Polycyclic aromatic hydrocarbons (PAHs) are major soil pollutants in many industrialized countries. Microbial degradation is considered to be the major route.
FACULTEIT LANDBOUWKUNDIGE EN TOEGEPASTE BIOLOGISCHE WETENSCHAPPEN
Academiejaar 2003 - 2004
PAH-BIODEGRADATION BY SPHINGOMONAS AND MYCOBACTERIUM : STUDY OF THEIR NATURAL ABUNDANCE, DIVERSITY AND NUTRIENT DEMANDS IN PAH-CONTAMINATED SOILS. PAK-BIODEGRADATIE DOOR SPHINGOMONAS EN MYCOBACTERIUM : STUDIE VAN HUN NATUURLIJKE VERSPREIDING, DIVERSITEIT EN NUTRIËNT EISEN IN PAK-GECONTAMINEERDE BODEMS. door ir. Natalie Leys
Thesis submitted in fulfillment of the requirements for the degree of Doctor (Ph.D.) in Applied Biological Sciences Proefschrift voorgedragen tot het bekomen van de graad van Doctor in de Toegepaste Biologische Wetenschappen op gezag van Rector: prof. dr. apr. A. DE LEENHEER
Decaan: prof. dr. ir. H. VAN LANGENHOVE
Promotoren: prof. dr. ir. W. VERSTRAETE prof. dr. ir. E. TOP dr. ir. D. SPRINGAEL dr. ir. L. BASTIAENS
ISBN 90-5989-016-7
Auteur en promotoren geven de toelating dit doctoraatswerk voor consultatie beschikbaar te stellen en delen ervan te kopiëren voor persoonlijk gebruik. Elk ander gebruik valt onder de beperkingen van het auteursrecht, in het bijzonder met betrekking tot de verplichting uitdrukkelijk de bron te vermelden bij het aanhalen van de resultaten van dit werk. The author and the promoters give the authorization to consult and to copy parts of this work for personal use only. Every other use is subjected to the copyright laws. Permission to reproduce any material contained in this work should be obtained from the author.
Gent, 28 maart 2004
De auteur: ir. Natalie Leys
De promotoren: Prof. dr. ir. Willy Verstraete, Dr. ir. Eva Top, Dr. ir. Dirk Springael, Dr. ir. Leen Bastiaens
" Research is what I'm doing when I don't know what I'm doing." - Wernher von Braun –
Dankwoord
~ THE END ~ Cast The Doctor…................................................................... Natalie Leys Partners in crime…………..……………………….….........Karolien & Zita Co-stars…....................................Cindy, Barbara, David, Jan & Joke Colleagues.....................................Annemie Ryngaert & the Vito crew Assistants................................................Students Els, Tine & Carlos Supporters...................................................................All my friends!
Boyfriend.….............................................................Joachim De Baer Anti-stress ball……........................................................Yazoo the cat Mom & Dad.…........................................Erna Van Hool & Louis Leys Sister & her family.…..................Isabella Leys & Ivan Eulaers & Cato The in-laws..........................................Hilda Ringoot & Hugo De Baer Catherine De Baer & Jeroen Stuur
Dankwoord
Review comity…..............Prof. H. Van Langenhove, Prof. P. Sorgeloos, Prof. E. Vandamme, Dr. K. Smalla, Prof. M. Höfte, Prof. P. De Vos Producer.............................................................Prof. Willy Verstaete Assistant Producer.........................................................Prof. Eva Top Director..................................................................Dr. Dirk Springael Assistant Director.................................................Dr. Leen Bastiaens Logistics supervisor......................................................Dr. Ludo Diels Production support.……….……………………………………The SCK crew
~ Thank you all very much !!!
~ That's all folks!
Summary
SUMMARY Polycyclic aromatic hydrocarbons (PAHs) are major soil pollutants in many industrialized countries. Microbial degradation is considered to be the major route through which PAHs are removed from contaminated environments and therefore bioremediation is considered as a feasible remediation technology for cleaning PAHcontaminated soil. Mycobacterium and Sphingomonas strains using polycyclic aromatic hydrocarbons (PAHs) as sole source of carbon and energy could be essential members of such PAH-degrading bacterial communities, as they are often isolated during enrichments of PAH-degrading bacteria from such soil. Therefore, for future optimization of bioremediation process, it is of interest to study more in detail the distribution and diversity and specific nutrient requirements of Mycobacterium and Sphingomonas in PAH-polluted soil. Four new culture-independent PCR-based detection methods targeting the 16S rRNA genes were developed to analyze PAH-degrading Mycobacterium and Sphingomonas communities in PAH-contaminated soils. Genus-specific primers were developed for PCR detection of either Sphingomonas species (Sphingo108f and Sphingo420r), or ‘fast-growing’ Mycobacterium species (Myco66f and Myco600r). The resulting amplicons were separated by Denaturing Gradient Gel Electrophoresis (DGGE) for generating Mycobacterium and Sphingomonas community fingerprints. Both Mycobacterium and Sphingomonas specific primer sets proved to be highly selective for the target group and single-band DGGE profiles were obtained for most strains tested. Strains belonging to the same species had identical DGGE fingerprints, and in most cases but not all, these fingerprints were typical for one species, allowing partial differentiation between species in a Mycobacterium or Sphingomonas population. Inoculated Sphingomonas and Mycobacterium strains could be detected at a cell concentration of 104 respectively 106 cells per gram of soil using the new primer set alone or 102 cells per gram of soil in a nested PCR approach in combination with eubacterial primers. In addition, 2 species specific primer sets were designed to detect bacteria related to Sphingomonas sp. EPA505 (EPAf and EPAr) and M. frederiksbergense (MYCOFf and MYCOFr). Using DNA extracts of a variety of inoculated PAH-contaminated soils, the EPA505 specific primer pair was able to
Summary
detect EPA505 in concentrations as low as 102 cells per gram of soil. The MYCOF primer set could detect M. frederiksbergense in soil at a cell concentration of 104 cells per g soil via direct PCR and subsequent DNA-DNA hybridization of the PCR products or at a cell concentration of 102 cells per g soil via a nested PCR approach. The new detection methods were used to rapidly asses the Mycobacterium and Sphingomonas population structure of several PAH-contaminated soils of diverse origin and different overall contamination profiles, pollution concentrations and chemical-physical soil characteristics. Using the Mycobacterium genus-specific detection method, fast-growing Mycobacterium species were detected in most uncontaminated soils and PAH-contaminated soils tested. By sequencing of cloned PCR products amplified from DNA from PAH-contaminated soil, well-known PAHdegrading species like M. frederiksbergense and M. austroafricanum were detected. However, in all PAH-contaminated soils bacteria were detected with 16S rRNA gene sequences related to the 16S rRNA gene of M. tusciae, a Mycobacterium species so far not reported in relation to biodegradation of PAHs. Using the species-specific detection method, M. frederiksbergense strains were detected in most PAHcontaminated soils, including soils in which no M. frederiksbergense strains were detected using the Mycobacterium genus-specific detection method. The new Sphingomonas specific PCR-DGGE method revealed the presence of Sphingomonas communities in all tested PAH-contaminated soils, with less diversity in soils containing highest phenanthrene concentrations. Sequence analysis of cloned PCR products revealed new 16S rRNA gene Sphingomonas sequences significantly different from sequences from known cultivated isolates. Sequences from environmental clones grouped phylogenetically with other environmental clone sequences available in data bases and possibly originated from several potential new species, not previously detected with culture-dependent detection techniques. In most of the tested PAH-contaminated soils, we detected also 16S rRNA gene fragments from Sphingomonas sp. EPA505 related strains. By adding different inorganic supplements of nitrogen (N) and phosphorus (P) affecting the overall Carbon/Nitrogen/Phosphorus-ratio of soil, we investigated the impact of soil inorganic N and P nutrient conditions on PAH degradation by PAHdegrading Sphingomonas and Mycobacterium strains by means of soil slurry
Summary
degradation tests. The general theoretical calculated C/N/P-ratio of 120/14/3 [expressed in mg] allowed rapid PAH metabolisation by Sphingomonas and Mycobacterium strains without limitation. In addition, PAH-degradation activity was not affected when circa 10 times lower concentrations of nitrogen and phosphorus were available, indicating that Sphingomonas and Mycobacterium strains are capable of metabolizing PAHs under low nutrient conditions. In addition, PAH-degradation was not affected by an excess of nitrogen and/or phosphorus unbalancing the C/N/P ratio in the soil. Supplements of nitrogen and phosphorus salts increased however the salinity of the soil slurry solutions and seriously limited or even completely blocked biodegradation. The results presented in this thesis suggest an important role for Mycobacterium and Sphingomonas species in the PAH-degrading bacterial communities naturally colonizing PAH-contaminated soils with very different contamination profiles and different origin. Sphingomonas populations seem to dominate in soils contaminated with high concentrations of more bioavailable and more easily degradable PAHs such as phenanthrene, while Mycobacterium populations may be better adapted to flourish in soils enriched in less bioavailable higher molecular weight PAHs. In addition, the results and conclusions of the 4 year research presented in this thesis will enable us to improve bioremediation of PAH-contaminated soils by stimulating as efficiently as possible the key biodegrading organisms.
“Truth is ever to be found in the simplicity, and not in the multiplicity and confusion of things.” - Sir Isaac Newton (1642-1727) -
Samenvatting
SAMENVATTING Polycyclische aromatische koolwaterstoffen (PAK’s) zijn belangrijke chemicaliën die voorkomen in verontreinigde bodems in vele geïndustrialiseerde landen. Biologische sanering door middel van bacteriën, is een milieuvriendelijke technologie voor de zuivering van gronden vervuild met PAK’s. Microbiële afbraak is het belangrijkste proces dat zorgt voor de natuurlijke verwijdering van PAK’s in het milieu. Bacteriën van het Mycobacterium genus en het Sphingomonas genus maken mogelijk een essentieel onderdeel uit van de bacteriële gemeenschap die zorgt voor PAK-afbraak in de bodem. Bodem isolaten die in staat zijn om PAK’s te gebruiken als enige bron van koolstof en energie zijn in het verleden immers herhaaldelijke geïdentificeerd als Mycobacterium of Sphingomonas. Voor de optimalisatie van biologische PAKafbraakprocessen, is het dan ook van belang om specifiek de verspreiding, diversiteit en specifieke voedingspatronen van deze groep van PAK-afbrekende Mycobacterium en Sphingomonas stammen verder in detail te bestuderen. Vier nieuwe cultuuronafhankelijke detectie methoden werden ontwikkeld voor de analyse van PAK-afbrekende Mycobacterium en Sphingomonas gemeenschappen in gecontamineerde bodems. Twee sets van genus-specifieke primers homoloog aan het 16S rRNA gen werden ontwikkeld en gebruikt in een PCR-DGGE methode voor de simultane detectie van alle species van het Sphingomonas genus of van alle ‘snelgroeiende’ species van het Mycobacterium genus. Stammen die behoren tot hetzelfde species toonden werden gekenmerkt door identieke DGGE-profielen, en meestal was één bepaald DGGE-profiel ook kenmerkend voor één bepaald species. Deze PCR-DGGE techniek liet toe de verschillende Mycobacterium of Sphingomonas species in een natuurlijke gemeenschap te ontwarren. Met behulp van de nieuwe genus-specifieke primer sets kon men in een enkelvoudige PCR een minimale concentratie van 104 Sphingomonas cellen respectievelijk 106 Mycobacterium cellen per gram bodem detecteren, of circa 102 cellen per gram bodem via een ‘nested PCR’. Twee species-specifieke primer sets werden ontwikkeld voor de selectieve detectie van bacteriën verwant met Sphingomonas sp. stam EPA505 en M. frederiksbergense, twee species gespecialiseerd in PAK-afbraak. Met behulp van de nieuwe speciesspecifieke primer sets kon men in een enkelvoudige PCR een minimale concentratie
Samenvatting
van 102 Sphingomonas sp. EPA505 cellen respectievelijk 104 M. frederiksbergense cellen per gram bodem detecteren, of circa 102 cellen per gram bodem via een ‘nested PCR’. De nieuwe detectie methoden werden toegepast om snel de Mycobacterium en Sphingomonas populatie te karakteriseren van PAK-gecontamineerde bodems van diverse oorsprong en met verschillende contaminatieprofielen. In het merendeel van ongecontamineerde
en
PAK-gecontamineerde
bodems
‘snelgroeiende’
Mycobacterium species gedetecteerd. In sommige PAK-gecontamineerde bodems werden stammen geïdentificeerd die sterk verwant waren aan wel gekende PAKafbrekende species zoals M. frederiksbergense en M. austroafricanum. In alle PAKgecontamineerde bodems werden 16S rRNA gen fragmenten gevonden met sterke gelijkenissen aan het 16S rRNA gen van M. tusciae, een Mycobacterium species dat tot op heden nog niet in verband werd gebracht met PAK-afbraak. De speciesspecifieke detectie methode onthulde de aanwezigheid van M. frederiksbergense stammen in bijna alle PAK-gecontamineerde bodems, zelfs in bodems waar de genusspecifieke methode geen Mycobacterium species had gedetecteerd. Daarnaast, werden ook in alle PAK-gecontamineerde bodems complexe Sphingomonas gemeenschappen gedetecteerd. In PAK-gecontamineerde bodems met de hoogste fenanthrene concentraties was de gedetecteerde Sphingomonas gemeenschap het minst gediversifieerd. Sequentie-analyse van gekloneerde PCR-fragmenten toonde een duidelijk onderscheid tussen de gedetecteerde Sphingomonas 16S rRNA genen en gekende genen van gecultiveerde Sphingomonas stammen. De gekloneerde PCRfragmenten groepeerden met ander klonen sequenties beschikbaar in de elektronische databanken en behoren mogelijk toe aan een aantal nieuwe, tot op heden niet gecultiveerde, species. In de meeste PAK-gecontamineerde bodems werden via de species-specifieke methode ook Sphingomonas gedetecteerd die sterk verwant zijn aan Sphingomonas sp. EPA505. De relatie tussen beschikbare concentraties van koolstof, stikstof en fosfor als voedingsbronnen in de bodem en het gedrag van PAK-afbrekende Sphingomonas en Mycobacterium stammen werd bestudeerd aan de hand van kleinschalige biodegradatietesten. Door toevoeging van anorganische stikstof (N) en fosfor (P) werden de concentraties en verhoudingen van koolstof/stikstof/fosfor (C/N/P-ratio)
Samenvatting
van enkele natuurlijke bodems bijgestuurd. De PAK-afbraak door Sphingomonas en Mycobacterium stammen was snel en volledig onder condities die de algemene theoretisch bepaalde optimale C/N/P-ratio gelijk aan 120/14/3 (uitgedrukt in mg) benaderde. De afbraak werd zelfs niet gelimiteerd wanneer tien keer minder stikstof en fosfor beschikbaar was dan algemeen voorgeschreven als optimaal, wat er op duidt dat Sphingomonas en Mycobacterium in staat zijn om te groeien ten koste van PAK’s in omgevingen met beperkte voedingsbronnen. Meer nog, de PAK-afbraak door Sphingomonas en Mycobacterium werd niet gehinderd door een onevenwicht van de C/N/P-ratio door een overmaat aan stikstof of fosfor. Niettemin, werd er een sterke inhibitie of zelfs stop van de afbraak waargenomen wanneer door de toevoeging van stikstof en fosfor supplementen ook de zoutconcentratie te drastisch werd verhoogd. De resultaten beschreven in deze thesis suggereren een brede verspreiding en een belangrijke rol voor Mycobacterium en Sphingomonas bacteriën in PAK-afbrekende microbiële gemeenschappen in gecontamineerde bodems. Sphingomonas populaties lijken te domineren in gecontamineerde bodems met hoge concentratie aan biobeschikbare PAK’s zoals fenantreen, terwijl Mycobacterium populaties zich beter lijken te handhaven in gecontamineerde bodems met lagere concentraties aan minder biobeschikbare hoogmoleculaire PAK’s. De resultaten en conclusies van het onderzoek beschreven in deze thesis, zal ons toelaten om in de toekomst biologische sanering van PAK-gecontamineerde bodems te verbeteren door selectief de belangrijkste organismen van de microbiële gemeenschap op te volgen en te stimuleren.
