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Capsule: Hydrophobicity and solid water distribution ratios inuence the interactionbetween PAHs and soil organic matter, thereby aecting microbial degradation.

1Abstract

Interest in the biodegradation mechanisms and environmental fate of polycyclic aromatic hydrocarbons (PAHs) is motivated bytheir ubiquitous distribution, their low bioavailability and high persistence in soil, and their potentially deleterious eect on human

health. Due to high hydrophobicity and solidewater distribution ratios, PAHs tend to interact with non-aqueous phases and soilorganic matter and, as a consequence, become potentially unavailable for microbial degradation since bacteria are known todegrade chemicals only when they are dissolved in water. As the aqueous solubility of PAHs decreases almost logarithmically withincreasing molecular mass, high-molecular weight PAHs ranging in size from ve to seven rings are of special environmental

concern. Whereas several reviews have focussed on metabolic and ecological aspects of PAH degradation, this review discusses themicrobial PAH-degradation with special emphasis on both biological and physico-chemical factors inuencing the biodegradationof poorly available PAHs.

2004 Elsevier Ltd. All rights reserved.

Keywords: Polycyclic aromatic hydrocarbons; Biodegradation; Bioavailability; Hydrophobicity; Persistence

1. Introduction

Polycyclic aromatic hydrocarbons (PAHs) are uniquecontaminants in the environment because they aregenerated continuously by the inadvertently incompletecombustion of organic matter, for instance in forest

res, home heating, trac, and waste incineration.Massive soil contamination with PAH originated fromextensive industrial coal gasication during most of the20th century. As gas works were typically located indensely populated urban regions to facilitate thedistribution of the coal gas, PAH contaminated sitesare mostly found in or near cities, thus representinga considerable public health hazard. PAHs are com-posed of fused, aromatic rings whose biochemicalpersistence arises from dense clouds of p-electrons on

1 Present address: Geological Survey of Denmark and Greenland,

Department of Geochemistry, :ster Voldgade 10, Kbenhavn K,Principles of microbial

Anders R. Johnsena,1, Lukas

aNational Environmental Research Institute, Department of En

PO Box 358, DK-40bSwiss Federal School of Technology Lausanne (EPFL), Sc

CH-1015 Lausa

Received 12 January 200

Environmental Pollution1350, Denmark.2 Present address: UFZ Centre for Environmental Research,

Department of Environmental Microbiology, Permoserstrae 15,

D-04318 Leipzig, Germany. Tel.: +49-341-235-2225; fax: +49-341-

235-2247.* Corresponding author. Tel: C41-21-6933773; fax: C41-21-

6935670.

E-mail address: hauke.harms@ufz.de (H. Harms).

0269-7491/$ - see front matter 2004 Elsevier Ltd. All rights reserved.doi:10.1016/j.envpol.2004.04.015AH-degradation in soil

. Wickb,2, Hauke Harmsb,2,*

onmental Chemistry and Microbiology, Frederiksborgvej 399,

Roskilde, Denmark

ol of Architecture and Civil and Environmental Engineering,

e, Switzerland

accepted 13 April 2004

33 (2005) 71e84

www.elsevier.com/locate/envpolboth sides of the ring structures, making them resistantto nucleophilic attack. Besides this, they possess physicalproperties, such as low aqueous solubility and highsolidewater distribution ratios, which stand againsttheir ready microbial utilization and promote theiraccumulation in the solid phases of the terrestrialenvironment. As rules of thumb, the aqueous solubility

entaand, as a consequence, the bioavailability of the PAHsdecreases almost logarithmically with increasing molec-ular mass. Of environmental concern are primarily thePAHs ranging in size from naphthalene (two rings,C10H8) to coronene (seven rings, C24H12). This reviewdiscusses the microbial degradation of environmentallyrelevant PAHs with an emphasis on biological strategiesto obtain poorly bioavailable, soil-sorbed or non aque-ous phase PAHs.

2. Growth on PAHs as sole carbon sources

Microbial degradation of PAHs and other hydro-phobic substrates is believed to be limited by theamounts dissolved in the water phase (Ogram et al.,1985; Rijnaarts et al., 1990; Volkering et al., 1992;Volkering et al., 1993; Harms and Bosma, 1997; Bosmaet al., 1997), with sorbed, crystalline, and non-aqueousphase liquid (NAPL)-dissolved PAHs being unavailableto PAH-degrading organisms. Bioavailability is consid-ered a dynamic process, determined by the rate ofsubstrate mass-transfer to microbial cells relative totheir intrinsic catabolic activity (Bosma et al., 1997;Harms and Bosma, 1997). It has been described bya bioavailability number, Bn, (Koch, 1990; Bosma et al.,1997), which is a measure of a microorganisms sub-strate degradation eciency in a given environment. Bnis dened as the capacity of an organisms or a pop-ulations environment to provide a chemical, divided bythe capacity of the organism or population to transformthat chemical. At high mass transfer rates, the overallbiodegradation rate is controlled by the metabolicactivity of the bacteria (Bn> 1), i.e. by both the specicactivity of the cells and the population density. AtBn 1, the biodegradation rate is equally controlled bythe physical transport and the microbial activity. Whenthe transport of the substrate decreases or the bacterialpopulation grows, the mass transfer becomes the factorthat limits the biodegradation (Bn! 1).

Bacteria growing in suspended, shaken cultures withcrystalline PAH in amounts exceeding the aqueoussolubility as the sole source of energy and carbon exhibitcharacteristic growth curves (Volkering et al., 1992;Volkering et al., 1993; Wick et al., 2001a; Mulder et al.,1998), which can be divided into three phases: (i) anexponential phase, (ii) a subsequent phase with pseudo-linear growth and nally (iii) a pseudo-stationary phase.A schematic illustration of the growth curves and someculture parameters is given in Fig. 1, whereas Fig. 2 showsan experimentally observed growth curve ofMycobacte-rium sp. LB501T with solid anthracene. The initial,exponential phase (I in Figs. 1 and 2) is similar to theexponential phase in growth curves obtained with highlywater-soluble substrates. Truly exponential growth, i.e.

72 A.R. Johnsen et al. / Environma constant specic growth rate over several generations,can be explained in two ways: in one scenario, thedissolved PAH concentration saturates the bacterialuptake system. The bacterial PAH-uptake remains sat-urated because PAH-dissolution is fast enough to keepup with the rising substrate consumption by thegrowing population. In this case, bacteria grow expo-nentially at their physiologically limited maximum rate.When the PAH consumption by the increasing popula-tion exceeds the PAH dissolution rate, the dissolved PAHconcentration drops below saturation and exponentialgrowth ceases.

