%) and thus still present at concentrations of 0.2–0.8 mg kg−1 dry matter.Thus, Kirchmann and Ewnetu (1998) proved that composting PAH-oil wasteswith horse manure was a suitable environmental approach. Horse manure ishigh in lignocellulosic residues, and at the temperatures cited considerableimpact would have expected from the natural mycoflora including membersof the Basidiomycotina.
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Loser et al. (1999) investigated the removal of PAHs during compost-ing of wood containing PAHs with liquid pig manure. They compared theremoval of PAHs in artificially contaminated pine wood (1000 mg phenan-threne kg−1 and 1000 mg pyrene kg−1 aged by autoclaving) and in real PAH-polluted waste wood (5485 mg PAH kg−1) in a pilot-scale percolator system.They inoculated both kinds of wood with 50 g decomposed wood kg−1, andmixed with 26 L liquid pig manure each. After 31 days of composting treat-ment, the PAH concentration of the real polluted waste wood was higherthan the concentration of the artificially contaminated pine wood (1470 mgkg−1 to 170 mg kg−1), which was probably due to the lower bioavailabilityof PAHs in the “naturally” polluted waste and thus slower biodegradation ascompared to the artificially contaminated wood. In addition, in pine wood93% of phenanthrene and 90% of pyrene were removed, but in “naturally”polluted waste wood 86% and 32% of phenanthrene and pyrene, respectively,were degraded. Another possible reason for the higher residual hydrocarbonratio in the “naturally” polluted soil could be the presence of higher molec-ular weight PAHs. In the “naturally” polluted soil two- and three-ring PAHswere decreased by 75 to 98% and four-ring PAHs by 40 to 45%, but five-and six-ring PAHs were reduced only by 15%. Loser et al. (1999) concludedthat remediation of PAH-polluted waste wood by means of microorganismsis possible. Despite using a different PAH-contaminated waste to enhancecomposting, Loser et al. (1999) corroborated the suitability of compostingapproaches to bioremediate highly PAH-contaminated wastes.
Potter et al. (1999) investigated the degradation of 19 PAHs from a Reillysoil (creosote manufacturing and wood preserving) during in-vessel compost-ing at the laboratory scale. Each of the test conditions in their experimentsutilized a 70% soil and 30% corncob mixture on a dry weight basis, amendedwith cow manure, a modified OECD fertilizer, or activated sludge to adjustthe nutrient content (CNP 100:5:1). Moisture content in each 208-L compostreactor was adjusted to 30–35%, and aerobic conditions were facilitated bya continuous vertical air flow. Samples for analyses were taken after 7, 14,28, 56, and 84 days of treatment. Temperatures in all reactors increased tothe upper mesophilic and lower thermophilic ranges (41–53◦C) during thefirst 15 days of treatment and subsequently decreased to ambient tempera-ture, which confirmed aerobic conditions. Compost reactors amended withsludge sustained higher biomass concentrations than those amended withcow manure during the first 28 days. In addition, the greatest amount ofbiomass appeared between the 15 days of composting corresponding to thehighest temperatures, and thus greatest aerobic activity. Following 56 daysof composting, all compost reactors contained similar amounts of biomass.Starting concentrations of total PAHs ranged from 1606 to 4,445 mg kg−1, andfinal concentrations ranged from 888 to 1556 mg kg−1 in the reactors. Theyreported that two- to three-ring PAH (small) were removed by an average of87% in all composters after 84 days of treatment, and four-ring PAH (medium)
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TABLE 7. Small, Medium, and Large PAH Concentrations at the Beginning andWeek 12 of Compost Treatments
Day 0 (mg PAH kg−1 ash) Day 84 (mg PAH kg−1 ash)and number of rings and number of rings
Note. Adapted from Potter et al. (1999). Values represent the average of duplicate reactorsrounded to the nearest whole number.aStd nutr + 1% cow manure.bModified OECD + 1% cow manure.cStd nutr + 1% activated sludge.dStd nutr + 5% activated sludge.eStd nutr + 5% autoclaved sludge.
were reduced by an average of 61%; however, none of the amendment condi-tions appeared effective in degrading five- to six-ring PAHs (large) (Table 7).Most of the concentration reduction occurred within the first 28 days of treat-ment, with a plateau forming by 56 days, suggesting first-order kinetics. Potteret al. (1999) reported removal rates of small PAHs ranging from 0.012 day−1
(5% activated sludge) to 0.081 day−1 (OECD corncobs +1% cow manure),and removal rates of medium PAHs ranging from 0.004 day−1 (5% autoclavedsludge) to 0.033 day−1 (1% cow manure) during the first 28 days. Potter et al.(1999) results showed a general similarity of final PAH concentrations acrossall treatments, which might reflect the recalcitrance of PAHs during the com-posting bioremediation process, whereby different types of amendments didnot significantly alter the final results.