“There is no higher or lower knowledge, but one only, flowing out of experimentation.” - Leonardo da Vinci (1452-1519) -
Contents
CONTENTS
INTRODUCTION _________________________________________________________________1 PAH-BIODEGRADATION BY SPHINGOMONAS AND MYCOBACTERIUM :STUDY OF THEIR NATURAL ABUNDANCE, DIVERSITY AND NUTRIENT DEMANDS IN PAHCONTAMINATED SOILS
CHAPTER 1 ______________________________________________________________________3 BACTERIAL PAH-BIODEGRADATION AND BIOREMEDIATION OF PAH-CONTAMINATED SOILS: A LITERATURE REVIEW
CHAPTER 2 _____________________________________________________________________61 OCCURRENCE AND DIVERSITY OF FAST-GROWING MYCOBACTERIUM SPECIES IN SOILS CONTAMINATED WITH POLYCYCLIC AROMATIC HYDROCARBONS (PAHs)
CHAPTER 3 _____________________________________________________________________83 MYCOBACTERIUM FREDERIKSBERGENSE, A MYCOBACTERIUM SPECIES SPECIALISED IN POLYCYCLIC AROMATIC HYDROCARBON (PAH) DEGRADATION, IS UBIQUITOUS IN PAH-CONTAMINATED SOILS
CHAPTER 4 _____________________________________________________________________97 OCCURRENCE AND PHYLOGENETIC DIVERSITY OF SPHINGOMONAS IN SOILS CONTAMINATED WITH POLYCYCLIC AROMATIC HYDROCARBONS (PAHS)
CHAPTER 5 ____________________________________________________________________ 117 OCCURRENCE OF SPHINGOMONAS SP. EPA505 RELATED STRAINS IN SOILS CONTAMINATED WITH POLYCYCLIC AROMATIC HYDROCARBONS (PAHS)
CHAPTER 6 ____________________________________________________________________ 129 INFLUENCE OF THE CARBON/NITROGEN/PHOSPHATE-RATIO ON PAH-DEGRADATION BY MYCOBACTERIUM AND SPHINGOMONAS STRAINS IN SOIL CHAPTER 7 ____________________________________________________________________ 145 GENERAL DISCUSSION AND PERSPECTIVES
BIBLIOGRAPHY ________________________________________________________________ 153
CURRICULUM VITAE ___________________________________________________________ 181
“The task is, not so much to see what no one has yet seen; but to think what nobody has yet thought, about that which everybody sees.” - Erwin Schrödinger (1887-1961) -
Introduction
INTRODUCTION PAH-BIODEGRADATION BY SPHINGOMONAS AND MYCOBACTERIUM : STUDY OF THEIR NATURAL ABUNDANCE, DIVERSITY AND NUTRIENT DEMANDS IN PAH-CONTAMINATED SOILS Polycyclic Aromatic Hydrocarbons (PAHs) are common pollutants of air, water and soil in many industrialized countries. Although PAHs are naturally present at low concentrations in the terrestrial environment, pollutions are mainly due to human activities. High concentrations of PAHs occur in contaminated soils at wood treating facilities, at sites formerly used to produce manufactured gas, petroleum processing plants and in river and harbor sludge. PAH-contamination is of environmental and public concern due to their toxic, mutagenic and carcinogenic properties. Microbial degradation is considered to be the major route through which PAHs are removed from contaminated environments and therefore bioremediation is considered as a feasible remediation technology for cleaning PAH-contaminated soil. Currently, in situ and ex situ bioremediation techniques are, however, still considered ineffective for the efficient removal of PAHs from contaminated soil due to general low biodegradation rates obtained. Biodegradation is mainly hampered by the low bioavailability of the hydrophobic PAHs which strongly adsorb to organic soil material or dissolve in non-aqueous phase liquids. In addition, mostly the soil is basically treated as a ‘black box’ with the inability to control and direct the biological processes. Not much is known about the key organisms involved in the degradation processes and about the specific needs of the biocatalysts with respect to nutrition and environmental conditions for their optimal activity in the soil environment. Different research data indicate that certain groups of soil bacteria are specialized in colonization of PAH-contaminated environments and may play the main role in the biodegradation process. PAH-degrading isolates almost exclusively belong to the genera Sphingomonas and Mycobacterium, which seem to possess not only the necessary enzyme machinery for degradation of PAH compounds but also seem to make use of original strategies to enhance PAH bioavailability. However, current knowledge is mostly based on cultivation–based isolated strains. Not much is know
-1-
Introduction
about the in situ occurrence, distribution and PAH-degradation activity of Sphingomonas and Mycobacterium strains and their specific nutrient requirements to become stimulated. The work presented in this thesis, aimed to elucidate the natural occurrence, diversity and nutrient demands of PAH-degradation by Mycobacterium and Sphingomonas in polluted soils. The first goal was the development of monitoring techniques to screen soils for the natural presence of bacteria belonging to the genera Sphingomonas and Mycobacterium and for monitoring their dynamics during bioremediation processes. The applicability of molecular techniques based on total soil DNA-extraction followed by specific 16S rRNA gene amplification by PCR and DGGE-analysis of the resulting amplicons was explored. The second aim was the determination of the impact of soil characteristics such as salinity, pH and nutrients available in the soil on the activity of PAH-degrading Sphingomonas spp. and Mycobacterium spp.. Different (in)organic supplements were added to the soil to change the overall C/N/P-ratio, the pH and/or the ionic strength of the soil and the degradation activity of added Mycobacterium and Sphingomonas spp. was followed. The manuscript is divided in 7 chapters. Chapter 1 provides an overview of the current literature on microbial degradation of PAHs with emphasis on (i) the ecology of PAH-contaminated environments, (ii) the phylogeny of the bacteria and the catabolic systems involved in biodegradation, and (iii) the environmental parameters influencing microbial PAH-biodegradation activity in soil during bioremediation processes. Chapter 2, Chapter 3, Chapter 4 and Chapter 5 describe the development and testing of molecular monitoring techniques based on PCR and DGGEfingerprinting to specifically detect Sphingomonas spp. and Mycobacterium spp. in soil. These new detection methods were used to study the occurrence and diversity of Sphingomonas and Mycobacterium communities in PAH-contaminated soil. In Chapter 6, we examined the different nutrition and environmental conditions to optimize the survival and activity of the PAH-degrading Sphingomonas spp. and Mycobacterium spp. in soil. Chapter 7 discusses the obtained results in the framework of the research objectives. Conclusions are drawn and perspectives for further research are presented.
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Literature Review
CHAPTER 1 BACTERIAL PAH-BIODEGRADATION AND BIOREMEDIATION OF PAH-CONTAMINATED SOILS: A LITERATURE REVIEW *†
* REDRAFTED
AFTER:
LEYS NATALIE, BASTIAENS LEEN, AND SPRINGAEL DIRK (IN
PREPARATION)
BACTERIAL BIODEGRADATION OF PAHS IN CONTAMINATED HABITATS: A LITERATURE REVIEW, CURR. ADV. APPL. MICROBIOL. BIOTECHNOL.
† REDRAFTED MICROBIAL
AFTER:
LEYS NATALIE, BASTIAENS LEEN, AND SPRINGAEL DIRK (IN
BIOREMEDIATION OF
PAH-CONTAMINATED
SOIL:
A
LITERATURE
PREPARATION)
REVIEW, CURR. ADV.
APPL. MICROBIOL. BIOTECHNOL.
INTRODUCTION Polycyclic aromatic hydrocarbons (PAHs) are a group of highly stable aromatic organic compounds, consisting of benzene rings in linear, angular and cluster arrangements containing only carbon and hydrogen (Table 1-1) (Harvey 1991). The more rings, the more hydrophobic the compounds are, the higher their octanol-water partitioning coefficients (Kow), i.e., the lower their water solubility, and the higher their melting and boiling points (Table 1-1) (Ernst 1995). PAHs are of governmental concern as they pose a risk for ecosystems and public health because of their possible cytotoxic, teratogenic, mutagenic and carcinogenic properties (Table 1-1) (Enzminger 1987; Harvey 1991). Most unsubstituted PAHs with four or fewer rings, are relatively harmless for humans, but can be toxic to marine diatoms, gastropods, mussels, crustaceans, and fish. Unsubstituted PAHs with five or six rings exhibit a wide range of carcinogenic activity and can cause cancers in mammalian animals and humans. Substituted PAHs, i.e., methylated or nitrated PAHs, are even of more of concern for public health. The pathways for human exposure to PAHs are inhalation (active/passive smoking and inhaling of polluted air), ingestion (contaminated food and drinks) and skin adsorption (Menzie 1992).
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Chapter 1
TABLE 1-1
PHYSICAL AND BIOLOGICAL CHARACTERISTICS OF SOME PAHS (Ernst; Enzminger 1987; Harvey 1991)
PAH Naphthalene Acenaphtene
Formula MW C10H8 128,17 C12H10 152,21
Melting Point (°C)
Water solubility at 30°C (ppb)
Carcinogenity
82
31.700
NR
95
3.930
NR
Fluorene
C13H10 166,22
116
1.980
NR
Phenanthrene
C14H10 178,24
101
1.290
iiST, iiiA
Anthracene
C14H10 178,24
218
73
ivST, ivA
Fluoranthene
C16H10 202,26
110
260
iiST, ivA
Benzo(a)fluorene
C17H12 216,28
190
NR
NR
Pyrene
C16H10 202,26
150
135
iST, ivA
Benzo(a)anthracene
C18H12 228,29
161
14
iiST, iiA
Chrysene
C18H12 228,28
256
2
iST, iA
Benzo(b)fluoranthene
C20H12 252,32
168
NR
iiiST, iA
Benzo(k)fluoranthene
C20H12 252,32
217
NR
iiiST, iA
Benzo(a)pyrene
C20H12 252,32
179
4
iiiST, iA
Dibenzo(ah)anthracene
C22H14 278,33
270
NR
iST, iA
Indenol(1,2,3-cd)pyrene
C22H12 276,34
164
NR
iiiST, iiA
Benzo(g,h,i)perylene
C22H12 276,34
273
0,26
iiiST, iiiA
Coronene
C24H12 300,28
442
NR
NR
i – sufficiently proven. ii – limited proven. iii – insufficiently proven. iv – not proven. A – through animal testing. ST – through short term experiment. NR – not reported.
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Literature Review
PAHs are common pollutants of air, water and soil in industrialized countries (Edwards 1983). In the Flemish part of Belgium, for example, sites contaminated with PAHs are concentrated around the city of Antwerp and Ghent and counted in 2002 3908 sites with contaminated top soil and 295 sites with contaminated top soil and ground water (OVAM 2003). PAHs present in the atmosphere, surface soil and waters may originate from many different sources (Ernst 1995). Natural forest fires and volcanic activity release occasionally PAHs in the environment but the most prominent sources of PAH-contamination are related to anthropogenic processing of fossil fuels such as crude oil and coal. PAHs are worldwide distributed in the whole ecosphere due to increasing traffic emissions. The concentration of PAHs in urban atmosphere depends on the number and types of local emissions sources, temperature, weather and seasonal variations (Dann 2001), but in general concentrations in urban and industrial areas are 10-100 times higher than in more remote regions (Harvey 1991). PAHs enter water environments directly through precipitation of contaminated dust from the air, through the runoff of polluted ground and through pollution of rivers and lakes by municipal and industrial effluents (Harvey 1991; Ernst 1995). In addition, many river and harbor sediments are severely contaminated with PAHs. Maintenance of such water transport routes yields yearly millions of tonnes of heavily polluted sludge, which needs to be treated or dumped in confined landfills. PAHcontaminated soil is often found around former gas production plants, wood treatment plants where creosote was used, incineration plants, petroleum refining plants, and asphalt production plants. Once introduced in the environment, PAHs accumulate in air, surface waters and sediments and soil and may persist for decades due to their hydrophobicity and recalcitrance. The fate of PAHs in the environment depends on many physical, chemical and biological interactions between the sorbate (PAHs), the sorbent (sediment/soil/tar), the solvent (NAPL/water) and the living organisms (Enzminger 1987; Heitzer 1993; Luthy 1997; Ramaswami et al. 1997; Gosh 2000). The relative importance of the individual processes is different for each PAH-compound and varies with the physicochemical properties of the site matrix and the environmental conditions. PAHs in the atmosphere are associated with airborne particles and are subject to various modes of chemical and photochemical degradation forming for example nitrogen and sulfur-containing heterocyclic compounds. Although the -5-
Chapter 1
concentration of PAHs in freshwater are usually low, PAHs dissolved in water can be ‘taken up’ by plants (Wild 1992), vertebrate fish and shellfish (Jackson 1994) and transferred in the food chain. Soils in particular are ‘sinks’ where PAHs tend to concentrate. The extractability and bioavailability of hydrocarbons in soils decreases with increasing soil hydrocarbon contact time, i.e., a process designated as ‘aging’ (Bauer et al. 1985; Weissenfels et al. 1992; Hatzinger et al. 1995; Macleod et al. 2000; Reid et al. 2000). The processes of hydrocarbon sequestration in soil are thought to be driven by (i) diffusion into soil micro pores, (ii) partitioning (ad- and absorption) into the soil organic matter (SOM), and (iii) accumulation in non-aqueous phase liquids (NAPL). Soils with higher organic carbon (OC) contents have a larger capacity to sequester hydrocarbons (Cornelissen et al. 1989; Weissenfels et al. 1992) and in soils with an OC content greater than 0.1%, partitioning into the SOM has been found to be the dominant sequestration process (Chiou et al. 1979; Means et al. 1980; Pignatello et al. 1996; Luthy 1997). Soil at manufactured gas plants are typically contaminated with dense NAPLs or coal tar particles, containing considerable amounts of PAHs (Efroymson 1994; Efroymson et al. 1995; Luthy 1997; Ramaswami et al. 1997). Mass transfer of PAHs between tar and the aqueous/solid phase is an extra process determining the fate of PAHs in such environment. Unlike physico-chemical sequestration or transport, biodegradation is the only natural process for actual removal of PAHs from the environment (Sims et al. 1983; Cerniglia 1984; Cerniglia 1992). Biodegradation can be done by bacteria, fungi, plants or algae or synergetic consortia between these living organisms. Microbial degradation is considered as a major route through which PAHs are naturally removed from contaminated environments (Bossert et al. 1984; Cerniglia 1984; Enzminger 1987; Cerniglia 1992). Therefore, bioremediation is considered as an economically and ecologically beneficial remediation technology for the treatment of PAH polluted sites (Erickson et al. 1993; Wilson et al. 1993; Luthy et al. 1994; Würdemann et al. 1995).
-6-
Literature Review
BACTERIAL BIODEGRADATION OF PAHS IN CONTAMINATED HABITATS Ecology of PAH-degrading microbial communities in the environment Microbial communities of contaminated soils can adapt to the contaminants and contribute to natural attenuation of pollutants such as PAHs. Different reports indicate that environmental stresses, including exposure to contaminants, lead to changes in microbial community structure through the selective enrichment of specific microorganisms that are more adapted to the new environment (Herbes et al. 1978; Spain et al. 1980; Spain et al. 1983; Wiggins et al. 1987; Bauer et al. 1988; Aeolin et al. 1989; Leahy et al. 1990; van der Meer et al. 1992; Tuhackova et al. 2001; Macleod et al. 2002). It has been shown that the community changes are controlled by the pollutant concentration, i.e., the number of specific degraders are higher and the degradation capabilities are more diverse in more contaminated soils (Spain et al. 1980; Spain et al. 1983; Wiggins et al. 1987; Aeolin et al. 1989; Leahy et al. 1990; Grosser et al. 1991; Grosser et al. 1995; Carmichael et al. 1997; Macleod et al. 2000; Tuhackova et al. 2001; Tuxen et al. 2002). Adaptation leads to reduced lag times and faster rates of PAH-mineralization and increases the extent of the degradation, i.e., a larger final proportion of PAHs was mineralized to CO2 (Bauer et al. 1988; Macleod et al. 2002). Microbial adaptation to PAHs in aquifers has been demonstrated by the mineralization of PAHs and contaminant-stimulated in situ bacterial growth (enrichment of PAH-degrading bacteria and enhanced numbers of Protozoa) in PAHcontaminated zones but not in adject uncontaminated zones (Madsen et al. 1991; Ghiorse et al. 1995). Not much is currently known about the global taxonomic composition of the bacterial communities degrading PAHs in the environment and the interaction between the different members during the degradation process. There are only a limited number of reports on culture-based or culture-independent total community analysis of PAHcontaminated niches. The composition of the microbial community in PAHcontaminated habitats has mainly been assessed by the enrichment and study of PAHdegrading isolates. -7-
Chapter 1
The identity of the isolated PAH-degrading bacteria depended more on the source environment and contamination profile than on the geographical origin of the source material. PAH-degrading bacteria have been isolated from many different anthropogenic contaminated environments (Bastiaens 1998; Bastiaens et al. 2000; Bicknell et al. 2001) as well as from undisturbed pristine environments (Heitkamp et al. 1989). PAH-degrading bacteria have been detected in biofilms on building stones (Ortega-Calvo et al. 1997), plant material (Juhasz et al. 2000a), top soil (road side soil) (Tuhackova et al. 2001; Johnsen et al. 2002), plant rhizosphere soil (Radwan et al. 1998; Daane et al. 2001), bulk soil (Song et al. 1986; Bouchez et al. 1995; Deziel et al. 1996; Dagher et al. 1997; Willumsen et al. 1997; Bastiaens 1998; Boonchan et al. 1998; Bastiaens et al. 2000; Juhasz et al. 2000a; Yuste et al. 2000; Rehmann et al. 2001; Baraniecki et al. 2002; Johnsen et al. 2002; Widada et al. 2002), subsurface soil and groundwater (Bicknell et al. 2001; Bakermans et al. 2002), fresh water and sediments rivers and lakes (Nortemann et al. 1986; Heitkamp et al. 1987; Cerniglia 1989; Dean-Ross et al. 2001), and, estuarine water and sediments from seas or oceans (Heitkamp et al. 1987; Heitkamp et al. 1988a; Cerniglia 1989; Geiselbrecht et al. 1998; Hedlund et al. 1999). Bacteria and fungi appear to be the dominant hydrocarbon degraders in soil environments while bacteria and yeast are the main degraders in aquatic ecosystems (Hanson et al. 1997). Via culture-independent analysis of 16S rRNA genes in soil-derived
13
C-labeled DNA, Pseudomonas, Acinetobacter and
Variovorax spp. were identified as the dominant active bacteria degrading 13C-labeled naphthalene in top soil (Padmanabhan et al. 2003). Oil- and PAH-utilizing rhizosphere isolates predominately belonged to the Arthrobacter genus (Radwan et al. 1998), Pseudomonas genus (Daane et al. 2001), Painibacillus genus (Daane et al. 2001) or Nocardia-Mycobacterium-Rhodococcus group (Daane et al. 2001). In moderated and heavily contaminated bulk soil, PAH-degrading isolates appeared to be mostly members of the Pseudomonas genus (Bouchez et al. 1995; Deziel et al. 1996; Dagher et al. 1997; Bastiaens 1998; Bastiaens et al. 2000), the Sphingomonas genus (Dagher et al. 1997; Mueller et al. 1997; Bastiaens 1998; Bastiaens et al. 2000; Ho et al. 2000; Baraniecki et al. 2002; Johnsen et al. 2002) and the Mycobacterium genus (Grosser et al. 1991; Bastiaens 1998; Bastiaens et al. 2000; Ho et al. 2000; Rehmann et al. 2001; Bogan et al. 2003), and in lesser extent members of the Rhodococcus -8-