I II III

Time

Biom

ass,

spe

cific

grow

th ra

te,

diss

olut

ion

rate

, diss

olve

d co

ncen

tratio

n

0

Biomass

Specific growth rate

Dissolution rate

Dissolved concentration

Fig. 1. Schematic representation of the four parameters biomass,

specic growth rate, PAH dissolution rate and PAH dissolved

concentration of a bacterial batch culture growing on solid PAH. A

short exponential phase (I) is followed by a prolonged pseudo-linear

phase (II) which approaches a pseudo-stationary phase (III). Details

are given in the text.

0.1

1

10

100

0

0.5

1

1.5

2

2.5

3

0 50 100 150 200

Bioa

vaila

bility

num

ber (

-) optical density at 578 nm

time (h)

pseudo-lineargrowth

II

exponentialgrowth

I

mass transfer limitation

microbial uptake limitation

pseudo-stationaryphase

III

Fig. 2. Batch growth (solid line) of Mycobacterium sp. LB501T on

30 g L1 of anthracene crystals of 0.2e0.5 mm as a function oftheoretically calculated anthracene bioavailability expressed by the

bioavailability number (Bn; dashed line) (adapted from Wick et al.,

2001a). Exponential growth (I) is reected by BnZ 1 (control bymicrobial activity), whereas pseudo-linear growth (II) and pseudo

stationary growth (III) are reected by Bn! 1 (mass transfer

l Pollution 133 (2005) 71e84limitation).

ntaAlternatively, the dissolved substrate concentrationmay not saturate the bacterial uptake systems, butremains constant despite a growing population. In thiscase, exponential growth occurs at a sub-maximum rate.This is known from bacterial growth on liquid hydro-carbons and may be explained by the physical contact ofthe active bacteria with their substrate. The consump-tion-driven, local substrate dissolution remains constantper cell as long as the surface of the substrate is not lledwith attached bacteria (Bouchez-Natali et al., 2001).Whatever the mechanism, the exponential phase in thePAH growth curve is short because at some point, thepotential metabolic demand of the increasing number ofcells exceeds the PAH-dissolution rate (i.e., Bn drops!1).

The subsequent pseudo-linear growth phase (II inFigs. 1 and 2) is explained by a physically limited,maximum dissolution of substrate which is convertedinto biomass. Typically, the PAH-crystals have a lowarea-to-volume ratio which remains constant overlonger periods leading to an almost constant ux ofsubstrate to the bacterial population. The specicgrowth rate decreases as the cell number increasesbecause a growing population shares a constant totalsubstrate ux. True linearity is unlikely because themaximum dissolution rate depends on the specicsurface of the crystals, which is likely to decrease duringdissolution. But the biomass formation will drop witha growing population, even when crystalline substrate isprovided in excess and the dissolution rate remainsconstant. This is because each cell needs substrate forcell maintenance, not resulting in additional biomass.The total maintenance substrate consumption will risewith a growing population and the fraction of substratethat is channeled into biomass formation will constantlydrop and approach zero (Pirt, 1965). In the resultingpseudo-stationary phase (III in Figs. 1 and 2), thesubstrate consumption of individual cells has reachedthe maintenance requirements of the cells and growthceases. The cells are neither deprived of substrate norinhibited by metabolic products as in the stationaryphase of a classical growth curve, but they reach a quasi-equilibrium where the entire PAH-ux is consumed formaintenance of the cells.

3. Growth on PAH in soil

The growth curves discussed above are obtainedunder idealized conditions that are characterized byunlimited bacterial access to the crystals and fastsubstrate transport by convection and diusion ina homogeneous, aqueous solution. Heterogeneousmedia, such as soil, do not have these characteristics.In soil, PAHs are heterogeneously distributed and may

A.R. Johnsen et al. / Environmebe absorbed inside of organic particles, located in smallpores that are inaccessible for bacteria, or otherwiseoccluded by the multitude of solid soil constituents. Incase of a massive contamination, PAHs will typicallyoccur as tar droplets with a low surface to volume ratio,which further limits the bacterial access to the PAHs.Mixing does not occur in soil and the eective diusionof molecules in soil may be orders of magnitude lowerthan in water, since the diusion is retarded by the solidphases, dead-end pores and the high tortuosity of thesystem. Bacterial cells are excluded from pores smallerthan about 0.2e0.8 mm. In addition, predation isbelieved to reduce bacterial biomass in pores largerthan 2 mm, thereby restricting the bacterial mobility(Postma and vanVeen, 1990; Harms and Bosma, 1997).Consequently, a large fraction of the PAH-degradingbacteria in soil is expected to be physically separatedfrom the PAH-sources, and to depend on diusivetransport of PAHs from the PAH-sources to the cells(Harms and Bosma, 1997). In soil, as opposed to well-mixed aqueous systems, the substrate consumption leadsmuch faster to mass transfer-limited conditions as thenumber of cells increases. Therefore, PAH-degradingpopulations in soil are probably mostly not growing, butthey are in a pseudo-stationary phase where transientgrowth only replaces decaying cells until the habitatsmass transfer-controlled carrying capacity is reachedagain.

In contrast, laboratory studies with PAH-degrading,pure cultures are almost always done with PAHs in highconcentrations. This is an unrealistic scenario in soil assoil-bacteria are generally believed to be carbon andenergy-starved. Under soil conditions, the bacteria maynot utilize a single carbon source, but rather co-utilizea number of available carbon compounds (Egli, 1995;Egli, 2002). The formation of enzymes catalyzing thedegradation of organic pollutants may be repressedwhen the cells grow on other substrates, but theconstitutive background level of expression is oftensucient for immediate consumption of the pollutant ifit becomes available in low amounts (Egli, 2002). Asa consequence, if the cells concomitantly take up severalcarbon sources to maintain their biomass, there will notnecessarily be a threshold concentration of PAH, belowwhich biodegradation stops. It should be noted that theoverall long-term benet to the cell of keeping a baselevel of PAH-degrading enzymes must be higher thanthe cost of producing the PAH-degrading enzymes, forthe trait to be evolutionarily stable.

4. PAH-metabolism

Since bacteria initiate PAH degradation by the actionof intracellular dioxygenases, the PAHs must be takenup by the cells before degradation can take place.

73l Pollution 133 (2005) 71e84Bacteria most often oxidize PAHs to cis-dihydrodiols by

ntaincorporation of both atoms of an oxygen molecule. Thecis-dihydrodiols are further oxidized, rst to thearomatic dihydroxy compounds (catechols) and thenchanneled through the ortho- or meta cleavage path-ways (Cerniglia, 1984; Smith, 1990).