Guerin (2000) investigated the removal of PAHs during bioremediationcomparing the use of mesophilic composting of soil with conventional landtreatment or land farming, both of them at field scale. This study was ofspecial importance because it was amongst the first to quantitatively com-pare the treatment of a highly PAH-polluted soil using a composting ap-proach with a different bioremediation technology. The treated contami-nated soil (4.3–6915 mg kg−1 total PAH) had a silty clay texture, visuallycontaminated with tar residues, and it was initially blended with commer-cially based slow-release nutrients. A ratio of green tree waste to manure tosoil of 15:5:80 was used, where green tree waste was fresh (<5 days old)Eucalyptus spp. leaf and stem waste. The initial concentration of naphtha-lene was 180–300 mg kg−1, phenanthrene was present at 70–230 mg kg−1,
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and benzo[a]pyrene was present at 58–71 mg kg−1. The soil compost mix-ture (130 m3) was placed in a windrow 4 m (wide) × 24 m (long) × 1.5 m(high) and regularly mixed during 224 days. The moisture was maintainedat 60–80% of the water-holding capacity of the treatment soils during thecourse of the field program. Soil composting temperature reached maximumtemperature (42◦C) after 35 days, while there was no self-heating of the soilobserved in the land treatment. The soil composting process conditions re-duced the total PAH concentrations to below the target level given by theregulatory body (500 mg kg−1) after 224 days and resulted in a final concen-tration of 120 mg kg−1, which was lower than that obtained by land treat-ment. Losses of the low-molecular-weight PAH from volatilization throughoutthe treatment period, as determined by a portable flame ionization detec-tor, were not detected. After 224 days of composting treatment, there wascomplete removal of the lower molecular weight PAH, 90% degradation ofmedium PAHs and 70% degradation of large PAHs. Indeno[1,2,3-cd]pyreneand benzo[g,h,i]perylene were most resistant to degradation, but approxi-mately 50% of each was lost. Guerin (2000) also investigated changes inmicrobial populations during 224 days of composting. Total heterotrophicpopulations remained in the range 107–109 per gram soil, showing slightlyhigher values in the composting treatment than in the land treatment. Mi-croorganisms capable of utilizing naphthalene (Pseudomonas sp.) remainedin the range of 107–108/g soil in the first 35 days; however, after 224 daysthey considerably decreased as expected due to the reduction in PAH con-centration, and in particular naphthalene concentration. In this study, com-posting proved to be a suitable field-scale environmental technology forPAH-contaminated soil treatment. Operational parameters such as appropri-ate amendment and ratio of amendment to soil, oxygen supply, and mois-ture were critical factors in achieving effective soil composting. Contaminatedsoils can vary greatly in distribution of contaminants; thus, special attentionshould be given to the appropriate soil preparation previous to compostingtreatment.
Ahtiainen et al. (2002) constructed two pilot compost piles (5 m3)with soil from a sawmill area heavily contaminated with creosote oil(23,600 mgPAHs kg−1 fresh weight soil) and metals (As, Cr, Cu), and theyadded spruce bark chips (mixture rate not available) as a bulking agent. Onepile was inoculated with Mycobacterium, and the other pile was left uninoc-ulated. At the start of composting, most of the PAHs (about 78%) were smallPAHs and hence rather quickly biodegradable. However, the total amountof PAHs was high enough to cause potential inhibition of biological activity.From their results, it was observed that soil native microbes were able todegrade PAHs under composting conditions (Table 8). Furthermore, inocula-tion of the compost with known PAH degraders did not speed up the processmarkedly, as some years before McFarland and Qiu (1995) reported using
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TABLE 8. Concentration and Reduction of PAHs after 17 Months of Composting Treatment in aControl Pile Compared to Mycobacterium-Inoculated Piles with and without Soil Pretreatmentwith Hydrogen Peroxide
Concentration, mg kg−1 (%PAH removal)
Compost PAHs Initial After 5 months After 17 months
P. chrysosporium as the inoculate. The increase in certain PAHs was probablydue to the heterogeneous nature of the compost. These results support therecommendations given by Guerin (2000) regarding the importance of soilhomogenization previous to composting treatment.