Literature Review
(Walter et al. 1991; Bouchez et al. 1995; Tongpim et al. 1996) and Burkholderia (Mueller et al. 1997) genera. That Mycobacterium strains are ubiquitous in PAHcontaminated soils from very different origin was also confirmed by cultureindependent 16S rRNA gene based community analysis (Cheung et al. 2001). Fresh water sediment PAH-degrading isolates were classified as a new species of the Polaromonas genus (Jeon et al. 2004). Obligate marine PAH-degrading isolates represented often new species and new genera of the γ-Proteobacteria such as Neptunomonas (Hedlund et al. 1999) and Cycloclasticus (Geiselbrecht et al. 1998; Kasai et al. 2002; Kasai et al. 2003). In addition, a relationship can be observed between the complexity, the hydrophobicity and recalcitrance of PAH compounds and the bacteria using the compound. Most bacteria selected on 2-ring PAHs such as naphthalene belong to the Pseudomonas genus (Dunn et al. 1973; Simon et al. 1993; Kiyohara et al. 1994; Deziel et al. 1996; Bastiaens 1998; Bastiaens et al. 2000), Sphingomonas genus (Fredrickson et al. 1995; Yrjala et al. 1998; Davidson et al. 1999a; Davidson et al. 1999b) or exceptionally to the Rhodococcus genus (Larkin et al. 1999), Arthrobacter genus (Ortega-Calvo et al. 1994), Ralstonia genus (Widada et al. 2002) or Neptunomonas genus (Hedlund et al. 1999). Strains utilizing 3-ring PAHs such as phenanthrene belong mainly to the Sphingomonas genus (Mueller et al. 1990; Kästner et al. 1994; Balkwill et al. 1997; Mueller et al. 1997; Gibson 1999; Ho et al. 2000; Pinyakong et al. 2000; Cho et al. 2001; Johnsen et al. 2002; Borde et al. 2003), Pseudomonas genus (Menn et al. 1993; Bouchez et al. 1995; Mueller et al. 1997; Ortega-Calvo et al. 1997; Garcia-Junco et al. 2001a; Borde et al. 2003), Aeromonas genus (Kiyohara et al. 1976), Burkholderia genus (Mueller et al. 1997), or gram positive genera such as Nocardioides (Saito et al. 2000), Planococcus (Ortega-Calvo et al. 1997), Bacillus (Ortega-Calvo et al. 1997), and the actinomycte genera Arthrobacter (Schwartz et al. 2000), Nocardia (Ortega-Calvo et al. 1997), Rhodococcus (Walter et al. 1991; Bouchez et al. 1995; Tongpim et al. 1996; DeanRoss et al. 2001), and Mycobacterium (Kleespies et al. 1996; Mueller et al. 1997; Churchill et al. 1999; Rehmann et al. 2001; Bogan et al. 2003). For the degradation of 4-ring PAHs such as
pyrene a few gram negative strains belonging to the
Proteobacteria genera Sphingomonas (Mueller et al. 1990; Weissenfels et al. 1991; -9-
Chapter 1
Mueller et al. 1997; Kästner 1998), Burkholderia (Juhasz et al. 1997a), Pseudomonas (Boonchan et al. 1998) and Stenotrophomonas (Boonchan et al. 1998; Juhasz et al. 2000c)
have
been
reported,
but
they
mainly
include
members
of
the
Corynebacterineae suborder, i.e., Gordonia (Kästner et al. 1994), Rhodococcus (Walter et al. 1991; Bouchez et al. 1995; Dean-Ross et al. 2001) and Mycobacterium (Grosser et al. 1991; Walter et al. 1991; Boldrin et al. 1993; Bouchez et al. 1995; Dean-Ross et al. 1996; Jimenez et al. 1996; Schneider et al. 1996; Thibault et al. 1996; Ho et al. 2000; Bogan et al. 2003; Gauthier et al. 2003). The apparent genus related preference for certain PAHs is probably due to increasing PAH-hydrophobicity with increasing number of aromatic rings. Some genera or species of PAH-degrading bacteria are more capable to interact with hydrophobic surface than others as also indicated by the effect of the isolation mode on the type of isolates. Using the same soils and the same PAHs compounds, different bacteria were isolated depending on the isolation procedure (Bastiaens 1998; Tang et al. 1998; Bastiaens et al. 2000; Friedrich et al. 2000; Grosser et al. 2000; Gauthier et al. 2003). Aqueous cultures using crystalline PAHs mainly led to the isolation of gram negative bacteria such as Pseudomonas and Sphingomonas species (Bastiaens 1998; Bastiaens et al. 2000). Two-liquid phase cultures selected for Mycobacterium, Bacillus, Microbacterium and Porphyrobacter strains (Gauthier et al. 2003). Enrichments using weakly sorbed PAHs (Amberlite beads) selected exclusively for Burkholderia species (Friedrich et al. 2000) whereas enrichments using strongly sorbed PAHs (hydrophobic Teflon membranes or polyacrylic beads) selected exclusively for Mycobacterium species (Bastiaens 1998; Bastiaens et al. 2000; Friedrich et al. 2000). Bacteria isolated on sorbed PAHs were more efficient in the degradation of the sorbed compound than species isolated from enrichment on non-sorbed PAHs (Tang et al. 1998), indicating that different PAH-degrading bacteria inhabiting the same soil may be adapted to different PAH-bioavailabilities. Analysis of microbial communities in sand-packed columns run with groundwater clearly showed compositional and functional difference in free-living and surface-associated bacteria that may reflect different roles of these distinct but interacting communities in the biodegradation of pollutants in aquifers (Lehman et al. 2002). The dispersal of catabolic capacity via mobile genetic elements (MGE) and horizontal gene transfer (HGT) plays a major role in bacterial adaptation to environmental - 10 -
Literature Review
stimuli, such as exposure to organic pollutants (Herrick et al. 1997; Stuart-Keil et al. 1998; Hohnstock et al. 2000; Park et al. 2003; Top et al. 2003; Wilson et al. 2003). Catabolic genes for PAH-degradation are often localized on conjugative plasmids and transposons (Park et al. 2003; Top et al. 2003; Nojiri et al. 2004). Dissimination of naphthalene-catabolic gene via conjugative plasmid transfer has been demonstrated in PAH-contaminated aquifers (Hohnstock et al. 2000). Conjugation frequency and natural HGT of PAH-catabolic plasmids was found to be stimulated by the exposure of the host to naphthalene (Hohnstock et al. 2000) and by the presences surfaces where conditions fostering stable, high-density cell-to-cell contact are more manifest (Park et al. 2003). The distribution and diversity of the catabolic genes involved in PAH-biodegradation in the environment has only been studied very recently. Culture-independent PCR and RT-PCR techniques have indicated that naphthalene metabolism genes are ubiquiteous and actively in situ-transcribed in the indigenous community of coal-tar contaminated sites (Ghiorse et al. 1995; Wilson et al. 1999). The naphthalene metabolism genes were similar to those found on the NAH7 plasmid of P. putida strain G7 (nahAc and nahR genes) but significant sequence polymorphism of the gene and mRNA PCR-products indicated the presence of divergent homologs of the initial naphthalene dioxygenase alfpha subunit gene nahAc in the naphthalene-degrading bacterial population (Ghiorse et al. 1995; Wilson et al. 1999). Sequence comparisons revealed two major groups of nahAc homologs related to the naphthalene dioxygenase genes ndoB and dntAc, previously cloned from pDTG1 plasmid from P. putida NCIB 9816-4 and Burkholderia sp. strain DNT, repectively (Wilson et al. 1999). Cultureindependent PCR and RT-PCR techniques have revealed much greater native naphthalene catabolic gene diversity than what was ever detected by culture-based approaches (Wilson et al. 1999). The ecological significance, relative distribution and transmission modes of the different naphthalene dioxygenase analogs was studied only rarely. One study on contaminated New Zealand soils indicated that the βProteobacterium phnAc allele, only found rarely in culture-based studies, may have a greater ecological significance in PAH-degrading communities than the extensively studied γ-Proteobacterium nah-like genotype (Laurie et al. 2000). A recent publication found that phnAc-like genes were mainly present in naphthalenedegrading isolates form a moderately contaminated hillside soil while nahAc-like - 11 -
Chapter 1
genes were found only among naphthalene-degrading isolates from an adjacent more heavily contaminated seep sediment (Wilson et al. 2003). Physiological characteristics of PAH-degrading bacteria A large number of PAH-degrading microorganisms have been isolated from PAHdegrading environmental communities and characterized. Almost all PAH-degrading isolates are aerobic and able to use PAHs as sole source of carbon and energy. PAHdegrading bacteria of different genera seem to have adapted very specifically to the PAH-substrates they are using. Especially the genera Pseudomonas, Sphingomonas and Mycobacterium are well known for their degradation potential towards recalcitrant compounds including PAHs and have acquired diverse capabilities to inhabit a wide range of environments. A typical adaptation of some Pseudomonas and Rhodococcus strains is the production of biosurfactants (e.g. glycolipids, sophorolipids, trehalose lipids, rhamnolipids, phospholipids and fatty acids) and bioemulsifiers (polysaccharides) which increase the aqueous solubility of PAHs (Deziel et al. 1996; Willumsen et al. 1997; Iwabuchi et al. 2002; Carcia-Junco et al. 2003). Some strains excrete extra-cellular surfactants but for most strains the surface activity is associated with the cell envelope (Willumsen et al. 1997). Another adaptation is a strong adhesion to the hydrophobic PAH-substrate and the formation of biofilms on PAH crystals, on hydrophobic surfaces coated with sorbed PAHs and organic non-aqueous phase liquids with dissolved PAHs (Ortega-Calvo et al. 1994; Garcia-Junco et al. 2001b; Carcia-Junco et al. 2003). Biofilm formation can increase diffusion or dissolution of the PAHs. Some Pseudomonas and Sphingomonas strains enhance biofilm formation on PAH-substrates by the production of viscous extrapolysaccharides (EPS) (Pollock et al. 1999; Johnsen et al. 2000). The efficient PAH-degrading capacities of Sphingomonas and Mycobacterium strains are probably linked to these specific hydrophobic cell wall properties which control the interaction with and the membrane transport of hydrophobic compounds (Nohynek et al. 1995; Kawai 1999; Wick et al. 2001; Wick et al. 2002a). Both Sphingomonas and Mycobacterium strains have a very particular lipophilic outer cell wall layer. Glycosphingolipids replace the normal Gram negative lipopolysaccharides in the - 12 -
Literature Review
outer membrane of Sphingomonas cells (Kawahara et al. 1999; Wiese et al. 1999; Yabuuchi et al. 1999) and Mycobacterium cells have an additional layer of mycolic acids (glycolipids) on top of the common Gram positive peptidoglycan cell wall (Sayler et al. 1994). Moreover, Mycobacterium cells respond to PAHs as growth substrates by changing the mycolic acid composition of their cell wall, i.e., they become more hydrophobic and more negatively charged (Wick et al. 2001; Wick et al. 2002a; Wick et al. 2002b; Wick et al. 2003b). While a more negative potential of cells would be expected to increase repulsive electrostatic interactions between cells and solids, a more hydrophobic cell wall might allow the bacterium to adhere better to hydrophobic surfaces such as Teflon and PAH crystals to form biofilms (van Loosdrecht et al. 1990a; van Loosdrecht et al. 1990b; Bastiaens 1998; Bastiaens et al. 2000; Wick et al. 2001; Wick et al. 2002a). PAH-degrading Pseudomonas, Sphingomonas and Mycobacterium strains have generally hydrophobic and negatively charged cell envelopes (Rijnaarts et al. 1993). The Gram positive Mycobacterium cells are in general more hydrophobic and more negatively charged than the Gram negative Pseudomonas and Sphingomonas cells (Rijnaarts et al. 1993). Many PAH-degrading bacteria seem to use very specific substrate uptake mechanisms. There are strong indications that PAHs are metabolized by intracellular enzymes located in the periplasmic space or bound to membranes, indicating that transport of these compounds through outer membranes is a requisite for their metabolism (Kawai 1999). A novel pit dependant endocytosic macromolecule transport system for hydrophobic polymers has been identified in a Sphingomonas strain growing on alginate, a natural HMW hydrophobic polymer (Momma et al. 1999). Electron microscopy revealed dynamic changes in both cell surface and membrane structure and mouth-like pits which open en close depending on the presence or absence of the substrate. Similar systems may be used for the incorporation of PAHs. Also Mycobacterium species seem to be well adapted to the low available PAH concentrations as they make use of high-affinity uptake systems (Miyata et al. 2004) and can shift to maintenance metabolism during growth on poorly available sorbed PAHs (Wick et al. 2001; Wick et al. 2002a). A specific ‘interfacial’ uptake system has been suggested to enhance PAH-biodegradation by Arthrobacter and Rhodococcus cells absorbed at the interface of non-water-soluble non-degradable solvent containing PAHs (Ortega-Calvo et al. 1994). - 13 -
Chapter 1
Some PAH-degrading strains are mobile and exhibit a chemotactic response towards PAHs (Pandey et al. 2002; Samanta et al. 2002). Pseudomonas putida G7, Pseudomonas sp. NCIB9816-4 and Pseudomonas putida RKJ1 are chemotactic attracted by naphthalene (Grimm et al. 1999; Samanta et al. 2000). For the first 2 strains, the naphthalene chemoreceptor, NahY, is encoded downstream of the naphthalene catabolic genes on the NAH7 plasmid and is co-transcribed with the catabolic genes (Grimm et al. 1999). In addition, also several Sphingomonas strains display phenotypic dimorphism and can adopt either a planktonic or sessile behavior in liquid media (Kovarova et al. 1998; Pollock et al. 1999; Johnsen et al. 2000). Sensing of environmental stimuli and genetic control over synthesis of the capsule are key events in alternating between these two phenotypes. In addition, the simultaneous utilization of multiple substrates (mixed-substrate growth) has been shown to be a heterotrophic bacterial strategy under oligotrophic conditions (Kovariva et al. 1997; Kovarova et al. 1998) that might also be utilized by PAH-degrading bacteria. Simultaneous degradation of different PAHs has been reported several times (Guha et al. 1999). In ternary mixtures of PAHs the biodegradation rates of the more degradable and abundant compounds are reduced due to competitive inhibition, but enhanced biodegradation of the more recalcitrant PAHs occurs due to simultaneous biomass growth on multiple substrates (Guha et al. 1999). Some specialize PAH-degrading bacteria even utilize PAHs such as anthracene in the presences of much more easily degradable organic compounds such as glucose (Wick et al. 2003a). It could be a response to the heterogeneous composition of aromatic structures in the fossil organic matter in the habitat from which the strains were isolated (Romine et al. 1999a). Taxonomy of PAH-degrading Mycobacterium and Sphingomonas strains The genus of Mycobacterium [Lehmann and Neumann 1896] belongs to the phylum of Gram positive eubacteria with high G+C content (>55%), the class of Actinobacteridae, the order of Actinomycetales, the suborder of Corynebacterineae and the family of the Mycobacteriaceae (Tsukamura et al. 1977). The Mycobacterium
- 14 -
Literature Review
genus currently groups 100 different registered species and is represented by the type species Mycobacterium tuberculosis. The 16S and 23S rRNA gene based phylogenetic trees of the Mycobacterium genus show 2 major taxa : the ‘fast-growing Mycobacterium species’ form a coherent line of descent, distinct from the more recently evolved ‘slow-growing’ Mycobacterium species (Rogall et al. 1990; Stalh et al. 1990; Pitulle et al. 1992; Stone et al. 1995; Tortoli 2003). The ‘fast-growing Mycobacterium species’ are a group of Mycobacterium strains, mostly of environmental origin, that are, based on growth and biochemical characteristics and infectious properties (i.e. Mycobacterium species of Bio Safety Level 1, growth within 7 days), very different from the pathogenic and facultative pathogenic more slowly growing species including the overt pathogens such as M. avium, M. tuberculosis, M. leprae or M. ulcerans (i.e. Mycobacterium species of Bio safety level 2 & 3, growth after more than 7 days). The phylogenetic relatedness within the slow- and fast-growing corresponds in general with the classification based on traditional phenotypical analyses (Stalh et al. 1990). However, the current rRNA gene based phylogeny does not fully agree with previous numerical phenetic classification or Runyon classification based on pigmentation (Rogall et al. 1990; Pitulle et al. 1992). Both fast- and slow-growing species can be scotochromogenic (pigmented in light and dark), photochromogenic (pigmented in light) or non-chromogenic (not pigmented) (Tortoli 2003). So far, all PAH-biodegrading Mycobacterium isolates have been placed in the phylogenetic branch of the ‘fast-growing’ Mycobacterium species (Guerin et al. 1988; Briglia et al. 1994; Godvidaswami et al. 1995; Bastiaens 1998; Churchill et al. 1999; Poelarends et al. 1999; Yagi et al. 1999; Bastiaens et al. 2000; Schrader et al. 2000; Solano-Serena et al. 2000; Willumsen et al. 2001a). Most PAH-degrading mycobacteria are scotochromogenic and produce smooth round yellow colonies on solid media. PAH-degrading Mycobacterium isolates have been, based on 16S rRNA gene sequence, often assigned to the species M. frederiksbergense (Willumsen et al. 2001a), M. gilvum (Boldrin et al. 1993; Bastiaens et al. 2000; Vila et al. 2001; Gauthier et al. 2003), M. austroafricanum (Bogan et al. 2003), M. vanbaalenii (Heitkamp et al. 1988b; Godvidaswami et al. 1995; Wang et al. 1995; Khan et al. 2001; Moody et al. 2001; Khan et al. 2002), M. hodleri (Kleespies et al. 1996), M. flavescens (Dean-Ross et al. 1996), M. anthracenicum (Wang et al. unpublished) and M. chelonae (Kanaly et al. 2000a; Kanaly et al. 2000b; Kanaly et al. 2002). - 15 -
Chapter 1
The Sphingomonas genus (Yabuuchi et al. 1999) was originally proposed to describe a group of bacterial strains isolated from human clinical specimens and hospital environments. The Sphingomonas genus belongs to the phylum of Proteobacteria, the class of α-Proteobacteria, the order of Sphingomonadales (i.e. the α-4 subgroup) and the non-photosynthetic family of the Sphingomonadaceae (Takeuchi et al. 1994; Kosako et al. 2000) and is represented by the type species Sphingomonas paucimobilis. The genus originally grouped 3 species but has quickly grown to 34 different registered species via reclassification of several old strains and addition of many new isolates (Balkwill et al. 1997; Yabuuchi et al. 1999; Hiraishi et al. 2000). 16S rRNA gene based phylogeny reflects classification of Sphingomonas based on chemotaxonomic characterization via polyamine patterns, polar lipid profiles and fatty acid
composition
(Busse
et
al.
1999).