The biological degradation of PAHs can serve threedierent functions. (i) Assimilative biodegradation thatyields carbon and energy for the degrading organismand goes along with the mineralization of the compoundor part of it. (ii) Intracellular detoxication processeswhere the purpose is to make the PAHs water-soluble asa pre-requisite for excretion of the compounds. Gener-ally, it seems that intracellular oxidation and hydroxyl-ation of PAHs in bacteria is an initial step preparingring ssion and carbon assimilation, whereas in fungi itis an initial step in detoxication (Cerniglia, 1984). (iii)Co-metabolism, which is the degradation of PAHswithout generation of energy and carbon for the cellmetabolism. Co-metabolism is dened as a non-specicenzymatic reaction, with a substrate competing with thestructurally similar primary substrate for the enzymesactive site. An example is the co-metabolization ofbenzo(a)pyrene by bacteria growing on pyrene (Boon-chan et al., 2000). Keck et al. (1989) noted that: In thecase of a pure culture, co-metabolism is a dead-endtransformation without benet to the organism. Ina mixed culture or in the environment, however, such aninitial co-metabolic transformation may pave the wayfor subsequent attack by another organism (Kecket al., 1989).

In spite of considerable eort, only a very limitednumber of bacteria have been isolated that can grow inpure cultures on PAHs with ve or more aromatic rings(high molecular weight (HMW) PAHs). A possiblereason is the high retention of these compounds by thesolid soil phase, resulting in mass-transfer rates ofHMW-PAHs to the bacterial cells too low to match thecells basic metabolic requirements. The low bioavail-ability of PAH may have prevented the evolution ofsuitable enzymatic pathways in soil bacteria. Accordingto Perry (1979), recalcitrant compounds generally donot serve as growth substrates for any single microbialorganism, but are thought to be oxidized in a series ofsteps by consortia of microbes.

Environmental, bacterial isolates often degrade onlya narrow range of PAHs (Bouchez et al., 1995) andpatterns of simultaneous degradation of PAH mixturesare complex. Cometabolism and inhibition phenomenahave been investigated by Bouchez et al. (1995). Withmixtures of two individually degradable PAHs, eitherpreferential degradation of one PAH or reduceddegradation rates of both PAHs indicating metaboliccompetition were observed. The interactions were morecomplex when the strains were growing on one PAH(PAH1) and a second non-mineralizable PAH (PAH2)

74 A.R. Johnsen et al. / Environmewas present. (a) Degradation of PAH1 could beunaected with no cometabolism of PAH2, (b) degra-dation of PAH1 could be inhibited with no cometabo-lism of PAH2, (c) PAH2 or derived metabolites could betoxic leading to no metabolism of PAH1, (d) degrada-tion of PAH1 could be unaected with cometabolism ofPAH2, and last, (e) cometabolism of PAH2 could havea synergistic eect increasing the degradation of PAH1.When a second bacterial strain able to degrade PAH2was introduced, the inhibitory eects of PAH2 on strainone were relieved. The authors concluded that degrada-tion of a PAH-mixture appears as a co-operative processinvolving a consortium of strains with complementarycapacities. A later study (Bouchez et al., 1999) showedthat mixed cultures of two or three strains, althoughpossessing the capacity to mineralize each of ve PAHs,achieved limited degradation of a ve-PAH mixture. Incontrast, an enrichment from a PAH-contaminated soilreadily mineralized the ve-PAH mixture. This demon-strates that the appealing idea of releasing bacterialconsortia from the lab to degrade PAH-contaminationsmay not be that ecient in vivo.

5. Bacterial adaptations that maximize the acquisitionof sorbed and separate phase PAHs

In the following, the naturally occurring, rate limitingmass transfer processes will be analysed in terms of thepossibilities of biological systems to optimize the PAHtransfer.

5.1. Diusion

The basic assumption is that microbial degradationof PAHs is limited by the PAHs dissolved in the waterphase (Volkering et al., 1992; Volkering et al., 1993;Harms and Bosma, 1997; Bosma et al., 1997). Thisassumption shall be considered in the context of FicksFirst Law of Diusion. Diusive mass transfer strives toequalize concentration gradients.

Q=t DAC0 Cx=x 1Here, Q is the quantity of substrate (mol) diusing

through areaA (m2) per unit of time t (s). (C0Cx)/x is theconcentration gradient where C0 is the concentration atthe source (mol m3), Cx is the concentration at the sink(mol m3) and x is the diusion path length (m), i.e. thedistance between source and sink. In the case of soil-sorbed PAH, C0 is typically the aqueous equilibriumconcentration of the PAH. D is the diusion coecient(m2 s1) expressing the resistance of the environment todiusion. The negative sign indicates that diusion is inthe direction away from higher concentration. Steadystate is to be expected when sorbents or NAPL provide

l Pollution 133 (2005) 71e84PAHs and bacteria act as a sink, i.e., the bacterial

the cell surface, steeper concentration gradients, and

ntaconsumption of PAH drives the dissolution (or volatili-sation through a gaseous phase) of PAH from the source(Fig. 3). From Ficks Law results that the aqueoussolubility limits the potential size of the gradient.

Ficks Law of Diusion also applies to bacteriagrown on PAHs in shaken batch cultures. The PAH-crystals are surrounded by a stagnant boundary layer ofmedium through which the mass transfer occurs bydiusion. The thickness of the boundary layer dependson the size and shape of the crystals and the shakingvelocity (Mulder et al., 1998). The smaller the crystalsand the higher the shaking velocity, the thinner is theboundary layer and the higher the PAH-ux from thecrystals to the bulk liquid. For naphthalene crystals ina stirred reactor, the thickness of the boundary layer wasfound to vary from 10 mm to 120 mm depending on thestirring velocity (Mulder et al., 1998). We can assumeuniform mixing outside the diusive boundary layer dueto turbulence. If the actively growing cells are dispersedin the bulk liquid, the value of x equals the thickness ofthe boundary layer.

5.2. Bacterial optimization of the diusion coecient(D) and the area (A)

According to Eq. (1), the PAH-ux could beincreased by increasing the diusion coecient D.However, D of individual molecules is a physicalparameter which cannot easily be increased by thebacteria. A biological way of increasing the eective Dof PAH molecules would be the excretion of biosurfac-tants as carriers. In a study of the inuence ofsurfactants on the bioavailability of solid dibenzofuran,Garcia et al. (2001) showed that micelles of thesurfactant Brij 35 facilitated the diusive transfer ofdibenzofuran through the boundary layer around di-benzofuran crystals. Calculations showed, and experi-mental results conrmed, that the eective diusion of

C0

x

Sorbed orentrapped PAH

Soil particles

Diffusion boundary

Cx

Fig. 3. Schematic drawing of a PAH-degrading bacterium growing

close to soil entrapped of sorbed PAH. C0 is the PAH-concentration at

the surface of the particle, Cx is the concentration at the cells and x is

the diusion path length.