Ahtiainen et al. (2002) also investigated the removal of PAHs duringcomposting on a larger scale, constructing one compost pile (100 m3) withsoil from an old wood-preserving facility heavily contaminated with creosoteoil (10,960 mg PAHs kg−1 fresh weight soil) and metals (As, Cr2, Cu). Thesoil was pretreated with 50% hydrogen peroxide to speed up the breakdownof the recalcitrant medium and large PAHs. After the hydrogen peroxide wascompletely degraded, a microbial inoculum of PAH degraders, nutrients, andspruce bark chips were added. The concentration of PAHs is summarized inTable 8. The results of the Ahtiainen et al. (2002) investigation showed thatby pretreating the soil with hydrogen peroxide, they could achieve similarremoval rates in shorter periods of time of about 96% of small PAHs, and57% of medium and large PAHs.
B. Treatment of PAHs-Contaminated Wastes With Compost
Compost can also be added to contaminated waste after its maturation forbioremediation purposes. The composts sustain populations of microorgan-isms with the potential to degrade a variety of organic contaminants (Mahroet al., 1994), and they can improve the contaminated soil environment for
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TABLE 9. Degree of Compost Maturity
Degree of Maximum O2 consumptionmaturity temperature (◦C) (mg g−1) Material status
indigenous or introduced microorganisms by changing the soil pH, nutrientstatus, aeration, and moisture retention characteristics.
The maturity of compost, distinguishing mainly among fresh and maturecompost, is important for application purposes (Korner et al., 2003). Usually,the degree of compost maturity is calculated by the maximum self-heatingtemperature in an isolated vessel (suitable for large-scale composting) or res-piration activity in the respirometer over a period of 4 days (suitable for lab-scale composting). Table 9 summarizes the relationship between the degreeof maturity, temperature, and oxygen consumption. In general, both meth-ods are comparable for mature compost, although during early and middlestages of composting the degree of maturity may vary and a classificationwith regard to the material status could be inaccurate.
One major concern of using compost as a bioremediation approach isthe problem of mixing noncontaminated material with contaminated soil, re-sulting in a greater quantity of contaminated material if the treatment doesnot succeed. In fact, dilution has frequently been practiced as a simple meansof getting contaminated sites within regulatory limits. Research in using com-post as a bioremediation approach is limited and most of it has been doneon the laboratory scale. A review of investigations on treatment of PAH-contaminated wastes with composts is presented chronologically.
Martens (1982) investigated the changes in concentration of four- to six-ring PAHs in fresh compost and in mature compost, which had been allowedto mature for 90–365 days in stacks. Information on the compost technologyand scale was not reported. Martens (1982) found a higher removal of four-to six-ring PAHs in the mature compost than in the fresh compost. Addi-tionally, when these composts were inoculated with 14C-labeled anthracene,benzo[a]anthracene, benzo[a]pyrene, and dibenzo[a,h]anthracene, higher lev-els of mineralization to 14CO2 in the mature composts were found. Maximalremoval for the four-ring 14C-labeled PAHs in fresh and mature compostwere 19% and 62% for anthracene, 8% and 58% for benzo[a]anthracene, 0.5%and 19% for benzo[a]pyrene, and 1.4% and 21% for dibenzo[a,h]anthracene,respectively, over a 70-day incubation period. From this early study, itseems that mature compost may be more efficient during treatment of PAH-contaminated wastes.
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Mahro and Kastner (1993) investigated the fate of pyrene in soil andsoil–compost mixtures over a period of 100 days, finding that the degra-dation of pyrene was enhanced significantly with the addition of maturecompost with >80% pyrene removal after 20 days in the presence and <5%pyrene removed in the absence of the compost. Further, Kastner et al. (1995)investigated the impact of mature compost addition on the fate of 14C-labeledanthracene in soil at laboratory scale. They used 3-L closed compost reactorsincubated at 21 ± 2◦C, continuously aerated with humidified air to keep 60%moisture. In soil–compost incubations, 23% of the 14C-labeled anthracenewas mineralized to 14CO2 and 42% was irreversibly sequestered/bound tothe soil–compost matrix after 103 days (they suggested biogenic binding).However, in soil-only incubations, approximately 88% of the PAH was re-coverable by solvent extraction, with the formation of bound residues beingless significant.