The
phylogenetic
tree
of
the
Sphingomonadaceae family is complex as the heterogeneous braches belonging to the Sphingomonas genus are intermixed with branches of α-4 subgroup aerobic photosynthetic genera such as Porphyrobacter, Erythromicrobium, Erythrobacter and Sandaracinobacter. In 2001 proposed Takeuchi et al. to divide the heterogeneous Sphingomonas genus in the genus Sphingomonas sensu stricto and three new genera, Sphingobium, Novosphingobium and Sphingopyxis on the basis of phylogenetic and chemotaxonomic analyses (Takeuchi et al. 2001). However, in 2002 Yabuuchi et al. concluded that the genus Sphingomonas should remain undivided at this time, as none of the physiological and biochemical characteristics considered (including cellular lipids and fatty acid composition) provided evidence for the division of the current genus Sphingomonas (Yabuuchi et al. 2002). Based on 16S rRNA gene sequence, most of the cultured PAH-degrading Sphingomonas isolates (Mueller et al. 1990; Weissenfels et al. 1991; Kästner et al. 1994; Balkwill et al. 1997; Bastiaens 1998; Bastiaens et al. 2000) are close relatives of the S. yanoikuyae, S. herbicidivorans and S. chlorophenolica species type strains. These three species cluster phylogenetically in a group formally described as ‘the Sphingobium genus’ (Takeuchi et al. 2001). Over the last years, many new PAH- and pesticide-degrading strains and even new species such as S. xenophaga (Stolz et al. 2000), S. chungbukensis (Kim et al. 2000) or S. cloaca (Fujii et al. 2001) have been added to this ‘Sphingobium’ cluster and have emphasized the environmental importance of this group of Sphingomonas strains. - 16 -
Literature Review
Bacterial dioxygenases involved in biodegradation of PAHs Bacterial PAH-degradation pathways There are 3 types of microbial PAH degradation: mineralization, co-metabolic transformation and non-specific oxidation (Cerniglia 1984). Mineralization reactions will transform the substrate to molecules which can enter the central metabolism (acetate, pyruvate, citrate and methanol) and will lead to complete mineralization of the substrate to inorganic end products such as CO2 and H2O. These reactions will produce energy and carbon molecules which can be used for the production of cell material and reproduction (anabolic reactions). Bacteria that use PAHs as the sole source of carbon and energy, can grow and multiply on sole expense of the PAH compound. In co-metabolism, the PAH compound cannot support microbial growth but is modified and as such degraded when another growth-supporting substrate, mostly another PAH compound, is present. The energy gained from this transformation process is limited and can only support very low growth rates. In most cases, the co-metabolised PAH is not completely mineralized and metabolites may accumulate. Co-metabolism is considered as an important mechanism to degrade PAHs with higher molecular weigths in mixtures of PAHs (Alexander 1980; Keck 1989). Non-specific oxidation of PAHs is based on the activity of highly non-specific enzymes such as the extracellular lignine peroxidase produced by white rot fungus Phanerochaete chrysosponium (Bogan et al. 1996). Chemically, the mineralization of PAHs is a chain of oxidation and reduction reactions. The most common used final electron-acceptor in PAH-degrading bacteria is molecular oxygen. Anaerobic PAH-biodegradation using nitrate or sulfate as final electron acceptor has been recently described but the enzymes and genes involved are not yet identified (Sharak Genthner et al. 1997; Zhang et al. 1997; Rockne et al. 2001; Chang et al. 2002; Eriksson et al. 2003). The complete biochemical pathways for aerobic microbial biodegradation of aromatic compounds including PAHs have been well described (Cerniglia 1984; Cerniglia 1989; Cerniglia 1992; Sutherland et al. 1995; Kanaly et al. 2000b) and can be consulted at the website of
- 17 -
Chapter 1
Biocatalys/biodegradation
Database
of
the
University
of
Minnesota
(http://umbbd.ahc.umn.edu). Initial PAH-hydroxylating dioxygenases Enzymatically, the initial step in the aerobic catabolism of aromatic rings by bacteria involves an oxygenase that catalyzes the ring oxidation producing vicinal cisdihydrodiols. Therefore bacteria produce monooxygenases (incorporate 1 oxygen atom) and dioxygenases (incorporate 2 oxygen atoms) while fungi and mammals only produce monooxygenases. Initial attack of PAHs by bacteria involves substrate specific dioxygenases. Bacterial aromatic ring-hydroxylating dioxygenases are multicomponent enzyme systems that consist of a reductase (flavoprotein), a ferredoxin, and a terminal dioxygenase (Figure 1-1). The reductase and the ferredoxin transport electrons to the terminal dioxygenase which catalyzes the reaction. In the course of the reaction two oxygen atoms, two electrons and two protons are consumed. The terminal dioxygenase, an iron sulphur protein (ISP), is composed of a large (α) and a small (β) subunit. The α-subunit (ISPα) is the catalytic component and contains 2 conserved regions: the Rieske [2Fe-2S] center and the mononuclear iron domain, which are involved in the consecutive electron transfer to the dioxygen molecule. Both α- and β-subunits of the ISP are necessary for function and in determining the substrate specificity of the dioxygenase. The initial ring oxidation with substrate specific dioxygenases is considered the rate-limiting step in the biodegradation of PAHs.
PAHs
dihydrodiol
FIGURE 1-1
ELECTRON TRANSPORT BETWEEN THE SUBUNITS OF A INITIAL PAH-DIOXYGENASE
- 18 -
Literature Review
The aromatic ring-hydroxylating dioxygenases responsible for the first step in the aerobic oxidation of PAHs are substrate and bacterium variable. Naphthalene dioxygenase have mainly been isolated from γ-Proteobacteria, i.e., from a few strains from genera such as Neptunomonas (Hedlund et al. 1999), Cycloclasticus (Geiselbrecht et al. 1998; Kasai et al. 2003; Hedlund et al. unpublished), Pseudoalteromonas (Hedlund et al. unpublished) or Marinobacter (Hedlund et al. unpublished), but mainly from the Pseudomonas genus (Schell 1986; Kurkela et al. 1988; Haigler et al. 1990; Kelley et al. 1990; Herrick et al. 1993; Simon et al. 1993; Dagher et al. 1997; Kosheleva et al. 1997; Geiselbrecht et al. 1998; Bosch et al. 1999b; Hamann et al. 1999; Mordukhova et al. 2000; Park et al. 2000; Ferrero et al. 2002; Park et al. 2002; Olivera et al. 2003) (Table 1-2). The operons coding for naphthalene dioxygenase in Pseudomonas strains are (i) nahAaAbAcAd genes (Schell 1986; Herrick et al. 1993; Simon et al. 1993; Bosch et al. 1999b; Ferrero et al. 2002; Park et al. 2002; Olivera et al. 2003), (ii) ndoAaAbAcAd genes (Kurkela et al. 1988; Simon et al. 1993; Yang et al. 1994; Pellizari et al. 1996; Hamann et al. 1999), (iii) pahAaAbAcAd genes (Kiyohara et al. 1994; Takizawa et al. 1994), and (iv) doxAaAbAcAd genes (Denome et al. 1993). In general, the geneAa encodes the reductase, the geneAb for the ferredoxin, the geneAc for the large α-subunit of the ISP and the geneAd for the small β-subunit of the ISP. Naphthalene dioxygenase genes have recently also been sequenced from β-Proteobacteria such as Commamonas (Moser et al. 2001; Jeon et al. 2003), Ralstonia (nagAaAbAcAd genes) (Fuenmayor et al. 1998; Zhou et al. 2001; Widada et al. 2002), Polaromonas (Jeon et al. 2003), Burkholderia (phnAaAbAcAd genes) (Wilson et al. 2003), Herbaspirillum (Wilson et al. 2003) or Bordetella (Parkhill et al. 2003a), from α-Proteobacteria such as Sphingomonas (bph genes or nsaAaAbAcAd genes) (Romine et al. 1999b; Conradt et al., unpublished), and from some Gram positive bacteria such as Mycobacterium (nid genes) (Khan et al. 2001) or Rhodococcus (narAaAbAcAd genes or bpfAaAbAcAd genes) (Larkin et al. 1999; Treadway et al. 1999; Andreoni et al. 2000; Kulakov et al, unpublished). Related genes have also been found in the genome of Mesorhizobium (Kaneko et al. 2000) or Agrobacterium (Wood et al. 2001) strains or even in an Archae bacterium Thermoplasma (Ruepp et al. 2000). Naphthalene dioxygenases are very versatile enzymes that can catalyze not only the degradation of naphthalene but - 19 -
Chapter 1
catalyze also many dioxygenation and monooxygenation reactions with other aromatic hydrocarbons, including substituted aromatic hydrocarbons and heterocyclic aromatic hydrocarbons (Resnick et al. 1996). The naphthalene dioxygenase of P. putida NCIB 9816/11, also catalyzes the fluorene monooxygenation or dioxygenation. The naphthalene degradation enzymes of other Pseudomonas strains are also involved in the transformation of phenanthrene and anthracene (Menn et al. 1993; Sanseverino et al. 1993; Yang et al. 1994). Genes encoding specific dioxygenases involved in the initial attack of fluorene have to our knowledge only been sequenced from one bacterium, i.e., the Actinomycetes Terrabacter (fln genes) (Habe et al. unpublished) (Table 1-2). Two different dioxygenation fluorene degradative routes which support growth by production of central metabolites have been described (Cerniglia 1992; Grifoll 1992; Boldrin et al. 1993; Monna et al. 1993; Resnick et al. 1996; Casellas et al. 1997; Wattiau et al. 2001; van Herwijnen et al. 2003b). For most strains, initial attack occurs by a monooxygenation at the C-9 of fluorene followed by an angular-carbon dioxygenation. Specific phenanthrene dioxygenase genes have mainly been sequenced from βProteobacteria such as Commamonas (Goyal et al. 1996), Burkholderia (phn genes) (Laurie et al. 1999a; Laurie et al. 1999b) and Alcaligenes (Kiyohara et al. unpublished). The genes and enzymes involved in phenanthrene degradation pathway have recently also been characterized from 2 α-Proteobacteria Sphingomonas strains (adhAaAbAcAd genes) (Pinyakong et al. 2003; Iwabuchi et al. unpublished) and 4 Actinomycetes of the genera Nocardioides (phdAaAbAcAd genes) (Iwabuchi et al. 1998; Saito et al. 2000) or Mycobacterium (pdoAaAbAcAd genes or nidAaAbAcAd genes) (Krivobok et al. 2003). In the phenanthrene and pyrene degrading Mycobacterium sp. 6PY1, two operons encoding a phenanthrene dioxygenase enzyme complex have been sequenced (pdo1AaAbAcAd and pdo2AaAbAcAd genes) (Krivobok et al. 2003). In contrast to the strict selective phenanthrene dioxygenases Phd and Pdo2, the Nid and Pdo1 dioxygenase can also catalyze the dihydroxylation of pyrene. Phenanthrene dioxygenases initiate phenanthrene degradation mostly by an initial dioxygenation at the 3,4-position (Cerniglia 1984). It has been demonstrated that in several Pseudomonas strains naphthalene and phenanthrene share a common - 20 -
Literature Review
upper metabolic pathway, i.e., a common set of enzymes is responsible for the conversion of phenanthrene to 1-hydroxy-2-naphthoic acid as well as that of naphthalene to salicylic acid (Kiyohara et al. 1994; Yang et al. 1994) (Table 1-2). The 1-hydroxy-2-naphthoic acid is oxidized to 1,2-dihydroxynaphthalene, which is further metabolized to salicylate via the naphthalene pathway (Evans et al. 1965; Balashova et al. 2001). In β-Proteobacteria, 1-hydroxy-2-naphthoic acid undergoes ring-cleavage and is further metabolized via o-phthalate and protocatechuate (Kiyohara et al. 1976; Iwabuchi et al. 1998; Pinyakong et al. 2000; Shuttleworth et al. 2000). Recent studies, have shown that for phenanthrene also dioxygenation at the 1,2-position followed by meta-cleavage is possible (Jerina et al. 1976; Pinyakong et al. 2000). In addition, the metabolism of phenanthrene by Streptomyces flavovirens and the marine Cyanobacterium Agmenellum quadruplicatum PR-6 is more similar to that reported in mammalian and fungal enzyme systems than those catalyzed by bacteria. Both oxidize phenanthrene to phenanthrene trans-9,10-dihydrodiol via a monooxygenaseepoxide hydrolase-catalyzed reaction rather than by a dioxygenase. Not much is known about the genetics of specific anthracene dioxygenases. The fluoranthene degradation genes of the α-Proteobacterium Sphingomonas paucimobilis strain TNE12 have been localized on a 240kb plasmid (Shuttleworth et al. 2000), but have not yet been sequenced. Only a few bacterial strains have been identified that can completely metabolize the 4ring PAH pyrene to CO2, and most of them belong to the Mycobacterium genus. Pyrene degradation enzymes and genes have been identified in pyrene degrading Mycobacterium strains M. vanbaalenii PYR-1 (nid genes), Mycobacterium sp. PAH2.135 (RIGII-135) (nid like genes), M. flavescens PYR-GCK (ATCC 700033) (nid like genes), M. gilvum BB1 (DSM 9487) (nid like genes), M. frederiksbergense FAn9 (DSM 44346T) (nid like genes) and Mycobacterium sp. 6PY1 (pdo2 genes). One copy of nidA and nidB like genes were detected in strain PAH2.135, while multiple copies of each gene were detected in the strains PYR-1, BB1, PYR-GCK and FAn9 (Brenza et al. 2003). The different copies could be essentially identical copies or different homologous genes coding for different ring-hydroxylating dioxygenases within the same strain, similar to phenanthrene and pyrene dioxygenase Pdo1 and phenanthrene specific dioxygenases Pdo2 that are co-expressed in Mycobacterium sp. - 21 -
Chapter 1
6PY1 (Krivobok et al. 2003). As mentioned before, for some Mycobacterium strains such as M. vanbaalenii PYR1 (Nid) (Khan et al. 2001) and Mycobacterium sp. 6PY1 (Pdo1) (Krivobok et al. 2003) it has been shown that the pyrene dioxygenases can catalyze dioxygenation of both phenanthrene and pyrene (Table 1-2). Furthermore, all Mycobacterium strains described so far to grow on pyrene also metabolize phenanthrene. In addition, in many Mycobacterium strains, pyrene degradation is induced by phenanthrene (Molina et al. 1999). These findings suggest that the same enzyme systems are involved in the catabolism of phenanthrene and pyrene in Mycobacterium strains, which is consistent with the current knowledge on the catabolic pathways for these 2 PAHs (Krivobok et al. 2003). The biosynthesis of the catabolic enzymes responsible for pyrene degradation in Mycobacterium strains seems to be under strict metabolic control, i.e., production is only induced by the presence of PAHs or their pathway intermediates (Krivobok et al. 2003). Pyrene degradation in Mycobacterium cells proceeds preferentially through the dioxygenation initiated Kivonara pathway (Heitkamp et al. 1988b; Cerniglia 1992; Boldrin et al. 1993; Krivobok et al. 2003). PAHs with more than 4 rings also called high-molecular-weight PAHs (HMW-PAHs) are very recalcitrant or even persistent. So far, no bacteria have been isolated to use these PAHs as source of carbon and energy. Such high-molecular-weight PAHs (HMW-PAHs) are removed from the environment trough cometabolism, using other low-molecular-weight PAHs (LMW-PAHs) as growth substrates. Cometabolic degradation of PAHs with more than 4 rings had been reported for a Mycobacterium (Schneider et al. 1996), a Burkholderia (Juhasz et al. 1997b), and some methanotrophs. Cometabolisation of the potent carcinogen benzo(a)pyrene is known to proceed through substituted pyrene intermediates (Schneider et al. 1996) and depends on the presence of pyrene as cosubstrate (Boonchan et al. 2000; Juhasz et al. 2002).