A.R. Johnsen et al. / Environmemicellar dibenzofuran was slightly faster than diusionof dissolved dibenzofuran. Surfactant micelles of as-sumed 4! 109 m diameter diused 7.9 times slowerthan a dibenzofuran molecule, while at equilibrium theycontained 14.6 dibenzofuran molecules. The productof dibenzofuran solubilization and reduced diusivity ofmicelles resulted in 1.85 times faster eective diusion ofmicellar dibenzofuran. In contrast, the bioavailabilityof sorbed dibenzofuran was found to be reduced afteraddition of Brij 35. This was explained by the ratelimiting intraparticle diusion of dibenzofuran thatremained uninuenced by the surfactant and the re-duction of the truly water-dissolved dibenzofuranconcentration by solubilization in micelles (Garciaet al., 2001).

Increasing the interfacial area A of the PAH-sourcewould increase the PAH-ux (Mulder et al., 1998). Forbacteria growing on alkanes or NAPL-dissolved PAH inaqueous systems, production of bioemulsiers wouldproduce a larger surface area of these substrates andsubsequently a higher substrate ux (Reddy et al., 1982).For example, addition of small amounts of biosurfac-tants (rhamnolipids) to Pseudomonas cultures growingon octadecane increased the octadecane dispersion byfour orders of magnitude and resulted in a highergrowth-rate (Zhang and Miller, 1992). Willumsen andKarlson (1997) reported that a high percentage of PAH-degraders isolated from PAH-contaminated soil had theability to emulsify diesel fuel when grown on crystallinePAHs, but there was no correlation between PAHdegradation and emulsication. Whereas emulsicationof NAPL-dissolved PAH appears to be a good strategyin aqueous systems with turbulent mixing, its eect isless obvious in soil systems, where physical mixing ismuch smaller.

5.3. Optimization of the concentration at the sink (Cx)

The concentration gradient and thus the ux of PAHcan be maximized by the cells keeping the PAH-concentration at the cell-surface (Cx) close to zero(van-Loosdrecht et al., 1990; Rijnaarts et al., 1990;Harms and Bosma, 1997; Wick et al., 2001b). The rate,at which bacteria degrade a substrate that is available atlow concentration, depends on their specic anity(a0A) towards the substrate. The specic anity isdened as the ratio of the maximal rate of substrateuptake (qmax) and the half saturation constant Km(Button, 1985) and equals the slope of the rst orderpart of the activity-versus-concentration plot. Hence, fora given substrate, a0A determines the eciency ofa bacterium to reduce the substrate concentration atits surface compared to the solubility of the substrateCeq at equilibrium. High specic anities therefore leadto ecient pollutant depletion at low concentrations at

75l Pollution 133 (2005) 71e84

ntahigher substrate transfer rates. This strategy has beendemonstrated for a Mycobacterium sp. growing onanthracene, having a specic anity for anthracene, i.e.a rst-order degradation rate constant, that was 30 timesabove values reported for a toluene-degrading Pseudo-monas and orders of magnitude higher than bacteriagrowing on readily water soluble substrates (Wick et al.,2001a; Wick et al., 2002a).

5.4. Optimization of the diusion path length (x)

From Ficks Law, it is seen that the PAH-ux (Q/t) isinversely proportional to the distance (x) between thePAH-source and the cells. Therefore, biolm formationon PAH-containing sorbents or separate phase PAHs isan ecient way of increasing the PAH-ux to the cells.The small distance between the PAHs and the biolmcells strongly favors the diusive mass-transfer of PAHto the cells by steepening the aqueous concentration-gradient (van-Loosdrecht et al., 1990; Rijnaarts et al.,1990). The thickness of the diusion layer aroundnaphthalene crystals was found to be in the range of 10to 120 mm (Mulder et al., 1998). An attached cellsituated less than 1 mm from a naphthalene crystalcould therefore dramatically increase the naphthaleneux to the cell compared to a cell growing in suspension.It should be noted, however, that when coupled tomicrobial degradation, the actual diusion ux is notinversely proportional to the distance, as Eq. (1) wouldsuggest. Bacteria are not perfect sinks whose uptake ratecan keep up with any substrate ux and the aqueoussubstrate concentration at the cell surface thus rises withincreased mass transfer, thereby partly diminishing thegradient. Attachment to solids containing absorbedcompounds was found to initially exert the samepositive eect. As degradation went on, intra-particlediusion as the rate-limiting step caused increasinglysubstrate-depleted outer regions of the sorbent, con-comitant with increasing distances between substrateand bacteria leading to reduced substrate availability(Harms and Zehnder, 1995).

A recent study suggests that bacterial attachment toPAH-crystals serving as sole carbon and energy sources,may depend on the solubility of the PAH. A gfp-labeledPseudomonas putida was grown in ow-channels withimmobilized uorene or phenanthrene crystals as thesource of carbon (Rodrigues et al., 2003). Confocal laserscanning microscopy revealed that the strain showeddierent behaviors depending on the kind of PAH itconsumed. When grown on phenanthrene, the cellsformed biolm on the crystal surfaces. In contrast, whengrown on the more soluble uorene or on both uoreneand phenanthrene, the strain formed biolm on the glasssurface in the neighborhood of the PAH crystals,indicating that attachment to more soluble substrates

76 A.R. Johnsen et al. / Environmewas more dicult. However, the fact that phenanthrenewas not colonized in the presence of uorene, providingadditional substrate, suggests that adhesion is regulatedas a function of the substrate bioavailability. Sucha mechanism was also suggested by Wick et al. (2001a,2002a), who showed that the anthracene-degradingMycobacterium strain LB501T formed a biolm onanthracene crystals. Cells at the crystal surfacesetched craters in the crystals due to consumption-driven PAH-dissolution on a micro-scale (Wick et al.,2002a) (Fig. 4). The biolm formation by strain LB501Tseems to be a well-regulated process as no biolm wasformed on anthracene in the presence of alternative,soluble carbon sources or when high amounts of solidanthracene, leading to high substrate uxes, weresupplied (Wick et al., 2002a). It was further found thatMycobacterium strain LB501T exhibited specic mod-ications of the cell envelope in response to solidanthracene, such as increased adhesion to hydrophobicsurfaces and changes in its mycolic acid prole (Wicket al., 2002b). This nding is of interest as mycolic acidsare believed to stimulate attachment to hydrophobic

Fig. 4. Scanning electron micrograph of an anthracene surface (A) and

an anthracene surface during early stages of biolm formation of

Mycobacterium sp. LB501T (B). Craters were absent from fresh,

untreated anthracene surfaces and thus can be attributed to anthracene

l Pollution 133 (2005) 71e84consumption by surface-associated cells. Scale bars= 10 mm.

ntasurfaces and hence increase the access to hydrophobicsubstrates (Bendinger et al., 1993; Borrego et al., 2000).