Following this investigation, investigations showed an interest in the un-derstanding of the interaction between the PAHs and the soil–compost ma-trix on biodegradation and binding mechanisms. Thus, Kastner and Mahro(1996) continued this work by investigating the degradation of naphthalene(500 mg kg−1), phenanthrene (100 mg kg−1), anthracene (100 mg kg−1), flu-oranthene (100 mg kg−1), and pyrene (100 mg kg−1) in soil (Ah horizon ofa para brown soil at a noncontaminated, rural area) and soil–compost (soilto compost weight ratio 3:1) incubations (25◦C, 60% of the water-holdingcapacity) at laboratory scale during 100 days. The authors found that thepresence of the compost enhanced the removal of the PAHs and that thepresence of the organic matrix of the compost was essential for enhanceddegradation. In contrast to the pure soil, naphthalene, phenanthrene, andanthracene were degraded after 20 days in the soil–compost mixture. Flu-oranthene and pyrene showed lag phases of 10 days, and then completedegradation occurred after 35 days. The authors suggested that the stimulat-ing effect on the PAH degradation was a function of the organic matrix ofthe compost (humic substances) and the ecological niches of the compost,which have to be colonized by the respective microorganisms. Kastner andMahro (1996) suggested that the addition of compost to contaminated soilsmay enhance the biodegradation and bioavailability of PAHs, retain certainvolatile PAHs, and reduce the sorptive effects in soils, which prevents thecompounds from being analytically detected.
Kastner et al. (1999) investigated the fate of [14C]anthracene(100 mg kg−1) in soil (particle size <2 mm) and in soil–compost mixtures(particle size <4 mm) in a continuously aerated bioreactor at laboratory scaleduring 176 days. Soil and compost were mixed at a ratio of 4:1 (dry weight),with a moisture content adjusted to 60%. They reported complete transforma-tion of the parent compound (anthracene). Although the amount of organiccarbon, which might act as an additional binding substrate, was larger inthe soil–compost mixture (12.7%) than in the native soil (1%), Kastner et al.
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(1999) reported less formation of bound residues from [14C]anthracene and ahigher mineralization in the soil-compost mixture (67.2% mineralized, 20.7%transformed into bound residues) than in the native soil (43.8% mineralized,45.4% transformed into bound residues).
Haderlein et al. (1999) also investigated the impact of humic matterpresent in the compost on the mineralization of PAH-contaminated wastes.They prepared mixtures of 640 g of a silty soil contaminated with aliphatichydrocarbons (total petroleum hydrocarbons (TPH) = 40,000 mg kg−1) andPAH (630 mg kg−1), 250 g maple leaves, 750 g alfalfa, and 80 g (w/w) CaCO3
with a moisture content about 50% (w/w), incubated at 55◦C, 50% mois-ture, and aerated either continuously or intermittently during 35 days andleft to mature for 90 days at ambient temperature. This mature compostedPAH-contaminated soil had a pyrene concentration of about 16 mg kg−1.Then Haderlein et al. (2001) mixed this resulting mature compost with PAH-contaminated soil (80% soil, 20% compost) and left the mixtures for furthercomposting during 100 days. Abiotic controls contained 0.4% NaN3. Pyrenewas rapidly mineralized (>50% mineralization after 15 days), whereas min-eralization in unamended soil was limited to <3% during the same period.Abiotic controls had a maximum total mineralization of 0.7% of the initiallyadded pyrene. Handerlein et al. (2001) focused their efforts to elucidate thelink between PAH mineralization and humic matter based on their prelimi-nary studies, where they found that pyrene mineralization potential duringthe composting of contaminated soil increases with time, as does humifica-tion. The addition of humic acid (previously extracted from the mature com-post) to the soil–compost mixture enhanced pyrene mineralization, reachingan increase of 18 ± 14% mineralization values at the end of the experiment.However, the addition of fulvic acid (previously extracted from the maturecompost) inhibited pyrene mineralization, probably due to the high contenton mineral salts remaining in the fulvic acid after extraction and high pH (8–10). From this study, it was corroborated that although humic acid content islikely not to be the sole reason why the addition of compost stimulates PAHmineralization, it is a major reason. This is probable due to the increasedbioavailability of contaminants sorbed to mineral–humic acid complexes assuggested by Laor et al. (1999) and that the sorption of the microorgan-isms and the PAHs to the colloidal surfaces of humic matter stimulates PAHbiodegradation.