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Literature Review
TABLE 1-2
GENES CODING FOR INITIAL PAH-HYDROXYLATING DIOXYGENASES IN BACTERIA
Catabolic genes‡ Host strains Naphthalene 1,2-dioxygenase Pseudomonas putida G7 nahAaAbAcAd Pseudomonas putida Pseudomonas putida PaW736 (NCIB 9816-4) Pseudomonas putida 3IIIA2, 5IIANH, 5IIIASal Pseudomonas putida 2IDINH, 3IA2NH, PR1MN1 Pseudomonas putida Cg1 Pseudomonas stutzeri AN10, AN11 Pseudomonas stutzeri B2SMN1, S1MN3, LSMN3 Pseudomonas stutzeri STMN3, ST27MN3, LSMN2 Pseudomonas stutzeri ST27MN2 Pseudomonas stutzeri 63, 67, 85 Pseudomonas balearica LS402, SP401, SP1402, LS401 Pseudomonas fluorescens A24, A88, I-16 Pseudomonas fluorescens LP6a Pseudomonas fluorescens Pseudomonas aeruginosa SCD-1, S1-1 Pseudomonas sp. PR3MN2, 8IDINH, LSMN7, 19IIDNH Pseudomonas sp. ND6 Pseudomonas sp. 5N1-1, 4N4-1, 4N1-3, 4N1-2, 4N1-1 Pseudomonas sp. SOD-3, SCD-3a, SCD-14b Burkholderia sp. SOD-5b, S1-17 Cycloclasticus pugetii PS-1 Cycloclasticus sp. W Neptunomonas napthovorans NAG-2N-126, NAG-2N-113 Mesorhizobium loti MAFF303099 Agrobacterium tumefaciens C58 Streptomyces coelicolor A3(2) Comamonas testosteroni GZ42 Polaromonas napthalenivorans CJ2 Marinobacter sp. NCE312 Bacillus sp. JF8 Bacterium sp. NK3, NK2, NJ2 Pseudoalteromonas sp. EH-2-1 Pseudomonas putida NCIB 9816 ndoAaAbAcAd Pseudomonas putida ATCC 17484 Pseudomonas fluorescens ATCC 17483 Pseudomonas sp. 30-2 Thermoplasma acidophylum Bordetella parapertussis 12822 Bordetella pertussis Tohoma I Pseudomonas sp. doxAaAbAcAd Pseudomonas putida OUS82 pahAaAbAcAd Comamonas testosteroni H Pseudomonas aeruginosa PaK1 Ralstonia sp. U2 nagAaAbAcAd Sphingomonas aromaticvorans F199 bphAaAbAcAd Rhodococcus opacus NCIB12038 bpfAaAbAcAd Rhodococcus sp. 1BN narAaAbAcAd Rhodococcus sp. P200, P400 Mycobacterium vanbaalenii PYR-1 nidAaAbAcAd Ralstonia sp. NI1 phnAaAbAcAd Burkholderia phenazinium Hg8, Hg10, Hg16, Hg14 Burkholderia glathei Hg2, Hg4, Hg11 Cycloclasticus sp. A5 Sphingomonas sp. BN6 nsaAaAbAcAd Burkholderia sp. DNT dntAaAbAcAd Fluorene dioxygenase Terrabacter sp. DBF63 flnAaAbAcAd Phenanthrene dioxygenase Burkholderia sp. RP007 phnAaAbAcAd Sphingomonas chungbukensis. DJ77 Alcalingenes feacalis AFH2 Sphingomonas sp. P2 adhAaAbAcAd Sphingomonas sp. AJ1 Nocardioides sp. KP7 phdAaAbAcAd
- 23 -
Location*
Reference
pNAH7 C pNPL1 pDTG1 C NR NR pCg1 C NR NR NR NR NR NR NR pLP6a NR NR NR pND6-1 NR NR NR NR NR NR NR NR NR NR NR NR NR NR NR NR NR NR NR NR NR NR NR chrom NR chrom pWWU2 pNL1 C NR NR NR NR NR NR NR NR NR NR
(Schell 1986; Harayama et al. 1989) (Boronin et al. 1989) (Simon et al. 1993; Park et al. 2002) (Ferrero et al. 2002) (Ferrero et al. 2002) (Park et al. 2003) (Bosch et al. 1999a) (Ferrero et al. 2002) (Ferrero et al. 2002) (Ferrero et al. 2002) (Olivera et al. 2003) (Ferrero et al. 2002) (Izmalkova et al. unpublished) (McFarlane et al. unpublished) (Min et al. unpublished) (Duncan et al. unpublished) (Ferrero et al. 2002) (Li et al. unpublished) (Bosch et al, unpublished) (Duncan et al. unpublished) (Duncan et al. unpublished) (Geiselbrecht et al. 1998) (Geiselbrecht et al. 1998) (Hedlund et al. 1999) (Kaneko et al. 2000) (Wood et al. 2001) (Bentley et al. 2002) (Jeon et al. 2003) (Jeon et al. 2003) (Hedlund et al. unpublished) (Miyazawa et al. unpublished) (Widada et al. unpublished) (Hedlund et al. unpublished) (Kurkela et al. 1988) (Hamann unpublished) (Hamann unpublished) (Panicker et al. unpublished) (Ruepp et al. 2000) (Parkhill et al. 2003b) (Parkhill et al. 2003b) (Denome et al. 1993) (Takizawa et al. 1994) (Moser et al. 2001) (Takizawa et al. unpublished) (Fuenmayor et al. 1998) (Romine et al. 1999b) (Larkin et al. 1999) (Andreoni et al. 2000) (Chen et al. unpublished) (Khan et al. 2001) (Widada et al. 2002) (Wilson et al. 2003) (Wilson et al. 2003) (Kasai et al. 2003) (Conradt et al. unpublished) (Leungsakul et al. unpublished)
NR
(Habe et al. unpublished)
NR NR NR NR NR NR
(Laurie et al. 1999a) (Kim et al. unpublished) (Kiyohara et al. unpublished) (Pinyakong et al. 2003b) (Iwabuchi et al. unpublished) (Iwabuchi et al. 1998)
Chapter 1
‡
*
Mycobacterium vanbaalenii PYR-1 NR (Krivobok et al. 2003) nidAaAbAcAd Pyrene dioxygenase Mycobacterium vanbaalenii PYR-1 NR (Krivobok et al. 2003) nidAaAbAcAd Mycobacterium sp. 6PY1 NR (Krivobok et al. 2003) pdoAaAbAcAd Mycobacterium sp. S65 NR (Sho et al. unpublished) The dioxygenase gene operon: geneAa encodes the reductase, geneAb encodes the ferredoxin, geneAc encodes the large αsubunit of ISP, geneAd encodes the small β-subunit of ISP. Location of the catabolic genes in the genome: chrom = on the chromosome, p = on a plasmid, NR = not reported; C = via conjugation transferable plasmid
Based on sequence similarity, the PAH-dioxygenases isolated from Gram negative Proteobacteria can be divided into three types which differ from taxa to taxa. Proteobacteria from the α-subcluster seem to have similar initial PAH-dioxygenases that differ from dioxygenases produced by members of the β-subcluster or the γsubcluster. The initial PAH-dioxygenase genes of PAH-degrading α-Proteobacteria were found to be highly conserved among many Sphingomonas species such as S. yanoikuyae strains B1 and Q1, S. aromaticivorans F199, S. ‘agestris’ HV3, S. chungbukensis DJ77, S. xenophaga BN6, Sphingomonas sp. TNE12, Sphingomonas sp. P2 and Sphingomonas sp. EPA505 but divers for the 2 other Proteobacteria gene systems (Bastiaens 1998; Laurie et al. 1999b; Pinyakong et al. 2003a). Similarly, the phenanthrene dioxygenase catabolic genes sequenced from β-Proteobacteria Commamonas strains and Burkholderia strains capable of growing at expense of naphthalene and phenanthrene, showed high homology among each other (Goyal et al. 1996; Laurie et al. 1999a; Laurie et al. 1999b) but low homology with the naphthalene degradation genes conserved within the γ-Proteobacteria. Naphthalene dioxygenase genes are highly conserved among γ-Proteobacterium Pseudomonas strains (87-93% similarity) (Meyer et al. 1999) but clearly different from naphthalene dioxygenase genes purified from Sphingomonas, Mycobacterium or Rhodococcus strains degrading higher molecular PAHs (Hamann et al. 1999; Larkin et al. 1999; Laurie et al. 1999b; Meyer et al. 1999; Treadway et al. 1999). In addition, phylogenetic analysis has indicated that all initial aromatic dioxygenases from Gram positive bacteria form a subfamily of enzymes distinct from the dioxygenases from Proteobacteria (Meyer et al. 1999; Saito et al. 2000; Khan et al. 2001). The large subunit of the initial phenanthrene dioxygenase Phd (PhdA) from Nocardioides was 57% identical to large subunit of Nid enzyme (NidA) sequenced from phenanthrene and pyrene degrading M. vanbaalenii PYR-1 (Khan et al. 2001; Khan et al. 2002; Brenza et al. 2003; Krivobok et al. 2003). All dioxygenases sequenced from
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Literature Review
Mycobacterium strains were found to be highly homologous (minimum 98% identity) (Khan et al. 2001; Khan et al. 2002; Brenza et al. 2003; Krivobok et al. 2003). The strong conservation of the nahAc homologs among the Gram negative naphthalene degrading isolates suggest in situ horizontal gene transfer (Herrick et al. 1997). Similarly, the high similarity and the exclusive presence of the nid genes in PAH-degrading Mycobacterium strains and not in non-degrading related strains of the same species, suggest a common origin of PAH-dioxygenase genes from Mycobacterium strains and that Mycobacterium strains obtained the nid genes later in evolution, possibly by horizontal transfer (Brenza et al. 2003). Catechol-hydroxylating dioxygenases In a second step, a dehydrogenase will transform the product of the oxygenation reaction, i.e., the cis-dihydrodiol to a dihydroxylated aromatic intermediate, i.e., a catechol (Cerniglia 1992; Juhasz et al. 2000b). These catechols may then be processed through either an ortho-cleavage type pathway (ring cleavage between the 2 hydroxyl carrying carbon atoms) or meta-cleavage type pathway (ring opening next to the hydroxyl carrying carbon atoms) (Figure 1-2) (Juhasz et al. 2000b). The final metabolites produced through both pathways (succinate, fumarate, pyruvate, acetate and acetaldehyde) will be further metabolized through central metabolic pathways for the synthesis of new cellular components or the production of energy and will be mineralized to CO2 and H2O. The ortho-pathway is catalyzed by intra-diol dioxygenases or catechol-1,2-dioxygenase and the meta-pathway by extra-diol dioxygenases or catechol-2,3-dioxygenase (C23O). The majority of bacteria growing with PAHs as sole sources of carbon and energy follow the meta-cleavage pathway using a catechol 2,3-dioxygenase (C230) (Cerniglia 1992; Grifoll 1992; Boldrin et al. 1993; Monna et al. 1993; Resnick et al. 1996; Casellas et al. 1997; Meyer et al. 1999; Wattiau et al. 2001; van Herwijnen et al. 2003b). However, enzymes of both the meta- and ortho-pathway of catechol degradation were shown to operate in the process of naphthalene degradation by Pseudomonas strains (Kulakova et al. 1989; Kosheleva et al. 1997). Moreover, P. putida strains degrading naphthalene via the ortho-pathway were found to be more
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Chapter 1
competitive than the P. putida strains degrading naphthalene via the meta-pathway in mixed chemostat cultures on naphthalene (Filonov et al. 1997). In Gram negative bacteria, anthracene is mostly metabolized via a meta-cleavage pathway analogous to that of naphthalene metabolism to yield salicylate and catechol (Cerniglia 1984; Cerniglia 1992; Sutherland et al. 1995). In Gram positive bacteria, on the other hand, anthracene is mostly mineralized via the ortho-cleavage pathway through o-phthalate and protocatechuate (Dean-Ross et al. 2001; van Herwijnen et al. 2003a).
FIGURE 1-2
THE ORTHO- AND META- DEGRADATION PATHWAYS FOR AROMATIC RING CLEAVAGE OF A CATECHOL
Catechol 2,3-dioxygenase gene sequences have been determined for PAH-degrading bacteria of the genera Pseudomonas, Sphingomonas (Yrjala et al. 1998; Meyer et al. 1999) and Burkholderia (Laurie et al. 1999b). Based on sequence similarity, the catechol-dioxygenases isolated from Gram negative Proteobacteria can be divided into three types which differ from taxa to taxa. Proteobacteria from the α-subcluster seem to have similar catechol-dioxygenases that differ from dioxygenases produced
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Literature Review
by members of the β-subcluster or the γ-subcluster. Except the catabolic genes for the lower
fluorene
degradation
pathway
of
Sphingomonas
sp.
LB126,
i.e.,
protocatechuate catabolic genes, which seemed to be more closely related to the genes previously found in lignin-degrading Sphingomonas sp. SYK-6 than to the corresponding genes of other PAH-degrading Sphingomonas strains (Pinyakong et al. 2003a; van Herwijnen et al. 2003b). It has been speculated that the ability of Sphingomonas strains isolated from the subsurface (deeply-buried sediments in the Atlantic ocean) to degrade a wide array of aromatic compounds including PAHs represent an adaptation for the utilization of sedimentary lignite, the major source of organic carbon for heterotrophic organisms in that environment (Fredrickson et al. 1999). Location, organization and regulation of the PAH-degradation genes In γ-Proteobacteria Pseudomonas strains, the naphthalene degradation genes have mostly been located in polycistronic operons (Burlage et al. 1989; Menn et al. 1993). However, the gene order of the catabolic genes of β-Proteobacteria Burkholderia and Commamonas is significantly different from the classical naphthalene degradation nah-like systems of Pseudomonas strains (Goyal et al. 1996; Laurie et al. 1999a). The genes for PAH-degradation in α-Proteobacteria Sphingomonas were often complexly arranged, i.e., the genes necessary for one degradation pathway were scattered over several operons and gene clusters (Feng et al. 1997). The metabolic pathways for different monocyclic and polycyclic aromatic hydrocarbon degradation were often linked through a grouped organization and co-regulation of the catabolic genes involved (Kim et al. 1999; Romine et al. 1999a). The catabolic genes for aerobic oxidation of PAHs such as naphthalene and fluorene are localized in the chromosome or on plasmids. In γ-Proteobacteria Pseudomonas strains, the naphthalene degradation genes are often located on plasmids such as pNAH7 (Schell 1986; Burlage et al. 1989; Selifonov et al. 1991; Menn et al. 1993), pNPL-1 (Kulakova et al. 1989; Selifonov et al. 1991; Kozlova et al. 1999), pBS2 (Kulakova et al. 1989; Selifonov et al. 1991; Kozlova et al. 1999), pBS216 (Kulakova et al. 1989), pBS217 (Kulakova et al. 1989), pBS3 (Selifonov et al. 1991), pBS4 (Selifonov et al. 1991), pDTG1 (Simon et al. 1993; Park et al. 2002), pCg1 (Park et - 27 -
Chapter 1
al. 2003), pND6-1 (Li et al. unpublished) and pLP6a (McFarlane et al. unpublished). It has been shown that the Pseudomonas PAH-catabolic systems located on plasmids are thermo-sensitive as the catabolic plasmids were eliminated at 41-42 °C. Some of these catabolic plasmids had even an inhibition effect on growth of Pseudomonas strains at an elevated temperature, as plasmid free mutants could grow much better on elevated temperature (Kochetkov et al. 1983). Plasmids containing PAH degradation genes have also been isolated from Sphingomonas strains such as S. aromaticivorans F199 (the 184kb conjugative pNL1 plasmid) (Romine et al. 1999b) and Sphingomonas sp. KS14 (a >500kb mega-plasmid). The presence of large plasmids, but with unknown function, has also been demonstrated in many other PAHdegrading Sphingomonas strains (Bastiaens 1998; Fredrickson et al. 1999; Romine et al. 1999a; Bastiaens et al. 2000). So far no plasmids carrying catabolic genes have been identified in Mycobacterium strains (Bastiaens 1998; Bastiaens et al. 2000; Brenza et al. 2003). The genes for PAH-catabolism from Mycobacterium strains are therefore thought to be localized on the chromosome. Although PAH-catabolic genes from Gram-positive and Gram-negative bacteria are phylogenetically different, they have nevertheless many similarities both in sequence, location and gene organization, suggesting that they have a common although distant, evolutionary origin (Harayama et al. 1992; Mason et al. 1992; Pinyakong et al. 2003a). Conclusions Microbial degradation is considered to be the major route through which PAHs are removed from contaminated environments and therefore is considered as feasible remediation technology. Currently, in situ and ex situ bioremediation techniques are, however, still very inefficient for removal of PAHs from contaminated soil. As any other technology engineered biological remediation of PAH-contaminated sites can only become successful if the active system is known and relative controllable. This requires more information about the identity of the degrading soil microorganisms and the enzymatic mechanisms they use. PAH-degrading microbial communities have been detected in many different contaminated habitats. Ecological analysis of microbial communities in PAH- 28 -
Literature Review
contaminated environments has shown that many different bacterial genera and species can be involved in the degradation process. The environment conditions and the type and concentration of PAHs seem to select for specific genera and species independent of the geographical location of the samples. Mostly Proteobacteria from the genera Pseudomonas (γ), Burkholderia (β) and Sphingomonas (α) or Corynebacterineae from the genera Nocardia, Rhodococcus and Mycobacterium seem to be involved in PAH-biodegradation in soil. The first group of Gram negative bacteria seems to be degrading preferentially lower molecular PAHs such as naphthalene and phenanthrene, while the second group of Gram positive bacteria is more specialized in the degradation of higher molecular PAHs such as pyrene. These PAH-degrading bacteria seem to have adapted to their hydrophobic PAH-substrates by showing low nutrient requirements and by making use of bioavailability promoting systems such as production of biosurfactants, high substrate affinity, mouth-like uptake systems or close contact biofilm formation. Mycobacterium species seem to be specialized in using hydrophobic sorbed or organic dissolved PAHs while Pseudomonas and Sphingomonas strains prefer aqueous liquid systems. Several aerobic microbial PAH-degradation pathways have been identified. The catabolic genes from Gram-positive and Gram-negative bacteria coding the dioxygenase enzymes are phylogenetically different, and are conserved on the genus level. In Pseudomonas and Sphingomonas strains, PAH-catabolic genes are often localized on conjugative plasmids while in Mycobacterium species they seem to be chromosomal. The genes of different genera have nevertheless many similarities both in sequence, location and gene organization, suggesting that they have a common although distant, evolutionary origin. The information on bacterial PAH-metabolic processes so far will allow the development of new tools that will enable researchers to analyze PAHs contaminated sites efficiently and to determine more rapidly strategies for treating them. For example, new nucleic acid probes and biologically based sensors could be constructed. It also provide the capacity to modify organisms for improved degradative performance through new metabolic pathway construction - either to complete a degradative process or to broaden its specificity to accommodate previously untreatable molecules - and to manipulate the genetic regulatory elements to improve efficiency.
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Chapter 1
MICROBIAL BIOREMEDIATION OF PAH-CONTAMINATED SOILS Technologies available for remediation of PAH-contaminated soil To clean-up PAH contaminated soil and sludge different soil remediation technologies are available. Some of these technologies allow in field treatment (in situ) while others demand excavation (ex situ) with (off site) or without (on site) transport. The choice of the most appropriate remediation technique for a given polluted site depends on (A) pollutant characteristics, such as (i) the type and concentration of the PAHs, (ii) the type and concentration of the co-contaminant, (iii) the pollution history, and (B) soil characteristics, such as (i) pollution of top soil or/and unsaturated zone or/and aquifer, (ii) soil permeability for air and water, (iii) the location of the soil, (iv) soil temperature, pH, natural chemical composition and presence of natural terminal electron acceptors, (v) soil nature microflora and (vi) stability of soil structures.