In another study, four sphingomonads, vemycobacteria and a Nocardia sp. were screened forattachment to PAH-crystals during growth on phenan-threne, uoranthene, pyrene or anthracene (Johnsen andKarlson, 2004). The potential respiration of cells attachedto the crystals and of cells living in the planktonic mode(i.e. in the bulk liquid) were tested with the respirationindicator WST-1. It was found for the mycobacteria andthe Nocardia that the more soluble the PAH, the higherthe percentage of respiration of cells in suspension. Sucha split-up in attached andplanktonic sub-populationswasnot found for the sphingomonads. Either all cells wereassociated with the crystals, indicating biolm formation,or all cells were planktonic. Sphingomonads and myco-bacteria growing on NAPL-dissolved PAHs also adherestrongly to the NAPL phase (Pia Willumsen, personalcommunication). Thus, it seems that attachment to- andbiolm formation on PAH-sources is a widespreadmechanism among bacteria to overcome mass-transferlimitations when growing on poorly soluble and stronglysorbed PAHs.

5.5. Release of biosurfactants

A bacterial strategy, which inuences the PAHtransfer in a more complex way, is the release ofbiosurfactants. Biosurfactants are small, detergent-likemolecules with a hydrophilic head and a lipophilic tail.They form spherical or lamellar micelles when thesurfactant concentration exceeds a compound-specic,critical micelle concentration (CMC). Hydrophobiccompounds become solubilized in the hydrophobiccores of the micelles, which leads to a transfer of PAHfrom solid, liquid, or sorbed PAH-pools into the waterphase. In terms of Fickian diusion, biosurfactantmicelles inuence the apparent solubility (C0), theeective diusion coecient (D), as was discussedabove, and both the diusion distance (x) and theinterfacial area (A). These eects are due to dispersionof the PAH. Surfactant molecules may furthermoreinuence the dissolution or desorption process byattaching to the PAHewater interface. Here, they formhemi-micelles which may accelerate the PAH-release(for a review, see Volkering et al., 1998) Moreover,biosurfactants seem to inuence the bacterial uptake ofsolubilized compounds. In many instances the uptake ofbiosurfactant-solubilized molecules was found to befaster than the uptake of truly dissolved (i.e. mono-dispersed) molecules. There are furthermore reports ofspecies-specic and energy-dependent uptake of bio-surfactant-solubilized compounds, which points at a di-rect interaction of biosurfactant micelles with cellmembranes (Beal and Betts, 2000; Noordman and

A.R. Johnsen et al. / EnvironmeJanssen, 2002). Seeing that many biosurfactants repre-sent constituents of cell envelopes (Neu, 1996), thepossibility of a fusion between micelles and cells isindeed not far-fetched.

It is well known thatmanybacteria growing on alkanesproduce biosurfactants to increase the bioavailability ofthese poorly available substrates (Itoh and Suzuki, 1972;Oberbremer andMuller-Hurtig, 1989). A similar strategywas suggested for PAH-degrading bacteria (Deziel et al.,1996). For example, biosurfactant produced by a Pseudo-monas aeruginosa growing on phenanthrene or naphtha-lenewas shown to increase the apparent solubility of thesePAHs, suggesting that themicroorganismwas promotingthe availability of its growth-substrate (Deziel et al.,1996). These results are in conict with earlier resultsshowing that supernatants of a naphthalene-degradingPseudomonas sp. culture did not aect naphthalenedissolution rates (Volkering et al., 1993). Willumsen andKarlson (1997) screened 57 PAH-degrading isolates forproduction of biosurfactants, but found no correlationbetween biosurfactant production and PAH mineraliza-tion. In a more recent study, most of 22 PAH degradersreduced the surface tension ofwater by only 0e4 mNm1

(Johnsen and Karlson, 2004), indicating that biosurfac-tants were either not produced or in concentrations farbelow the CMC. None of the tested sphingomonadsproduced biosurfactants, while a few mycobacteria andnocardia produced somebiosurfactant.However, none ofthe tested isolates were convincing biosurfactant pro-ducers. It seems that biosurfactant production is not verycommon among PAH degraders and obviously notessential for obtaining PAHs under environmentalconditions. The occasional accumulation of biosurfac-tants in aqueous cultures might be a side eect ofregulated bacterial cell surface modications. Biosurfac-tants may change a hydrophobic surface to be hydro-philic, and vice versa (Neu, 1996). It has been suggestedthat these hydrophobicity/hydrophilicity shifts area mechanism for cells to regulate attachment to- anddetachment from substrata (Neu, 1996). Itmaybe that therelease of surface active compounds is a way forhydrophobic cells of Mycobacterium and Nocardia toleave hydrophobic surfaces such as PAH-sources, forinstance, when cells at the bottom of the biolm becomeoxygen-limited or when the PAH-source they grow on isdepleted. It is furthermore known that microorganismsalso produce surface-active molecules for other purposes,e.g., as antibiotics (Nielsen et al., 2000; Nielsen et al.,2002).