Wischmann and Steinhart (1997) investigated the removal of PAH/N-PAHin soil–compost mixtures (415 g mixture) as compared to unamended soil(Ah/Al horizon soil material; 400 g) in a 1-L bioreactor at laboratory scalefor 180 days. Soil and soil–compost mixtures were spiked with a PAH/N-PAH standard solution in dichloromethane, resulting in concentrations from28 to 181 mg kg−1 dry weight. The moisture of the mixture was kept to50% during the length of the treatment. The compost used in this inves-tigation had a degree of maturity V. Abiotic removal of PAH/N-PAH was
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investigated in poisoned controls (soil autoclaved 35 min at 130◦C, 1.7 bar,1 g kg−1 HgCl2). Samples were taken after 1, 3, 8, 14, 21, 28, 42, 54, 77,105, and 180 days. In unamended soils, only PAHs up to three-ring PAHwere degraded over 105 days. In the soil–compost mixture there was a lagphase of 8 days, and then two-ring PAHs were depleted within the follow-ing 49 days. Three-ring PAHs were eliminated to <3% during 105 days. Ofthe PAHs with more than three-rings, only fluoranthene and pyrene werealmost completely transformed within 105 days. The residual concentrationof benzo[a]anthracene, chrysene and benzo[a]pyrene decreased to 2, 3, and27% compared to the initial estimated amounts within 180 days of treat-ment. PAH removal in the poisoned soil showed similar elimination rates tothose in the unamended soil, which suggested abiotic losses predominatedin the unamended soil. Microbiological analyses indicated nonsterile condi-tions in the poisoned soil–compost mixture, which suggested potential bioticlosses, although they occurred to a lesser extent as compared to the soil–compost mixtures. Longer treatment times were required in the investigationby Wischmann and Steinhart (1997) than in the investigation by Kastner andMahro (1996) for complete removal of two- and three-ring PAHs, althoughboth investigations used mature compost and not aged soils. These differentresults might be explained by the use of different operational parameters anddifferent origins of soil and compost.
Carlstrom and Tuovinen (2003) investigated the mineralization ofphenanthrene in yard-waste compost in biometers to assess the impact ofthe origin of the compost on the mineralization of PAHs. The compost wascollected in four different sampling events from the interior thermophiliczone and from the exterior mesophilic zone of compost piles between 3 and6 months old. Replicates of 5 g compost were spiked with phenanthrene(100 mg kg−1) and incubated in the dark at 22 ± 2 or 60 ± 2◦C during90 days. Carlstrom and Tuovinen (2003) reported a dominant effect due tothe heterogeneity of the yard-waste samples, which obscured the possible ef-fect of surfactant addition, particle size, and moisture content. Nevertheless,they reported an effect of the origin of the compost on the mineralizationof phenanthrene. Yard-waste compost samples collected from the 50–60◦Cthermophilic interior zone and incubated at 60 ± 2◦C yielded 1–2% min-eralization, whereas their incubation at 22 ± 2◦C resulted in 17% averagemineralization of phenanthrene. These results suggested that phenanthrene-mineralizing microorganisms present in the biometer were not thermophiles,but were able to survive in the thermophilic zone of the compost, proba-bly spore formers. Carlstrom and Tuovinen (2003) were probably the firstto report thermophilic biomineralization of PAHs. Samples collected from anouter 30–40◦C mesophilic zone and incubated at 60 ± 2◦C showed negligi-ble mineralization, while parallel subsamples incubated at 22 ± 2◦C yieldedabout 8% mineralization. The removal of phenanthrene was attributed to mi-crobial activity because sterile samples showed negligible 14CO2 evolution
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(<1%). The rate constants were in the range of 0.083–0.033 day−1 for themineralizable fraction of phenanthrene.
Sasek et al. (2003b) investigated the bioremediation of an MGP site soilcontaminated with PAHs by amending with mushroom compost in mid-phaseI (consisting of wheat straw, chicken manure), and gypsum during 54 daysin a thermally insulated composting chamber (∼1000 kg mixture) followedby a further 100 days of maturation in windrows. The total concentration of12 U.S. EPA PAHs in the soil was 610 mg PAH kg−1 dry mass of soil. Themixture comprised 64% soil and 36% compost on a dry weight basis, andthe mixture moisture was maintained at 64%. Changes in the temperatureof compost were monitored, showing an initial increase of temperature upto 70◦C, and after 12 days of composting the temperature progressively de-creased, indicating the different stages of composting. The degradation ofindividual PAHs was in the range of 20–60% at the end of 54 days of com-posting, followed by further PAH removal (37–80% maximum) after another100 days of maturation (Table 10). During composting, the outgoing air waspassed through a filter and the filter was analyzed for possible volatilizationlosses of PAHs. The amount of PAHs retained in the filter was below de-tection limits, indicating that the removal of PAHs during composting wasdue to either compost microflora metabolism or irreversible sorption to thecompost matrix.