Physico-chemical soil remediation The conventional techniques used for remediation have been the excavation of the contaminated soil and transport it to a landfill or containment of the contaminated areas by capping. However, the first approach moves the contamination elsewhere and may create significant risks during the excavation, handling and transport of hazardous material. Additionally, land fill sites are increasingly expensive for disposal of the material. The ‘cap and contain’ method is only an interim solution since the contamination remains on site, requiring long term monitoring and maintenance of the isolation barriers. A better approach is to completely remove the pollutants from the contaminated soil matrix or if possible, to destroy the pollutants by transforming them in innocuous substances. Mainly ex situ and off site thermal, physical and chemical extraction and decomposition techniques have been used. In the ‘thermal desorption’ technique the polluted soil is heated to 600 °C and PAHs are partially degraded by pyrolyse and transported from the solid phase to the vapor phase (Sutherson 1997). The airflow removing the PAHs is treated in a controlled combustion reaction completely - 30 -
Literature Review
oxidizing the PAH compounds to CO2 and H2O. ‘Soil washing’ is a combination of physical and chemical extraction methods to separate PAHs from the soil, or at least separate highly contaminated soil fractions from less contaminated fractions (Sutherson 1997). Water and chemicals are intensively mixed with the polluted soil, improving desorption and solubilisation of the PAHs in the water phase. Sometimes active carbon particles are mixed to enhance extraction of PAHs by sorption on the carbon particles. Organic substances and small soil particles are removed based on their particle size and density by pumping the soil slurry through a series of hydrocyclons and flotation beds. The pollutant concentrations in the ‘washed soil’ fraction are reduced below legal limits and can be reused while the separated fraction, highly enriched in PAHs, is incinerated or deposited. Such ex situ physico-chemical techniques are very effective in reducing contaminant concentrations but have several disadvantages, such as their technological complexity, the high cost for small-scale application, and the lack of public acceptance. Soil washing will generate a highly pollutend final soil fraction wich still needs to be dumped or incinerated. Incineration, has been protested as it may increase the exposure for both the workers at the site and nearby residents. Incineration will also result in combustion of all organic material and sterilization of the soil. The treated soil will loose all biological life and a large part of its economical value and can only be used in the construction industry. In addition, the large scale size of many contaminated areas makes the physico-chemical approach highly costly and unfeasible. Alternatively, for some sites in situ and on site extractive remediation techniques can be used. In porous soils, the light molecular weight volatile fraction of PAHs can be removed from soil by soil vapor extraction combined with air infiltration. Clean air is pumped in the soil above (‘airventing’) or below (‘airsparging’) the ground water table under pressure while soil air is extracted from the dry zone under vacuum. In addition, a soil water extraction (‘pump and treat’) approach can be used to remove the water-soluble but non-volatile PAH fraction from contaminated aquifers in porous sandy soils. Pump-and-treat can be combined with the infiltration of water containing detergents to leach out the pollutants (‘soil leaching’ method). When high concentrations of PAHs are trapped in a non-aqueous phase (NAPL) floating on the water table, selective extraction of NAPL under vacuum can be used (‘slurping’ method) combined with air and water extraction techniques applied on the capillary - 31 -
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zone. ‘Chemical oxidation’ has proven to be an efficient in situ treatment for remediation of PAH-contaminated soil and ground water (Barton et al. 2000; Siegrist 2000). However, oxidation with peroxide, ozone, permanganate or ultrasonic radiation will only partially degrade the PAH-compounds.
Biological soil remediation Bioremediation has emerged the last decennia as an alternative technology for the clean up of hydrocarbon contaminated soils (Vidali 2001). Bioremediation is defined as the process in which organic contaminants are biologically degraded under controlled conditions to an innocuous state, or to levels below concentration limits established by regulatory authorities (Vidali 2001). Microorganisms naturally present in the soil such as bacteria and fungi are stimulated to degrade or transform pollutants to less toxic compounds. As such, bioremediation is a relatively low-cost and lowtechnology approach, which generally has a high public acceptance and often can be carried out on site (Wilson et al. 1993) (Table 1-3). The efficiency of bioremediation is however site- and pollutant- dependent (Table 1-3). The range of contaminants on which it is effective is limited, the time scales involved are relatively long, and the residual contaminant levels may not always be appropriate (Raffi et al. 1994) (Table 1-3). In addition, considerable expertise may be required to design and implement a successful bioremediation program, due to the need to thoroughly assess a site for its suitability and to optimize conditions to achieve a satisfactory result (Table 1-3). TABLE 1-3
ADVANTAGES & DISADVANTAGES OF BIOREMEDIATION
Advantages • high public appreciation • relatively low cost • can often be carried out in situ or on site • (minimal disturbance, lower risks for contaminant distribution due to transport & handling) • useful for a wide variety of organic contaminants complete destruction of the contaminant • residues are usually harmless (CO2, H2O, biomass)
Disadvantages • extrapolation from bench & pilot-scale studies to full-scale field operations is very difficult • relatively long treatment time required • very much dependent on environmental factors and contaminants mixture and dispersion • limited to those compounds that are biodegradable • residues may be more persistent or toxic than the parent compound
Several ex situ bioremediation approaches have been developed to solve the contamination problem more rapidly (Table 1-4). These techniques involve the
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Literature Review
excavation of contaminated soil from ground and its transport to a ‘bioremediation facility’. Soil remediation ‘bioreactors’ allow full control and optimization of the degradation process (Vidali 2001) (Table 1-4). Depending on the reactor type the contaminated water, soil or sludge is treated as pure liquid (aqueous reactor), in slurry with 5 to 40 % dry weigth (slurry reactor) or as dry soil (dry soil reactor or DSR). Extensive mixing conditions allow maximum interaction between bacteria, soil particles, contaminants and nutrients stimulating biodegradation. Full control of degradation parameters such as aeration, temperature, nutrient addition and bacteria inoculation could severely reduce the remediation time from months to weeks or days (Cookson 1995) (Table 1-4). In general, the rate and extent of biodegradation are higher in a bioreactor system than in in situ or ex situ solid-phase system because the contained environment is more manageable and hence more controllable and predictable. The use of ex situ bioreactors on lab or pilot scale has been found relatively effective for remediation of soil containing a complex PAH mixture (Wilson et al. 1993; Bastiaens 1998). For real field applications, the running costs for bioreactors are generally higher than other ex situ and in situ treatments (Figure 1-4). ‘Landfarming’, ‘biobeds’ or ‘biopiles’ are ex situ techniques where the contaminated soil is spread out on an impermeable underground (Vidali 2001). Supplements to neutralize soil pH and/or to improve soil texture (sand, straw, compost or wood chips) and sometimes extra nutrients and/or bacteria are mixed with the soil. Tilling of the biobeds on a regular time scale improves oxygen and nutrient distribution. In addition, water is percolated and air can be pumped through the soil to improve biodegradation conditions. To reduce atmospheric pollution by volatile hydrocarbons, biobeds can be placed indoors or covered with plastic to capture and treat the air. When high concentrations of non-hazardous organic additives such as manure, straw or agricultural wastes are mixed with the soil, elevated temperature can be reached in the biobed due to degradation of organic material and the term ‘composting’ may be used instead of landfarming (Semple et al. 2001). In general, these techniques have relatively low monitoring and maintenance costs and relatively high clean-up liabilities compared to other ex situ treatments (Vidali 2001) (Table 1-4) (Figure 1-4). Landfarming and composting has been successfully applied on a large scale to different soil types contaminated with low to middle distillated hydrocarbon mixtures such as diesel fuel, jet fuel and gasoline (Tien et al. 1999; Semple et al. 2001). - 33 -
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However, biodegradation of PAHs by landfarming has been much less successful (Wilson et al. 1993). Landfarming methods could reduce PAHs contamination levels within a reasonable period of time but only PAHs with three or fewer aromatic rings were degraded (Wilson et al. 1993; Semple et al. 2001). In addition, in situ bioremediation techniques can be applied to the soil and groundwater (Table 1-4). These techniques are generally the most desirable bioremediation option to remediate sites in use and/or sites with low-level contamination of surface soil over a vast area due to lower cost and the minimal disturbance avoiding excavation and transport of contaminants. The currently most applied engineered in situ bioremediation technologies include ‘bioventing’ and ‘biosparging’ (Radway et al. 1996) (Table 1-4). Through extraction and infiltration of air and water in the soil, it is possible to increase the bioavailability of the PAHs and the oxygen supply for the aerobic PAH-degrading bacteria in the vadose respectively water-saturated zone (Vidali 2001). ‘Biological permeable reactive barriers’ or ‘bioscreens’ can be installed at the borders of the contaminated site perpendicularly to the groundwater flow to contain and to treat the PAH contaminated ground water in order to protect downstream areas. The PAHbiodegradation is locally stimulated by the addition of nutrients and oxygen in the bioscreen zone. After passing through the bioscreen the ground water will be free of contaminants. As this technique requires no active pumping the operational costs are low (Figure 1-4). A new in situ technique currently in development combines the use of electrokinetics to transport water, ions and bacteria through PAH polluted clay soils to stimulate in situ biodegradation (Ho et al. 1995). Distribution control and natural attenuation, i.e., biodegradation without human interference, could be an alternative in situ bioremediation strategy (Table 1-4). Such passive in situ bioremediation processes can be economically very attractive but requires regular monitoring and longer times to reach desired pollutants levels (Figure 1-4).
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TABLE 1-4
SUMMARY OF BIOREMEDIATION STRATEGIES
Application Ex situ
Technology Slurry reactors Bioreactors
Ex situ
Landfarming Composting
In situ
Bioventing Biosparging Natural attenuation
Benefits Rapid degradation kinetics Optimized environmental parameters Enhanced mass transfer Effective use of inoculants and additives Cost efficient Low cost Can be done on site Most cost efficient Noninvasive Relatively passive Treats soil & water
COST
Limitations Soil requires excavation Relatively high cost capital Relatively high operating cost Soil requires excavation Space requirements Extended treatment time Need to control abiotic loss Environmental constraints Geological constraints Extended treatment time Monitoring difficulties
Slurry reactor Bioreactor Intensive landfarming Extensive landfarming Active In situ bioremediation Passive In situ bioremediation
TIME
FIGURE 1-4
RELATIONSHIP
BETWEEN TREATMENT TIME AND COST FOR BIOREMEDIATION
METHODS
Assessment and monitoring of biodegradation potential A first step in a bioremediation strategy would be ‘a feasibility assessment and field evaluation’. The microbial community naturally present in the polluted soil is analyzed and tested for its potential to degrade the PAHs under predetermined laboratory conditions to identify limiting factors and recommend ways to mitigate these limitations in the field (Balba et al. 1998). The intrinsic metabolic potential of a contaminated soil is mostly assessed by degradation tests in the lab using laboratory microcosm systems (Heitzer 1993). Contaminated soil samples are incubated in the lab under different conditions to
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screen bioremediation treatments and select the most appropriate strategy for large scale application (Balba et al. 1998). These microcosms can vary in complexity from simple static soil jars to highly sophisticated mini bioreactor systems containing dry soil or soil slurry. There are different ways to assess the PAH-degradation potential in such microcosm systems. A first method is by measuring the total dehydrogenase activity in the soil as biological oxidation of organic compounds is generally a dehydrogenation process that is catalyzed by dehydrogenase enzymes (Balba et al. 1998). The most widely used method for the determination of soil dehydrogenase activities is the colorimetric method, involving the used of 2,3,5-triphenyl tetrazolium chloride (TTC) which acts as an electron acceptor for many dehydrogenase enzymes and which is reduced to form the red compound triphenyl formazan (TPH). The intensity of the red color produced from the dehydrogenase assay is a good index for microbial activities within the tested soil. However, several soil compounds such as nitrate, nitrite and ferric ions may interfere with the dehydrogenase activity by ions acting as alternative electron acceptors. An alternative method to assess the total biological activity in the sample is by measuring the total O2 consumption or total CO2 production rate (respirometry). This approach provides rapid, relatively unequivocal time-course data suitable for testing different biological treatment options, such as the effect of nutrient supplementation, microbial inoculation, etc. on microbial activity. To investigate more specifically the PAH-degradation potential, the mineralization of freshly added
14
C-labeled PAHs is
measured (radiorespirometry) (Spain et al. 1980; Spain et al. 1983; Grosser et al. 1991; Carmichael et al. 1997; Reid et al. 2001). To assess the relative contribution of bacteria and fungi to the mineralization of PAHs, selective inhibitors such as cycloheximide (fungal inhibitor) or a mixture of penicillin B and tetracycline (bacterial inhibitors) can be added to identical sets of respirometers, along with the 14
C-labeled PAHs (Macleod et al. 2002). However, added inhibitors are only limited
effective due to natural resistance to them and their short period of action due to deactivation and sorption to the soil matrix (active for 2-4 days depending on the soil type) (Anderson et al. 1973; Anderson et al. 1975; Stamatiadis et al. 1990). In addition, specific removal of competitors by adding selective inhibitors may result in - 36 -
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a faster growth of surviving microorganisms (Anderson et al. 1975). In addition, also the residual concentrations of PAHs and their degradation products can be monitored during treatment in a sub-sampling or a batch set-up (liquid or gas chromatography). A more direct approach to asses the PAH-degradation potential of a soil, is to look for the presence of bacteria with known PAH-degrading capacities or known bacterial PAH-degradation enzyme systems. Culture-dependent detection of bacteria is mostly done by microbial enumeration by plate counting on appropriate media, for example media with PAH as sole C-source (Balba et al. 1998). However, the agar plate microbial-counts technique has several limitations particularly when dealing with non-culturable microorganisms in soil. Culture-independent molecular methods based on the detection of bacterial DNA, RNA or enzymes involved in PAH-degradation are often preferred to give more rapidly information about the presence and activity of PAH-degrading in the soil. DNA or RNA extracted from soil is used as template in specific PCR reactions. Specific 16S rDNA based detection methods allow to screen the soil for the presence of unknown types of bacteria in general or certain groups of bacteria in specific. The catabolic DNA coding initial PAH ring-hydroxylating dioxygenases and the catecholcleaving dioxygenases have been used as targets to detect the presence of PAH degradation potential at the DNA level. A limited number of specific or degenerate probes and PCR primers have been designed based on known DNA sequences encoding the initial ring-hydroxylating dioxygenases and the C23O meta ringcleaving dioxygenases of certain species (Hamann et al. 1999; Meyer et al. 1999). However, PAH degradation enzymes can be quite different at the DNA level without affecting their function. Many catabolic genes with different sequence may exist and may not be detected (Widada et al. 2002), leading to an underestimation of the degradation potential. It is therefore necessary to gain more sequence information for different PAH-degrading strains of different taxa to establish a universal or group specific PCR and RT-PCR protocols for the detection of PAH degradation genes and mRNA, respectively (Meyer et al. 1999; Widada et al. 2002). The total size and activity of a bacterial community can be assessed by measuring the 16S rRNA:rDNA ratio of the cells as they grow in the soil (Ka et al. 2001). To follow the activity of certain species or strains, this technique can be combined with specific 16S rDNA oligonucleotide hybridization probes. - 37 -
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Limitations and strategies to improve engineered bioremediation of PAHcontaminated soil Currently, mostly the soil is basically treated as a ‘black box’ during bioremediation. Not much is known about the key organisms needed for PAH-biodegradation and the specific needs of PAH-degrading bacteria with respect to nutrition and environmental conditions for their optimal activity in the soil environment. Bioremediation can only be effective where environmental conditions permit microbial growth and activity. Bioremediation therefore often involves the manipulation of environmental parameters to allow microbial growth and degradation to proceed at a faster rate. Bioaugmentation The first criterion for successful bioremediation is the presence of microorganisms in the soil that actually are capable of degrading the pollutant, i.e., microorganisms that posses appropriate enzymatic systems to catalyze the degradation reactions. It is assumed that whenever the pollutants are biodegradable and the soil conditions are favorable, suited bacterial communities to do the job will most likely have naturally evolved (Ide et al. 1996). Most characterized PAH degrading isolates are able to degrade 1 or a few different PAHs. However, consortia of different bacteria or maybe new engineered strains with as wide spectrum of degradation enzymes are needed for the degradation of complex PAH mixtures in soil (Bouchez et al. 1995; Bastiaens 1998). A cell density of 106 to 108 CFU g-1 dry weight is mentioned as a minimum cell concentration to obtain suitable biodegradation rates (Ramadan 1990). If not ‘sufficient’ bacteria capable of mineralizing PAHs are present in the soil, a bioaugmentation strategy can be an option, i.e., the addition of pollutant degrading bacteria indigenous or exogenous to the contaminated site. The positive effects of bioaugmentation on small scale are reported many times (Grosser et al. 1991; Hickey et al. 1993; Weir et al. 1995; el Fantroussi et al. 1997). Different authors have shown in laboratory and some pilot scale experiments the successful biodegradation of artificially PAH-contaminated soils by the addition of PAH degraders belonging to the genera Mycobacterium and Sphingomonas. 2methylnaftalene, phenanthrene, pyrene and benzo(a)pyrene degradation was increased - 38 -
Literature Review
by the inoculation of sediment and water microcosms with M. vanbaalenii PYR1 (Heitkamp et al. 1989). The reintroduction of 107 to 108 Mycobacterium sp. PAH135 cells per gram of soil could significantly increase the mineralization of pyrene (Grosser et al. 1991). The degradation of pyrene in soil was enhanced 10 times by adding a Mycobacterium gordonae-like strain or a Sphingomonas paucimobilis strain (Kästner 1998). A fluorene degrading Sphingomonas strain was succesfull inoculated in a ‘Dry Soil Reactor (DSR)’ containing 70 kg of a non sterile PAH-contaminated soil (70 % dry weight) and allowed relatively fast PAHs removal in a relatively short treatment time (Bastiaens 1998). The high degradation capacity and the bioavailability promoting characteristics make consortia of Sphingomonas and Mycobacterium strains especially suited for bioaugmentation of PAH polluted soil. Reports of successful in situ or ex situ large scale bioaugmentation however, are less numerous. Points limiting the use of bioaugmentation in real site bioremediation are (i) the inefficient large scale cultivation of the bacteria to inoculate, (ii) the inefficient inoculation and distribution of the bacteria in the soil, and (iii) the limited colonization, survival and activity of the inoculated bacteria under the given soil or remediation conditions. To treat large amounts of polluted soils huge amounts of bacteria are needed. Therefore inoculum bacteria need to be efficiently and economically cultivated in large volume batch bioreactors to obtain sufficient biomass. This asks for cheap and well defined media satisfying the needs of the selected inoculum bacteria. However, using rich media allowing quick growth to high densities migth create bacterial cells that have difficulties to adapt and survive the harsh oligotrophic soil environment. Cultures pregrown on or in the presence of the target pollutants containing actively working enzyme machinery would be far more suited. Very little research has been done to develop good propagation conditions for large scale inoculation of bacteria in soil. So far, almost exclusively batch cultures have been used, but continuous cultivation could be economically more interesting. In addition, in most bioaugmentation cases only 1 single pure strain is used instead of a well adapted consortium of several strains attacking mixed pollutions. Using mixed inoculum cultures increases the need for efficient propagation methodologies. Bioaugmentation attempts were often also lacking a good inoculation methodology. Lab scale and pilot scale experiments have shown that the introduction of bacteria in - 39 -
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real environment can lead to rapid reduction of viability and activity (Roszak et al. 1987; Ramadan 1990; Weissenfels et al. 1992; Weir et al. 1995; Kästner 1998). Many inoculation parameters can influence the behavior of the added bacteria such as (i) the size of the inoculum, (ii) the inoculation medium and procedure, and (iii) the inoculum distribution in the soil. It was found that small inocula were incompetent (Ramadan 1990; Weir et al. 1995). In many lab experiments degradation activity increased linear with the inoculum size (Goldstein et al. 1985; Roszak et al. 1987; Kästner 1998). In general, cell concentration of 107 to 108 cells per gram soil is presented as the inoculum size limit for positive results in bioaugmentation. Bacteria can be added to the soil in liquid medium through simple sprinkling or mixing or could be added immobilized on for example polyurethane foam (Manohar et al. 2001). Mostly inoculum cells are suspended in buffers or media but not without any risks. It has been shown that the introduction of PAH-degrading bacteria suspended in a mineral medium had a negative influence on the degradation capacity of the indigenous and introduced microflora, while inoculation as a pure aqueous suspension did not affect the degradation activity (Kästner 1998). Immobilization has been reported to enhance the survival without decreasing the activity of added bacteria in soil and waste water during bioremediation processes. Sphingomonas cells bound to alginate or chitosan beads were successfully re-used several times in a labscale bioreactor for the treatment of nonylphenol contaminated industrial wastewater (Fujii et al. 2000). In addition a good mixing is critical for the uniform distribution of the cells and for the efficiency of the remediation process. For in situ bioaugmentation, subsurface irrigation techniques have proven to be much more efficient than surface irrigation as a microbial delivery tool (Mehmannavaz et al. 2002). Studies have shown that under natural conditions most motile endogenous and added soil bacteria have a limited mobility. For example several 2,4-dichlorphenol or p-nitrophenol degrading motile Pseudomonas strains were incapable to move through the pores of the soil matrix to locations with higher polluent concentrations (Goldstein et al. 1985). As a consequence, pollutants will only be degraded in the near vicinity of the inoculation points.