5.6. Production of extracellular polymericsubstances (EPS)

The bacterial access to PAHs in the environment may

77l Pollution 133 (2005) 71e84be highly variable. This diculty could possibly be

ntamitigated by accumulation of PAHs in times of amplesupply. Organisms capable of extracting substrate pulseswould have a selective advantage. This has been shownfor diclofop methyl (methyl 2-[4-(2,4-dichlorophenoxy)phenoxy]pyruvate, a two-ring chlorinated herbicide)which accumulated in biolms by sorption to microbialexopolymers (Wolfaardt et al., 1994). The accumulateddichlofop was metabolized by the biolm communityduring starvation (Wolfaardt, 1995). The biolm regionsresponsible for diclofop accumulation also bound thehydrophobic, uorescent dye Nile Red, demonstratingthe presence of hydrophobic binding sites (Wolfaardtand Lawrence, 1998). In another study Spath andWuertz, 1998 found that biolm extracellular polymericsubstances (EPS) contained about 60e70% of accumu-lated BTX (benzene, toluene, and xylenes); whereas, theBTX content in the cells was below 20%. Sorption oflipophilic PAHs to the highly hydrated and seeminglyhydrophilic EPS is counterintuitive, but may beexplained by hydrophobic attractions between hydro-phobic and hydrophilic moieties in water (van-Oss,1995). Sugar monomers of polysaccharides can behydrophobic or hydrophilic, depending on their degreeof hydroxylation (Neu and Poralla, 1990) and theirthree-dimensional conformation (van-Oss, 1995). Sorp-tion of PAHs to EPS has indeed been reported. Twenty-four out of 28 microbial polymers tested acted assorbents for phenanthrene and facilitated the transportof phenanthrene in sand columns (Dohse and Lion,1994). The emulsion-stabilizing factor Alasan, excretedby Acinetobacter radiodurans, was shown to increase theaqueous solubility of PAHs (Barkay et al., 1999). Wespeculate that EPS may also be involved in PAH-storagein biolms. This could, for instance, be of advantagein soil where rain or bioturbation could supply micro-organisms with pulses of water-dissolved PAHs. PAH-degrading biolm-organisms could then extract water-solubilized PAHs by sorption to EPS or to other cellsurface constituents. In such a manner, acquired PAHscould help to survive periods of famine. A hint pointingat this is the nding that members of the genusSphingomonas produce sphingans, a group of exo-polysaccharides with a common repeating unit (Pollock,1993), which sorb PAHs in a way that they remainreadily bioavailable (Johnsen and Karlson, 2004).Although the binding of PAH did not aect the PAHtransfer rates, it may result in better competitivenessof sphingan-producing organisms. It is, however,unclear whether sorption of PAHs to EPS materialsserves to increase the PAH bioavailability rather thanbeing a non-specic side eect of EPS production forother purposes. In this respect, it is known that EPSare involved in anchoring biolms to hydro-phobic surfaces. EPS was for instance found to be indirect contact with anthracene crystals in Mycobacteri-

78 A.R. Johnsen et al. / Environmeum sp. LB501T biolms (Wick et al., 2001a) andwith phenanthrene crystals in Pseudomonas biolms(Rodrigues et al., 2003).

6. PAH-degrading bacteria are well-adaptedto oligotrophic conditions prevailing in soil

It has been observed that PAH degradation in soil isdominated by bacterial strains belonging to a verylimited number of taxonomic groups such as Sphingo-monas, Burkholderia, Pseudomonas and Mycobacterium(Kastner et al., 1994; Mueller et al., 1997; Ho et al.,2000; Bastiaens et al., 2000; Johnsen et al., 2002).Among these taxonomic groups a high proportion of thePAH-degrading isolates belong to the sphingomonadssensu lato (Mueller et al., 1997; Ho et al., 2000;Bastiaens et al., 2000; Johnsen et al., 2002), whichrecently have been split into the genera Sphingomonas,Sphingobium, Novosphingobium and Sphingopyxis(Takeuchi et al., 2001). Members of these genera appearto be specialized in the degradation of aromaticchemicals (e.g. Romine et al., 1999; Wattiau, 2002).Indeed, a high percentage of environmental isolates withthe capability to degrade a variety of environmentallyhazardous compounds, including PAHs (Fredericksonet al., 1995; Mueller et al., 1997; Ho et al., 2000), dioxincompounds (Halden et al., 1999) and chlorinatedphenols (Nohynek et al., 1995; Edere et al., 1997;Copley, 2000), belong to the former genus Sphingomo-nas. Sphingomonads are also widespread in unpollutedenvironments such as water distribution systems (Kos-kinen et al., 2000) and marine environments. In autumnsamples from the northern Baltic Sea, for example, onefourth of the bacterial cells were sphingomonads(Pinhassi et al., 1997). Estuarine and marine sphingo-monads are adapted to their oligotrophic habitat byhaving high-anity uptake systems and an ability tosimultaneously take up mixed substrates (Pinhassi andHagstrom, 2000; Schut et al., 1993; Eguchi et al., 1996;Schut et al., 1997). Some are even considered obligateoligotrophs (Cavicchioli et al., 1999; Schut et al., 1997).It seems possible that PAH-degrading sphingomonadsin soil are in fact adapted to the oligotrophic environ-ment rather than being particular specialists for thedegradation of hydrophobic aromatic compounds. Inthis regard, it is interesting to note that the closestrelative to a numerically dominant phenanthrene de-grader in a PAH-polluted soil was Sphingomonas sp.strain Bal5 from the Northern Baltic Sea (Pinhassi et al.,1997; Johnsen et al., 2002).

Besides being oligotrophic, a bacterium also needsthe appropriate catabolic genes to be a good PAHdegrader. Genes for the degradation of PAHs are oftenlocated on plasmids. Examples are the genes of theupper and lower pathways of naphthalene of several

l Pollution 133 (2005) 71e84pseudomonads which are borne on iso-functional

ntaNAH-plasmids (Yen and Serdar, 1988). In some cases,the enzymes encoded by NAH-plasmids have broadspecicities allowing the host to grow on several two-and three ring PAHs as the sole sources of carbonand energy (Foght and Westlake, 1996). A set ofnaphthalene degradation genes on the Pseudomonasplasmid NAH7 is part of a defective transposon, whichbecomes mobile in the presence of a Tn4653 transposase(Tsuda and Iino, 1990). In a Pseudomonas sp., thechromosomal gene-clusters encoding the upper- andlower naphthalene pathways were anked by genes withtransposase-like sequences, suggesting that the host hadacquired the naphthalene-degrading ability by trans-position (Bosch et al., 1999; Bosch et al., 2000). Also,a 184 kb conjugative, catabolic plasmid from Sphingo-monas aromaticivorans has been sequenced (Romineet al., 1999). Seventy-nine out of the 186 open readingframes were predicted to encode enzymes associatedwith the complete catabolic pathways of biphenyl,naphthalene, m-xylene and p-cresol. Other open readingframes were predicted to encode functions associatedwith DNA-recombination (invertase-, integrase- andtransposase-like sequences), replication and transfer ofthe plasmid. The presence of PAH-degradative genes onmobile, genetic elements may indicate the easy spreadingof PAH-catabolic abilities among bacteria in pollutedsoil as a result of conjugative gene-transfer. For a tar-contaminated soil, strong evidence is provided thattransfer of plasmid-encoded NAH-genes has indeedoccurred between phylogenetically dierent membersof the bacterial community (Herrick et al., 1997; Stuart-Keil et al., 1998).

In conclusion, it is conceivable that an ecient PAH-degrader is characterized by its intrinsic oligotrophy,a natural tendency to form biolms on surfaces inoligotrophic environments and the ability to acquire andexpress degradative genes.