Sasek and coworkers investigated the same composting/compost ap-proach in the treatment of a soil collected from an area of a former tar-producing plant with a total concentration of 16 PAH listed by the U.S. EPA
TABLE 10. Concentration of PAHs in Manufactured Gas Plant Soil (610 mg kg−1)before and after Composting and after Maturation
mg PAH kg−1 dry soil (SD) (%PAH removal)
Soil–compost After composting After maturationPAHs mixture 0 days 54 days 54 + 100 days
at a higher concentration of 2832 mg kg−1 (Cajthaml et al., 2003). The mix-ture comprised approximately 47% soil and 53% compost on a dry weightbasis, and the mixture moisture was maintained at 64%. Their results showed42–68% removal of three- to four-ring PAHs, and 35–57% removal of five- tosix-ring PAHs after 42 days of composting. However, an additional 100 daysof compost maturation in open air did not result in a further decrease of PAHconcentrations (Table 11).
Lau et al. (2003) investigated the effect of temperature on the biodegra-dation of naphthalene, phenanthrene, benzo[a]pyrene, and benzo[g,h,i]-perylene, using a compost approach in the laboratory. One gram of steri-lized garden soil was spiked with 1 ml acetone containing the PAHs andmixed with straw spent mushroom compost (Pleurotus pulmonarius), whichis a combination of wheat straw, dried blood, horse manure, and and groundchalk composted together. The moisture content of the mixture was adjustedto 60%, and the sample was incubated at 4 to 80◦C at 200 rpm. Removalof the PAH under investigation occurred due to biodegradation and sorp-tion mechanisms. Under the experimental conditions of 1% spent mushroomcompost treating 100 mg PAH L−1 at room temperature, the removal of PAHsvaried between 82% naphthalene and 59% phenanthrene. The highest sorp-tion removal (46%) was with phenanthrene. Lau et al. (2003) reported anincrease in PAH removal as temperature was increased. At 50◦C, three PAHsbut not phenanthrene were completely removed. At 80◦C, 5% of the spentmushroom compost completely degraded the four PAHs at 200 mg kg−1 soil.Thus, this mechanistic study indicates that increasing the temperature during
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bioremediation of PAH-contaminated waste using compost enhance the re-moval of PAHs, reporting an optimal removal at 80◦C.
IV. DISCUSSION AND CONCLUDING REMARKS
Most investigations of the bioremediation of PAH-contaminated waste us-ing composting approaches conducted to date have been at laboratory orpilot-plant scale. Furthermore, efforts have focused on operational consider-ations rather the fundamental physical, chemical, and biological mechanismsthat underpin bioremediation-composting technologies. Results from theseinvestigations were difficult to compare due to the use of different experi-mental conditions—(1) temperature, (2) moisture, (3) soil/waste to amend-ment/compost ratio, (4) aeration, (5) inoculation versus noninoculation—and to different composting technologies, with windrows versus in-vesselapproach, and different scales of operation: laboratory versus pilot versusfield. Treatment of wastes/soils of different origin with different total PAHsconcentration, distribution of PAHs, bioavailability of PAHs, homogenization,variations in sources, and types of organic wastes were further complications.
Most of the reviewed investigations, regardless of the composting/compost approach, technology, or scale used, agreed on the importance ofoptimizing the operations conditions during the application of the biore-mediation technology. Emphasis is given to temperature, moisture, andmaintenance of aerobic conditions, but particularly to the ratio of wasteto amendment used. The majority of the laboratory-scale investigations onthe use of composting/compost of PAH-contaminated wastes were kept inthe mesophilic or lower thermophilic ranges (Joyce et al., 1998; Kirchmannand Ewnetu, 1998; Potter et al., 1999), with a few investigations in the ther-mophilic phase (Carlstrom and Tuovinen, 2003; Lau et al., 2003). Moisturediffered from one investigation to another, ranging from 40% to 80%, althoughthe majority of the studies emphasized the importance of keeping the mois-ture of the mixture constant throughout the treatment. Different technolo-gies require different approaches to aeration, but has commonly achievedby aerating the mixture with humidified air. By doing so, both moisture andaerobic conditions are maintained. Although some of the earliest work oncomposting/compost bioremediation approaches suggested the importanceof finding a suitable mixture ratio between the contaminated waste and theamendment/compost (Crawford et al., 1993) there are still few reports wherethe optimal mixture ratio has been investigated (Civilini, 1994), and in gen-eral it does not exceed 80% soil content on a dry weight basis (Haderleinet al., 1999; Kastner et al., 1999).