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After introduction, the inoculum bacteria will have to survive in the soil environment. The main cause for quick reduction in cell number in non sterile soil is probably the lack of competitiveness of many pure strain inocula adapted to lab cultivation conditions (Goldstein et al. 1985; Vidali 2001). Inocula will have to compete with the indigenous microflora for nutrients or colonization of niches. Only bacteria with high substrate affinity (specific biodegrading enzymes, biosurfactants, etc.) will be able to grow on poorly bioavailable pollutants such as PAHs in soil environments. It has been shown that the success of bacterial inocula to degrade PAH in soil was also depending on the concentration of organic material limiting PAH-bioavailability (Weissenfels et al. 1992). For some inoculum bacteria, polluent concentrations may be too low to deliver enough carbon and energy to sustain growth or to induce the enzymatic machinery needed for polluent mineralization (Goldstein et al. 1985; Ramadan 1990; Colores et al. 1999). On the other hand, high pollutant concentrations could inhibit biodegradation through nutrient- and oxygen depletion or through toxic effects of the pollutant on the inoculum (Leahy et al. 1990; Volkering et al. 1995).
Biostimulation Apropriate PAH-degrading microorganisms in contaminated soil are not necessarily present in numbers sufficient for bioremediation of the site within a reasonable time period. The slow rate of PAH-biodegradation in soil in comparison to degradation rates in laboratory culture conditions, suggests that microbial activity is seriously constrained in the natural environment. The inherent properties of the PAH compounds and the soil properties are seriously affecting the biodegradation (Manilal et al. 1991; Wilson et al. 1993) (Table 1-5). To solve this problem and improve bioremediation efficiency, a better understanding of the interaction between the soil environment, the PAH-compounds and the biocatalysts is needed. For successful bioremediation of PAH-contaminated sites a system must be created in which the right soil microorganisms can be promoted to create and maintain sufficient biomass and to be metabolically active. In addition, process control is not only necessary to produce sufficient bacterial biomass but also to direct them towards optimal PAHbiodegradation activity.
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Nutrient availability for PAH-degrading bacteria in soil In the environment, bacteria depend on the availability of several elements or nutrients in order to allow the cells to survive and build up new biomass (Baker et al. 1994; Sutherson 1997). There are three categories of nutrients based on the quantity and essential need by the microorganism, i.e., macro-, micro-, and trace nutrients. The macronutrients carbon (C), nitrogen (N) and phosphorus (P) are known to comprise 50%, 14% and 3% dry weight, respectively, of a typical microbial cell (Vidali 2001). The micronutrients sulfur (S), potassium (K) and sodium (Na) comprise 1% dry weight and calcium (Ca), magnesium (Mg) and chloride (Cl) 0.5% dry weight of a cell but play important roles in membrane and enzyme stability (Bailey et al. 1986). The most common trace elements are iron (Fe), manganese (Mn), cobalt (Co), copper (Cu) and zinc (Zn). These trace nutrients are only present in low quantities in the cells but are essential for enzyme functioning. Based on this approach, it was suggested that optimal nutrient mixes for biomass formation should have an overall C/N/P-ratio of approximately 120/19.4/3.9 [w/w/w] (Paul et al. 1989). However, in real field situations mostly the C/N/P-conditions in soil are not optimal for the growth and activity of bacteria. Organic and inorganic substrates (carbon, nitrogen, phosphate, sulfur) can be added ex situ or in situ to balance the Carbon/Nitrogen/Phosphorus concentration ratio in soil. A C/N/P-ratio of 120/10/1 [mg], has been reported as optimal for cell growth on contaminants in general in a review about bioremediation of soil contaminated with PAHs (Wilson et al. 1993) (Table 1-5). More specific for optimal PAH-degradation in soil, an optimal overall C/N/P ratio of 120/2/0.15 [mg] was mentioned. Lately, most researchers recommend for bioremediation applications an optimal C/N/P mole-ratio of 100/10/1 [mole], i.e., a 120/14/3 ratio [mg] (Hoeppel et al. 1994; Bouchez et al. 1995; Cookson 1995). •
PAHs as carbon source
The principle of PAH-biodegradation is that heterothrophic soil microorganisms will use the carbon supplied by the PAHs for growth and energy. As exception to what
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Literature Review
was previously said, carbon is the only substrate that is mostly needed at higher quantities then the chemical composition of the cell would predict. In general, a simple growth substrate can sustain a microbial yield of 60%, meaning that from 100mg of C only 60mg is incorporated into biomass and the other 40mg is used for energy production and transformed into CO2 (Paul et al. 1989). However, C from most pollutant compounds is mostly not substantially incorporated into microbial biomass, i.e., a larger fraction is mineralized to CO2 for the production of energy (Paul et al. 1989). This leads to higher pollutant concentrations needed to sustain microbial biomass (Paul et al. 1989). The limited bioavailability of the PAH-compounds itself is believed to be the major bottleneck for efficient biological treatment of PAH-contaminated soil (Weissenfels et al. 1992; Erickson et al. 1993; Luthy et al. 1994; Beck et al. 1995; Würdemann et al. 1995). The rate at which microorganisms can mineralize hydrocarbons depends on (i) the rate of transfer of the compound to the microorganism and (ii) the rate of uptake and metabolism of the compound in the cell (Bosman et al. 1997) (Figure 1-5). Currently, most researchers believe that microorganisms can use PAHs only in dissolved state. Consequently, PAH solubilisation (desorption) and/or diffusive transport through the aqueous solution to the cell surface will affect PAHbiodegradation (Figure 1-5). For naphthalene and phenanthrene it was indeed shown that the growth rates of PAH-degrading bacteria were independent of the total PAH surface area presented to the bacteria, i.e., independent of the amount of solid-PAH, but limited by the solubility of the PAH compounds (Bouchez et al. 1995). It was shown that limited solubilisation and transport caused very slow growth and linear growth curves, indicating mass-transfer limited conditions, for bacteria that grew on 3-chlorodibenzofuran (Harms 1995). Based on this dissolved-based utilization, bioavailability and biodegradation of PAHs in soil is seriously limited by their hydrophobicity and concurrent sorption effects. PAHs tend to ad/absorb to the soil organic matter or accumulate strongly in non aqueous phase liquids. Sorbed PAHs desorb slowly and are therefore believed to be less available for biological uptake and degradation (Volkering 1992; Weissenfels et al. 1992; Volkering et al. 1993). The bioavailability of hydrocarbons in soils decreases with increasing molecular mass of the component (Cerniglia 1992) and with increasing soil hydrocarbon contact time, i.e., aging (Bauer et al. 1985; Weissenfels et al. 1992; Hatzinger et al. 1995; Macleod - 43 -
Chapter 1
et al. 2000; Reid et al. 2000). As such, even massive contaminations may be oligothrophic from a microbe’s perspective. Theoretical considerations show that for extremely hydrophobic contaminants as PAHs, the low solved available pollutant concentrations may not be sufficient to sustain the population size or to induce the degradation enzymes needed to achieve environmental clean-up (Goldstein et al. 1985). However, not all the researchers completely agree with this hypothesis and suggest that sorbed compounds are available to microorganisms without prior desorption (Guerin 1992; Volkering 1992; Mihelcic 1993). Some researchers state that the sorption of PAHs as thin films to the soil particles can lead to a better degradation by bacteria in comparison to the crystalline PAHs, due to increased substrate surface. The positive interaction between increasing substrate surface and a higher degradation and growth rate, due to higher mass transfer rates, has been described many times for mixed and pure cultures (Thiem et al. 1994; Tongpim et al. 1996; Bastiaens 1998).
FIGURE 1-5
MICROBIAL UPTAKE OF A SORBED POLLUTANT
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It is known that many PAH-degrading bacteria use original strategies to enhance bioavailability of the highly hydrophobic compounds. Some bacteria can increase the transport rate to and the uptake rate in the cells so that biodegradation rates exceeds the normal dissolution rates of the PAHs from solid or organic-solved state (Manilal et al. 1991; Bouchez et al. 1997; Bastiaens 1998; Tang et al. 1998; Garcia-Junco et al. 2001a; Wick et al. 2001; Carcia-Junco et al. 2003). These bioavailability promoting properties are (i) the excretion of biosurfactants to solubilize the PAH (Deziel et al. 1996; Willumsen et al. 1997; Yuste et al. 2000; Garcia-Junco et al. 2001a), (ii) high specific substrate affinity to increase PAH uptake rate (Button 1985; Wick et al. 2002a), (iii) attachment to the hydrophobic PAH crystal, the sorbed PAH, or the degradable or non-degradable organic phase solvent containing the PAH to decrease the cell–molecule distance (Ortega-Calvo et al. 1994; Harms 1995; Tongpim et al. 1996; Bouchez et al. 1997; Garcia-Junco et al. 2001a; Wick et al. 2002a; Wick et al. 2002b; Rodrigues et al. 2003). Up till now, still little is known about the physiological and biochemical mechanism involved in bacterial uptake of pollutants in the sorbed state. Even the specific physico-chemical characteristics of the substrate itself seem to determine the bacterial strategy regarding uptake, as some PAH induces biofilm formation and other not in the same strain capable of growing on both (Rodrigues et al. 2003). For some PAH-degrading Mycobacterium strains, there are strong indications based on electron microscopic observations that they posses special mechanisms to make very close contact with PAH-crystals (Wick et al. 2001; Wick et al. 2002a). As such, the persistence/bioavailability of PAHs in the environment thus not only depends on the physical and chemical characteristics of the PAHs, the composition and chemical characteristics of the sorbent but also on the specific properties of the PAH-degrading microorganisms present. During engineered bioremediation the limitations of PAH-solubility can be reduced or solved in different ways. A chemical pretreatment consisting of an oxidation of the PAH structure with permanganate or ultrasonic radiation can partially degrade the PAH-molecules and enable microorganisms to metabolize them at a rate suitable for remediation purposes (Barton et al. 2000). The addition of synthetic nonionic surfactants, such as Triton X-100, Tergipol NPX, Brij 35 Igepal CA-720 or the oleophilic fertilizer Inipol EAP-22, in concentrations above their CMC to soil could increase the apparent solubility and therefore the - 45 -
Chapter 1
bioavailability and the biodegradation of PAH-compounds. However, experimental observations on the effects of surfactant addition on microbial degradation of hydrophobic hydrocarbons are not consistent (Aronstein et al. 1991; Edwards et al. 1991; Laha et al. 1992; Wilson et al. 1993; Efroymson et al. 1995; Volkering et al. 1998). The application of a surfactant has been reported to be beneficial (Volkering et al. 1995; Grimberg et al. 1996) or detrimental (Thibault et al. 1996) to microbial substrate utilization rates and growth yields. Possible beneficial effects are (i) the increase of aqueous solubility of the hydrocarbons, (ii) the reduction of interfacial tension that promotes formation of more interfacial area, (iii) the emulsifying action of the surfactant that overcomes interfacial area limitation and that permits effective contact between cells and substrate, or (iv) the liposome facilitated substrate transport through the microbial cell wall. Possible causes of negative effects could be (i) the toxicity of the surfactants to the bacteria, (ii) the sorption of the surfactant onto the soil, (iii) the CMC effect, i.e., the unavailability of the PAHs trapped in the micelles to the bacteria, (iv) the bacterium-surfactant interactions that effect cell membranes or prevent the bacteria to adhere to the hydrophobic hydrocarbon surfaces, or (v) the preferential metabolisation of nontoxic organic surfactants. In addition, provision of high surfactant concentrations on large scale is relative expensive and therefore preferably avoided. An alternative way to meet the limited C-source availability in PAH-contaminated soils could be the addition of extra C-substrates to the soil to maintain sufficient biomass and an active soil population. Historically, Broadbent and Norman (Broadbent et al. 1946) reported the stimulation of mineralization of one compound by added extraneous C-substrates already in 1946 and coined the term ‘priming effect’ to describe it. Non-inducer organic nutrients added to the soil could stimulate the soil microbial community and support co-metabolic transformation of PAHs. Shen and Bartha (Shen et al. 1996) demonstrated that the addition of glucose during bioremediation tests can stimulate the mineralization of organic carbon in soil. The stimulation of survival and PAH-degradation in soil by addition of extra C-substrates like skim milk has been reported by Weir (Weir et al. 1995) for a seeded Pseudomonas sp. that mineralizes phenanthrene. However, most researchers found that easy metabolisable organic nutrient amendments such as glucose or amino acids had no effect or interfered negatively with the degradation activity (Goldstein et al. 1985; Swindoll et al. 1988; Heitkamp et al. 1989; Carmichael et al. 1997). The supplements - 46 -
Literature Review
probably limit the metabolisation of the PAHs because of rapid growth of non-PAH degrading microorganisms which out compete the PAH-degrading bacteria. Alternatively, the inhibitory effect could be due to ‘diauxic growth’, i.e., the preferential use of the supplement as carbon source above the pollutant by the same microorganisms. However, diauxic growth could be prevented if the substrates are added in multiple small doses (Harder et al. 1982). At low substrate concentrations (oligotrophic conditions) mixed-substrate growth is more common then diauxic growth (Kovarova et al. 1998). To stimulate a certain group of microorganisms, sometimes some specific organic pathway intermediates, i.e., inducers, are added. The survival of six naphthalenedegrading strains was significantly increased by adding salicylate, an intermediate in the naphthalene degradation pathway, to non-sterile soil samples (Ogunseitan et al. 1991). The biodegradation rate of naphthalene and phenanthrene in artificially contaminated soils was enhanced by addition of salicylate, while the degradation of pyrene was enhanced by the addition of gentisate, phtalate, cinnamate and propionate (Carmichael et al. 1997). These supplements had no effect however on PAHdegradation in natural historically contaminated soils or soils containing high PAHconcentrations. Moreover, at high supplement concentrations inhibition of degradation was observed due to accumulation of upstream metabolites. Steffensen and Alexander (Steffensen et al. 1995) suggest that the effect of one biodegradable substrate on the metabolisation of the second substrate is mostly related to the competition for inorganic nutrients in a mixed bacterial community.