7. Bacterial-eukaryotic consortia

It is tempting to take for real that PAH degradationin soil proceeds as in laboratory assays where a purebacterial culture or dened mixed culture mineralizesa PAH to CO2 and water without much excretion ofintermediates or interference of other substrates or otherorganisms. In a natural setting, however, various co-metabolic side-reactions will act on the PAHs and bringabout a multitude of metabolites. These metabolitesgenerally possess higher polarity than the mothercompounds and one can expect that they, at least partlyenter the pool of dissolved organic carbon. PAH-metabolites have also been found covalently linked tomacromolecular soil organic matter (Richnow et al.,1997) and only the turnover of the organic matter would

A.R. Johnsen et al. / Environmelead to the mineralization of the PAH carbon.Numerous ligninolytic and non-ligninolytic fungipossess the ability to oxidize PAHs (Cerniglia, 1992).Since fungal mycelia constitute a large fraction of thesoil biomass, they may contribute considerably to thetransformation of PAH-molecules in soil. Lignin con-tains a variety of aromatic structures formed in plantcell walls by oxygen-radical coupling-reactions of4-hydroxy cinnamyl, 3-methoxy-4-hydroxy cinnamyland 3,5-dimethoxo-4-hydroxy cinnamyl, resulting ina variety of intermonomer linkages (Hammel, 1992).Ligninolytic fungi oxidize lignin extracellularly by theaction of lignin peroxidases, Mn-dependent peroxidasesand laccases. These enzymes are unspecic and oxidizea wide variety of organic compounds. The lack ofselectivity is reasonable, considering the random struc-ture of lignin (Hammel, 1992). The peroxidases requirethe presence of peroxides (e.g. hydrogen peroxide) foractivity (Bollag, 1992). Electrons are removed fromlignin aromatic rings to give cationic radicals, which arestabilized by cleavage of CeC bonds in the ligninskeleton. The radicals then react with O2 or water togive hydroxy- or keto-derivatives (Bollag et al., 1998).The products of the peroxidase-catalyzed PAH-oxida-tions are PAH-quinones (Hammel, 1992). Laccases usemolecular oxygen to oxidize phenolic compounds tovery reactive, free radicals (Bollag, 1992). The presenceof primary, mediating substrates extend the substraterange of laccases to non-phenolic aromatics by formingpotent radicals which co-oxidize non-phenolic lignincompounds (Bourbonnais and Paice, 1990) and alsoPAHs (Pickard et al., 1999). PAHs oxidized by fungirange in size from naphthalene to benzo(a)pyrene(Cerniglia, 1992) and a variety of white rot fungalgenera are capable of oxidizing these PAHs (Cerniglia,1984; Field et al., 1992). For instance, laccase puriedfrom Coriolopsis gallica was shown to oxidize thestructurally dierent PAHs acenaphthene, phenan-threne, biphenylene, anthracene, 2-methylanthracene,9-methylanthracene and benzo(a)pyrene (Pickard et al.,1999).

The initial attack on HMW-PAHs in soil by fungalexoenzymes appears to be more likely than attack bybacterial intracellular enzymes. Fungal exoenzymeshave the advantage that they may diuse to the highlyimmobile HMW-PAHs. This is in contrast to bacterialPAH-dioxygenases, which are generally cell-boundbecause they require NADH as a co-factor. Oxidationproducts of PAHs are more soluble than the parentcompounds and therefore more bioavailable to themicrobial community as the example of the ligninolyticwhite rot fungus Bjerkandera sp. strain BOS55 shows(Kotterman et al., 1998). In pure-culture, 14C-benzo(a)-pyrene was mobilized to a great extent (74%) by BOS55,but only limited amounts of 14CO2 were produced bythe fungus. Addition of soil, sludge or LMW-PAHenrichment cultures led to a rapid increase in 14CO

79l Pollution 133 (2005) 71e842

entaproduction, indicating that some polar 14C-benzo(a)py-rene fungal metabolites were readily biodegraded,whereas others persisted.

One way to increase the extent of PAH bioremedi-ation in soil, is to mix PAH-contaminated soil withorganic matter containing mycelia of white rot fungi(Lestan and Lamar, 1996). Large quantities of myceliaare available from the commercial production of fruitingbodies of several species of white rot fungi forconsumption. The strategy has been investigatedby mixing Pleurotus ostreatus refuse ( producingMn-peroxidases and laccases) with creosote-contami-nated soil and sh oil in microcosms under optimalconditions in the laboratory (Eggen, 1999). Theseexperiments showed high degradation of 3-ring PAHs(89%), 4-ring PAHs (87%) and 5-ring PAHs (48%)within seven weeks. Mesocosm experiments withP. ostreatus refuse under eld conditions, however, haveshown little or no reductions (Hestbjerg et al., 2003).One should be aware of possible unwanted eects of theunspecic oxidation. Some metabolites may be dead-endproducts which may leak into groundwater (Kottermanet al., 1998).

Other fungi attack PAHs in the course of intracellulardetoxication processes. During detoxication, PAHsare oxidized to epoxides by cytochrome P-450 mono-oxygenase (Cyt P-450). The epoxides are then non-enzymatically rearranged to phenols, which can beconjugated (Cerniglia, 1984). Alternatively, the epoxidesare converted to trans-dihydrodiols by epoxide hydro-lases (Cerniglia, 1984). The best studied non-ligninolyticfungus Cunninghamella elegans oxidizes numerousPAHs to phenols by oxidation at Cyt P-450 (Cerniglia,1992). The produced epoxides are hydrolyzed to trans-dihydrodiols or to phenols which are subsequentlyconjugated with sulphate, glucoronic acid or glucose.

Also, the fungal Cyt P-450 transformation of PAHsmay mobilize PAHs for further degradation by bacteria.Fungalebacterial co-cultures containing the non-ligni-nolytic fungus Penicillium janthinelum VUO 10,201 havebeen shown to degrade chrysene, benzo(a)anthracene,dibenzo(a,h)anthracene and benzo(a)pyrene (Boonchanet al., 2000). Only low amounts of the PAHs weredegraded when incubated with either the fungus or thebacteria. Inoculation of soil microcosms with fungalebacterial co-cultures have resulted in improved degra-dation of HMW-PAHs and reduced carcinogenicity ofthe soils (Boonchan et al., 2000).

The ubiquitous co-existence of bacteria and fungi insoil and their known catabolic cooperation suggest thatphysical interactions between them may be of impor-tance for PAH degradation. Recently, the formation ofbacterial biolms around hyphae of mycorrhizal fungihas been demonstrated (Nurmiaho-Lassila et al., 1997;Biancotto et al., 1996; Perotto and Bonfante, 1997). It

80 A.R. Johnsen et al. / Environmmay thus be hypothesized that fungal hyphae may act asvectors to mobilize bacteria upon fungal growth andthat the creation of voids and the provision ofcontinuous surfaces by fungal hyphae could facilitatethe displacement of bacteria in soil.