Composting of PAH-contaminated wastes has received more attentionthan the treatment of PAH-contaminated wastes with compost. One majorconcern of using compost as a bioremediation technology is the problem
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of mixing noncontaminated compost with contaminated waste, resulting ina greater quantity of contaminated material if the treatment does not suc-ceed. Nevertheless, if using compost as a bioremediation technology provesto achieve similar removal levels of PAHs as using composting as a bioreme-diation technology, the application of compost bioremediation technologymay offer important operational advantages (i.e., homogenization) duringthe application of the technology at a field scale. Consequently, more mech-anistic studies comparing the bioremediation of PAH-contaminated waste us-ing composting and compost approaches applying the same technology andscale at optimal operational conditions are hitherto necessary. If fresh com-posts are mixed with the soil, a composting process may occur (Cajthamlet al., 2003; Sasek et al., 2003b), and then a more appropriate definitionwould be composting/compost approach rather than composting or com-post approach.
Composting and the use of compost have both been successfully appliedto the bioremediation of PAH-contaminated wastes. The main mechanisms ofPAH removal under both bioremediation regimes were mineralization, bind-ing, and volatilization. Mineralization was the more significant of the mech-anisms of removal reported in most studies to date. Mineralization of PAHsmay be enhanced by increasing their availability to microbial attack duringbioremediation processes. In order for the pollutants to become availableto microbial attack, desorption from the waste–compost matrix may have totake place. In terms of composting, increases in temperature may enhancethe rate of PAH desorption from the matrix and thus transfer to the aqueousphase (Pignatello and Xing, 1996). In terms of compost, enhanced PAH min-eralization of PAHs present in soil–compost mixtures may occur due to thepresence of humic matter in the compost matrix. This is supported by stud-ies where either humic acid and/or fulvic acid has been added to a soil withvery low endogenous humic acids/fulvic acids (Bogan and Sullivan, 2003;Conte et al., 2001; Haderlein et al., 2001; Laor et al., 1999) and an increasein the bioavaliable fraction has been observed. Another favorable factor in-herent to composting approaches is that normally large amounts of organicmatter, considered a major factor in the “locking up” of organic pollutants,are added to the system (Brusseau et al., 1991; McFarland and Qiu, 1995).Thus, by utilizing composting approaches, the bioavailable fraction will ide-ally be mineralized and the unavailable fraction will be “locked up” in thesoil–compost matrix, reducing the overall risk. However, the long-term fateof the “locked up” organic contaminants remains uncertain.
While more research on comparisons of composting/compost bioreme-diation technologies versus conventional technologies remains limited, theinvestigations reviewed have proved that composting, regardless of the ap-proach used, is a good environmental technology that may be used to re-move PAH from contaminated wastes. These bioremediation technologieshave proved to be beneficial for the amelioration of contaminated wastes by
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reducing the concentration of PAHs, which promotes soil sustainability andsoil reuse in contrast to landfill or incineration approaches. Guerin (2000)reported a higher PAH removal from contaminated soils using compostingapproaches compared to conventional landfarming.
Further study should be focused on the optimization of operational pa-rameters at laboratory and pilot scales using naturally PAH-contaminatedwastes, which will facilitate the highest possible removal of the target PAHs.Complementary studies will be needed to obtain a better understanding ofthe different mechanisms contributing to PAH mineralization during compost-ing, and in particular of the interactions between the different componentsin the waste–compost matrix together with a greater understanding of themicroflora involved with particular emphasis on the impact of lignocellulyticfungal decomposers. Commercial industrial-scale operations will undoubt-edly benefit from the results obtained from such studies.
ACKNOWLEDGEMENTS
We are grateful to Cleanaway Ltd and London Remade for providing supportfor this study through the Entrust scheme.
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