Multicellular animals such as nematodes, micro-arthropodes (e.g. springtails and mites) and annelids(earthworms) are abundant in soil. By moving around inthe soil, these animals come in close contact with PAHsand may take up PAHs through the body surfaces.More important may be their PAH-uptake throughfeeding, since PAHs sorb to the soil organic detritus,which the animals feed on. Detoxication reactions arewidespread among animals and known, e.g. in isopods(Stroomberg et al., 1999), star sh (DenBesten et al.,1992), lobsters (Li and James, 1993) and polychaetes(Driscoll and McElroy, 1996). A well-studied example isthe formation of pyrene metabolites in the hepatopan-creas and gut of the common soil isopod Porcellio scaber(Stroomberg et al., 1999). Pyrene was oxidized to 1-hy-droxypyrene, 1-hydroxypyrene sulfate, 1-hydroxypyreneglycoside and three unidentied 1-hydroxypyrene con-jugates. This was demonstrated for animals exposed topyrene in the lab as well as eld-exposed animals. Itseems likely that PAHs ingested with the food will betaken up by the soil animals, Cyt P-450-detoxied byhydroxylation and conjugation and excreted to thesurroundings where it would be available to bacteria. Itmay be argued that heavily PAH-polluted soils are tootoxic for PAH-mobilizing animals to live in, but this iscertainly not the case for the diuse pollution in citycenters and along roads. It may therefore be speculatedthat many bacteria involved in PAH degradation in soillive on eukaryotic PAHmetabolites rather than onPAHs.

8. Anaerobic PAH-degradation

It has been suggested that biodegradation of PAHs,both eukaryotic and prokaryotic, require the presence ofmolecular oxygen to initiate the enzymatic attack on thePAH molecules (Cerniglia, 1992). This would have wideimplications for PAH-contamination of anaerobic sedi-ments, water-logged soils and aquifers since contamina-tion would practically last forever. Fortunately, there isincreasing evidence of anaerobic PAH-degradation withnitrate and sulfate as terminal electron acceptors. It hasrecently been demonstrated with pure cultures, thatanaerobic growth on naphthalene (as the sole source ofcarbon and energy) is possible when coupled todissimilative nitrate reduction (Rockne et al., 2000).Two naphthalene-utilizing species were isolated frommarine sediment, a denitrifying Pseudomonas sp. anda Vibrio sp. which reduced nitrate to nitrite (Rockneet al., 2000). Anaerobic PAH-degradation is not re-stricted to two-ring PAHs. A denitrifying enrichment

l Pollution 133 (2005) 71e84culture from a uidized bed reactor showed nearly

ntacomplete mineralization of phenanthrene (96%)(Rockne and Strand, 2001). Only small amounts ofphenanthrene-derived carbon was incorporated intobiomass, as opposed to 57% of the added naphthalene(Rockne and Strand, 2001). In addition, pseudomonadsisolated from both contaminated and uncontaminatedsoils were shown to degrade phenanthrene, anthraceneand pyrene under denitrifying conditions (McNallyet al., 1998). PAH-degradation has also been demon-strated with sulfate as the terminal electron acceptor.Examination of the anaerobic biodegradation potentialof river sediments showed that all of 15 PAHs testedwere degraded to some extent (8e79%) and that PAH-degradation was probably linked to sulfate reduction(Johnson and Ghosh, 1998). In a marine, sulfate-reducing sediment, uorene, phenanthrene and uoran-thene were mineralized, while pyrene and benzo(a)pyrene were not (Coates et al., 1997). Sulfate-reducingenrichment cultures from an anaerobic, coal-tar-con-taminated, sulfate-rich, aquifer grew with naphthaleneas the sole source of carbon and energy (Bedessem et al.,1997). Addition of molybdate to the cultures, a specicinhibitor of sulfate reduction, inhibited the mineraliza-tion of naphthalene, demonstrating the link betweennaphthalene mineralization and sulfate reduction(Bedessem et al., 1997). Together, these studies suggestthat the potential for PAH-degradation in anaerobicenvironments may be greater than previously recognized.

9. Implications for bioaugmentation of contaminatedsoil with PAH-degrading bacteria

Various studies have investigated the possibility ofbioaugmentation of PAH-polluted soils with PAH-degrading consortia or pure strains, and enhancedPAH-degradation in soil slurries and soil-microcosmshas indeed often been observed (Grosser et al., 1991;Madsen and Kristensen, 1997; Kastner et al., 1998).However, to our knowledge, enhanced PAH-biodegra-dation in large scale experiments as a result of in-oculation with PAH-degrading lab-strains has neverbeen demonstrated. It may be speculated that manyfailures of bioaugmentation in the early years of thistechnique may have been due to a lack of attention tothe principles of PAH degradation. This negligencecertainly also includes the focus on the intrinsicbiochemical capacity of the inocula without consider-ation of their physicochemical characteristics and theircapacity to colonize soil. Costerton and Lappin-Scott(1995) state: The planktonic single species laboratoryculture exerts a powerful selective pressure on a bacterialgenome that eventually produces a stripped down celllacking in protective and adhesive surface structures thatsimply cannot survive in natural environments where

A.R. Johnsen et al. / Environmeadhesion and protection is of paramount importance.Bioaugmentation should be more than only theaddition of a metabolic function. It may inuence thebioavailability of pollutants when the applicationmethods involve homogenization, slurrying, or intensiveushing of the system, or when the bacteria added dierfrom the indigenous population with respect to theirspecic anity for the contaminant, maintenancerequirements, ability to co-utilize natural substrates,active or passive mobility, adhesion behavior, or abilityto produce biosurfactants and to ingest surfactant-solubilized chemicals. The bioavailability may also beaected when genetic information responsible fordegradation activity of the introduced bacteria is trans-ferred to indigenous recipient bacteria, which deviatewith respect to above characteristics.

Acknowledgements

We are grateful to Ulrich Karlson and ParmelyPritchard for their critical revision of the manuscript,and to all the partners of the BIOVAB/BIOSTIMULconsortia. This work was supported by the EuropeanCommission (contracts BIO4-CT97-2015 and QLRT-1999-00326) and the Danish Strategic EnvironmentalResearch Program (BIOPRO).

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Principles of microbial PAH-degradation in soilIntroductionGrowth on PAHs as sole carbon sourcesGrowth on PAH in soilPAH-metabolismBacterial adaptations that maximize the acquisition of sorbed and separate phase PAHsDiffusionBacterial optimization of the diffusion coefficient (D) and the area (A)Optimization of the concentration at the sink (Cx)Optimization of the diffusion path length (x)Release of biosurfactantsProduction of extracellular polymeric substances (EPS)

PAH-degrading bacteria are well-adapted to oligotrophic conditions prevailing in soilBacterial-eukaryotic consortiaAnaerobic PAH-degradationImplications for bioaugmentation of contaminated soil with PAH-degrading bacteriaAcknowledgementsReferences