•
Inorganic macronutrients: Nitrogen and Phosphorus sources
Microorganisms need besides organic compounds as carbon sources (C) also macronutrients such as nitrogen (N) and phosphorus (P) for cell growth. Competition for nitrogen by plant roots, mycorrhizal fungi and microorganisms keep inorganic N and P concentrations in soil low. It is even suggested that the competition for inorganic nutrients among the microbial population can reduce the number of active PAH-degrading bacteria to an insufficient level, since the slower growing pollutant degrading bacteria are outcompeted (Cerniglia 1992; Steffensen et al. 1995). Nitrogen (N) is primarily used for cellular growth (NH4+ or NO3-) and as an alternative electron
- 47 -
Chapter 1
acceptor (NO3-) instead of oxygen. Bacteria will use nitrogen to build amino acids and proteins for catabolism and to incorporate it in the peptidoglycan of their cell walls. In most soils, ammonia released during soil organic matter decomposition is rapidly transformed to nitrate through microbial nitrification processes. Phosphorus (P) is needed in the bacterial cell for the production of nucleic acids, phosphate sugars and phospholipids in the cell membrane. Phosphorus exists in nature in a variety of organic and inorganic forms but primarily in either insoluble or poorly soluble inorganic forms. The P-solubility in soil is regulated by a complicated equilibrium controlled by ion-ion association reactions, pH-effects and soil texture. Phosphorus tends to absorb to clay minerals. It exists mainly as apatites, with the basic formula M10(PO4)6X2. Commonly the mineral (M) is calcium, iron or aluminum and the anion (X) is a fluor, cloor, hydroxy or carbonate ion (Paul et al. 1989). These bound forms are poorly extractable and therefore also considered as poorly bioaccessible. The portion that is extractable with water, diluted acids or bicarbonate solutions is designated as ‘available’ for uptake by living organisms. The concentration of free phosphorus in the soil solution is mostly of the order 0.1-1 ppm. Addition of inorganic nutrients to balance N- and P-concentrations with the CPAHconcentrations in soil has been practiced to eliminate possible nutrient limitations in soil. Nitrogen is the nutrient most commonly added during bioremediation. Mostly nitrogen is added as urea (NH2CONH2), also called carbamide or carbonyldiamide, or as ammonium chloride (NH4Cl), but is also supplied as an ammonia-/nitrate- salt (e.g. NaN03) or ammonium nitrate (NH4NO3). All these forms are readily assimilated in bacterial metabolism. However, many studies have indicated that ammonium, already being in the reduced state required for incorporation into amino acids, is preferred to the oxidized form nitrate which first has to undergo reduction. Even low levels of ammonia often repress the enzymes required for nitrate reduction (Paul et al. 1989). Nevertheless, some features of the supplements may be important depending on the soil situation. The application of urea in the unabsorbed form such as in manure will lead to high volatilization. Surface manure in the fields may lose up to 50% of its nitrogen due to volatilization (Paul et al. 1989). The ammonium ion (NH4+) creates an increased oxygen demand due to nitrification processes. In addition, while the anionic nitrate is very soluble and freely mobile in soil solution, the cationic ammonium is retained by soil cation exchange sites on clays (Bohn 1985). Ammonia also reacts - 48 -
Literature Review
with soil organic matter to form quinone-NH2-complexes. In addition, plants prefer ammonia to nitrate and thus ammonium can also be lost to plants. Nitrate is more susceptible to losses through leaching and denitrification (Paul et al. 1989). Fertilizer nitrogen, added as urea (NH2CONH2), ammonia (NH3) or the ammonium ion (NH4+) form, is also subject to nitrification (Paul et al. 1989). Gaseous nitrogen oxide (N2O) addition under high vapor pressure has been used as an alternative to distribute nutrients better in situ throughout soil (Radway et al. 1996; Bogan 2001). This new approach has been utilized in a patent process but it has not, however, been documented as a means of enhancing the remediation of PAH-contaminated soils (Bogan 2001). Phosphorus is the second most commonly added nutrient in bioremediation and is supplied to serve as a source for cellular growth. As additional inorganic phosphorsource mostly solid monophosphate salts (also called orthophosphate salts) (e.g. KH2PO4, K2HPO4, K3PO4) or polyphosphate salts (also called pyrophosphate salts) (e.g. KH3P207) with potassium or sodium are added. Gaseous phosphate under the form of triethylphosphate (C3H9O4P) (TEP) and tributylphosphate (C12H27O4P) (TBP) have been used as an alternative for liquid or solid phosphate amendments (Radway et al. 1996; Bogan 2001). Although mildly toxic and corrosive irritants, TEP and TBP are considered the safest phosphorus compounds, which can readily be gasified. When added to soil, phosphate quickly absorbs, however, to iron and aluminum oxide surfaces and may even form precipitates with iron, aluminum, manganese and calcium (Brady 1990). Such reactions will limit its transport and causes it to be unavailable for biological activity (Johnson et al. 1990; Aggerwal et al. 1991; Ward et al. 1999). The addition of potassium phosphate may also accelerate the cleavage of hydrogen peroxide (H2O2) sometimes added simultaneously to the soil as an oxygen source for aquifer bioremediation. If the H2O2 cleaves immediately, this may generate microbial colonies growing only near the injection well and the oxygen source may be depleted before it reaches the contaminated zone (Ward et al. 1999). In most geochemical environments, 10 mg l-1 is the maximum orthophosphate addition to avoid significant precipitation, peroxide cleavage or toxicity (Riser-Roberst 1998).
- 49 -
Chapter 1
The use of inorganic N and P supplements have been tested with a variety of organic pollutants, pure and mixed microbial cultures and environmental matrices. Unfortunately, the effects of nutrients on pollutant degradation rates in soil are very inconsistent. Addition of inorganic nutrients have been found to shorten the adaptation period for microbial degradation (Wiggins et al. 1987; Swindoll et al. 1988) or to increase the extent and rate of microbial metabolism of some pollutants in soil and water (Bossert et al. 1984; Swindoll et al. 1988; Manilal et al. 1991; Alexander 1994; Baker et al. 1994). Sometimes stimulation of pollutant degradation rates only appeared until days or weeks after nutrient addition (Jobson et al. 1974; Bossert et al. 1984). Many articles report no apparent effect on metabolism (Swindoll et al. 1988; Johnson et al. 1990) or even a decreased metabolism in other cases (Johnson et al. 1990; Manilal et al. 1991; Morgan et al. 1992). Lag time, initial rate, and degradation extent were affected singly or as a group by nutrient additions (Thorton-Manning et al. 1987; Swindoll et al. 1988). Moreover, in several reports there is a significant difference between the effects caused by the addition of N-salts, P-salts or both (Swindoll et al. 1988; Manilal et al. 1991). Phosphorus might be more often the limiting macronutrient in soils. A vague trend could be that P-supplements have more often a positive effect while N-supplements cause more frequently inhibition of biodegradation of organic pollutants. Inorganic salts of nitrogen and phosphorus can be very effective in closed systems but have the tendency to wash out in real outdoor field applications (Leahy et al. 1990). Very few or non overriding generalizations emerge from the studies that could deliver some practical information for real field applications. Differences in the bioavailability of nutrients added to soil may explain a portion of the observed differences in the response of degradation rates to nitrogen and phosphorus supplements. Just as sorbed carbon substrates are generally considered to be unavailable to soil microbes, sorbed inorganic nutrients also may be less bioavailable than nutrients dissolved in the soil solutions. Morgan and Watkinson (Morgan et al. 1992) suggested that high concentrations of supplements could lead to inhibition of microorganisms that are used to live in oligothrophic soil environments. Another important conclusion from all these studies may be that the responses to nutrient addition are largely depending on the soil that is used.
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Literature Review
•
Micronutrients & Trace-elements
Next to macronutrient-concentrations, also micronutrients like sulphur (S), calcium (Ca), and magnesium (Mg) and trace-elements like the metals iron (Fe), manganese (Mn), cobalt (Co), copper (Pb) and zinc (Zn) or vitamins (V) and amino acids (AA), are essential for the survival and activity of soil bacteria (Swindoll et al. 1988). It has been reported that biostimulation (stimulation of indigenous bacteria) and bioaugmentation (adding bacteria) are performed best in the presence of relative high levels of micronutrients (Ward et al. 1999) (Table 1-5). Soil properties effecting microbial PAH-degradation in a non-nutritional way In addition, biodegradation rates are controlled by characteristics of the soil that are not directly involved in bacterial nutrition. The availability of oxygen (O2) and redoxpotential (Eh), humidity (H2O), acidity (pH), salinity (IS) and temperature (T) play an important role in the survival and activity of soil bacteria (Table 1-5). Control and tuning of these soil parameters such as temperature, pH, moisture or redox-potential is much more difficult. All supplements added to a soil may have serious wanted or unwanted impact on one or more of these factors. Therefore, careful consideration must be taken in determining the quantity and type of nutrients to add so that optimal IS, pH and Eh conditions are maintained or created.
•
Oxygen & Redox-potential
The biodegradation of PAHs is mostly an aerobic process in which molecular oxygen (O2) functions as final electron acceptor (EA) in energy generating biochemical pathways. Moreover, oxygenases incorporate oxygen atoms (O) in the aromatic ring of the PAHs. The availability of oxygen for PAH-degradation in soil, however, is controlled by many parameters such as the soil type, the water saturation level of the soil, the presence of other metabolisable substrates that can lead to oxygen depletion and the microbial consumption rate. In soil, aerobic degradation can only take place if a minimum of 10% of the pore space is filled with air (Sims et al. 1993; Wilson et al. 1993; Hurst et al. 1996) (Table 1-5). If air fills less then 1% of the pore volume only
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Chapter 1
anaerobic processes can take place. Low oxygen concentration in soil has been identified as one of the major parameters limiting the biodegradation of petroleum in saturated soil and ground water (Leahy et al. 1990). Oxygen availability is also expressed in the redox-potential, which defines the total terminal electron availability in the environment. The redox-potential affects the oxidation states of hydrogen, carbon, nitrogen, oxygen, sulfur, etc. The redox-potential of submerged sediments may range from –300 mV for anoxic sediments up to +700 mV for highly aerobic sediments. The lack of oxygen in reduced or anaerobic sediments causes obligate and facultative anaerobic microorganisms to utilize other compounds such as nitrate, sulfate or iron ions as electron acceptors. Although aromatic ring reduction and hydrolytic cleavage under anaerobic conditions has been reported for monoaromatic and polyaromatic hydrocarbons, in general, low redox-potentials decrease the rate of PAH-biodegradation (Cerniglia 1984; Ward et al. 1999). For an aerobic environment, the redox-potential must be above +50 mV. Values for Eh of 100 mV to 400 mV generally indicate relative low oxygen concentrations, but are acceptable for aerobic biodegradation (Sims et al. 1993; Barden 1994). High positive Eh values (400 mV to 800 mV) indicate well-aerated conditions that are optimal for aerobic degradation (De Laune 1981). To increase the oxygen amount in the soil it is possible to till or sparge air. In some cases, oxygen releasing compounds such as hydrogen peroxide or magnesium peroxide (e.g. ORC®) can be introduced in the environment (Rau et al. 2001). Also the supplemental nutrients may react with the compounds in the subsurface or provide an alternate TEA, thus changing the Eh. •
Moisture
Soil moisture is essential to biodegradation since the majority of microorganisms live in the water film surrounding soil particles. Biological degradation processes can only take place if sufficient water is present to keep the soil microorganisms active. Moisture is for soils generally expressed in terms of ‘water activity’ (aw) or ‘water holding capacity’. The humidity of soil regulates the equilibrium between the waterand gas phase within the pores of the soil matrix and many other physical and chemical soil processes like transport, adsorption, complexation of nutrients, etc.. Tests have shown that a humidity of 70-90% water holding capacity produces optimal - 52 -
Literature Review
metabolic rates with a variety of organic pollutants (Hurst et al. 1996; Carmichael et al. 1997) (Table 1-5). A soil humidity range of 70-90% of the water holding capacity is also reported to be optimal for oil and PAH-biodegradation (Sims et al. 1993; Wilson et al. 1993; Barden 1994). Higher water contents is probably one of the reasons why slurry reactors perform better then dry solid reactors or in situ treatments for PAH-degradation. Nutrients and PAHs are more dissolved and more available for bacterial uptake when more water is present. However, in these systems an appropriate mixing or aeration system is crucial since in 90% humidity most of the soil pores are filled with water so that oxygen transfer can become the limiting factor for optimal degradation. In situ, irrigation may be needed to achieve the optimal moisture level (Vidali 2001).
•
Acidity
Soil pH is an indicator of hydrogen ion activity in the soil. The pH plays a major role in many physical en chemical soil processes. Since PAHs do not posses acidic or basic groups in their elemental structure, the pH will not directly affect the PAHs. Nevertheless, the behavior of PAHs in soil can be seriously affected by the pH. If the pH rises, humic acids in the soil become more negatively charged, they will loosen their 3-dimensional structure and the sorption of the PAHs will diminish (Chiou et al. 1979; De Laune 1981; Thurman 1985; Murphy 1994; Bastiaens 1998). Most soil organisms can live with pH ranging from 5.5 to 8.5. A pH in the range of 5 to 9 is generally acceptable for biodegradation; a pH of 6.5 to 8.5 is considered optimal (Cookson 1995). In specific, a pH range of 7.0-7.8 was suggested as optimal for microbial PAH-degradation (Wilson et al. 1993). Low pH could negatively affect bacterial growth and thus degradation activity (Baath 1996; Kästner 1998; Alden 2001). Soil pH also affects the availability of other inorganic nutrients. The solubility of phosphorus is maximized at a pH value of 6.5 (Barden 1994). If a soil is too acidic, it is possible to adjust the pH by adding lime (e.g. CaCO3). Neutralization of soil is generally discussed to be favorable for the biodegradation of mineral oil components and PAHs (Leahy et al. 1990; Shen et al. 1996) (Table 1-5).
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Chapter 1
•
Salinity
Salinity is mostly expressed in terms of ‘ionic strength’ (I). The ionic strength refers to the total concentration of ions in solution. The contribution of ions to the ionic strength of the solution increases with the charge of the ion (Formula 1-1). The ionic strength is usually measured as electrical conductivity and calculated via a simplified linear relationship (Formula 1-2).
FORMULA 1-1
I = ½ (Σci*Zi2)
with
ci the molar concentration of ion i Zi the charge of ion i
FORMULA 1-2
I = 0.000016 * conductivity
with conductivity in µS/cm or µmhos/cm
The ionic strength is important for bacteria as it effects electrostatic (Coulombic) interactions between molecules in many biochemical reactions. Proteins have charged surfaces and electrostatic forces determine the interaction of the proteins with their substrates. High salt concentrations such as 0.1 to 0.5 M KCl (i.e. I of 0.1-0.5) can cause dissociation of bacterial membrane proteins (Alden 2001). For bacterial cells the solution may not exceed the maximum salt concentration of 4% w/v NaCl, i.e., I of about 0.700 (Wilson et al. 2003). Experimentally it has been shown that increasing the I of the aqueous solution also increases the extent of bacterial sorption to a variety of natural and artificial surfaces, and enhances the bacterial retention in sand columns in transport experiments (van Loosdrecht et al. 1990a; van Loosdrecht et al. 1990b; Fontes 1991; Gannon 1991; Mills 1994). In addition, the salinity of the groundwater can seriously influence the solubility and the sorption of PAHs to soil-particles (Shiaris 1989). The solubility of naphthalene, phenanthrene and benzo(a)pyrene decreases with increasing I (Eganhouse et al. 1976). In addition, an increase in I also leads to an increase of the sorption-coefficients of PAHs to organic material (Karickhoff et al. 1979). In case of a high I of the soil solution, the negatively charged groups of the humic acids will be shielded by salt ions what allows increased - 54 -
Literature Review
absorption of hydrophobic components such as PAHs to the organic particles (Engebretson et al. 1994). In general, for biological processes it is favorable that the soil solution does not exceed an electrical conductivity of 2000µS/cm, i.e., an I of approximately 0,032 (Barden 1994) (Table 1-5). The negative effect of high salinity on the PAH-degradation activity of soil bacteria has been reported several times in estuarine sediments (Kerr et al. 1988; Shiaris 1989) or inoculated soil (Kästner 1998). •
Temperature
Temperature directly affects the rate at which PAH are degraded by microorganisms in natural ecosystems. The soil temperature influenced the biodegradation of organic pollutants by affecting (i) the chemical and physical properties of hydrocarbons, (ii) the rate of metabolic processes in microorganisms and (iii) the composition of the microbial community (Leahy et al. 1990). Seasonal fluxes in heterotrophic activity and PAH degradation rates have been detected several times with the highest activities in summer (20°-25°C) and the lowest activities in winter (Shiaris 1989; Barden 1994). It is unclear whether temperature-related differences in hydrocarbon degradation result from seasonal selection of psychrophilic or mesophilic hydrocarbon-degrading microorganisms or from low-temperature suppression of the degradative capacity of stable PAH-degrading microbial populations that persist during the year. In general, the biodegradation rate increases with increasing soil temperature, and most soil microorganisms show highest activity between 20 and 30°C (Cerniglia 1984; Bauer et al. 1985; Wilson et al. 1993) (Table 1-5). Naphthalene, phenanthrene or anthracene mineralization was not detected at extremes incubation temperatures of 5 and 45°C (Cerniglia 1984; Bauer et al. 1985). It has been shown that the Pseudomonas PAH-catabolic systems located on plasmids are thermosensitive as the catabolic plasmids were eliminated at 41-42 °C. Some of these catabolic plasmids had even an inhibition effect on growth of Pseudomonas strains at an elevated temperature, as plasmid free mutants could grow much better on elevated temperature (Kochetkov et al. 1983). For on site and in situ techniques, plastic covering can be used to enhance solar warming in late spring, summer and autumn (Vidali 2001).
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Chapter 1
•
Mineralogy & granularity
The mineralogical composition of soils can have a great influence on the bioavailability of PAHs. PAHs sorb strongly to humic acids or clay (Ortega-Calvo et al. 1998). In addition, mineralogy can also affect the availability of other nutrients such as nitrogen and phosphorus as discussed previously. Soil structure controls the effective delivery of air, water and nutrients. Hence soils with low permeability may not be appropriate for in situ clean-up techniques. To improve soil structure in an ex situ approach, materials such as gypsum or organic matter can be added (Vidali 2001).
•
Presence of co-contaminants
Most successful PAH-biodegradation experiments in soil are on lab-scale and use an artificial contaminated soil polluted with one or a selected group PAHs and a selective group of soil bacteria added to the soil. In reality, PAH-contaminated soils mostly contain a complex mixture of pollutants that could interact with each other and with the bacteria degrading the PAHs. PAH contamination always includes a mixture of many PAHs. The quick and complete degradation of mixture of PAHs always require a consortium of different bacteria mineralizing different PAHs. The degradation process of one PAHs by one specific bacterium can be, however, seriously inhibited by the presence of another PAH compound (Bauer et al. 1985; Bouchez et al. 1995), although this inhibition effect was less significant in soil (Bastiaens 1998). A strong similarity in chemical structure between PAHs allows interactions on many different steps in the degradation process, i.e., competition for active zones of the degradation enzymes, accumulation of toxic waste products or interaction on the level of enzyme induction. Heavy metals can be additional pollutants in PAH-contaminated soils. Such heavy metals could disrupt biodegradation processes (Springael et al. 1994). Some PAH-degrading bacteria have combined PAH-degradation capacities with resistance to heavy metals (Kozlova et al. 1999). Also cyanides are very frequent present in PAH-contaminated soils from for example coal gasification plants.
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Literature Review
•
Soil microflora
Other microorganisms present in the soil could have a deleterious effect on PAHdegrading community (Goldstein et al. 1985). They could produce natural antimicrobial substances such as antibiotics or toxins. In addition, bacteriophages and protozoa are wide spread in the environment and could attack or graze on degrading bacteria (Ramadan 1990; Leser et al. 1995).
TABLE 1-5
Parameter
ENVIRONMENTAL CONDITIONS AFFECTING PAH-DEGRADATION IN SOIL
Conditions required for general microbial activity
Optimum values for Oil & PAH-degradation
Nutrient content
N and P for microbial growth
C/N/P 120/19/4 [mg] 120/10/1 [mg] 100/15/3 [mg] 100/10/1 [mg]
Oxygen content
Aerobic degradation process: minimum air filled pore space of 10%
10-40% O2 Buried lifts of bed bioreactor 2% to 5%
Redox potential
Aerobes and facultative aerobes: 50 mV > RP < 500 mV
50 mV > RP < 500 mV
Soil pH
5.5-8.8
6.5-8.0
Temperature
15-45 °C
20-37°C
IS
0 - 0,056
< 0,032
Soil moisture
25-85% of water-holding capacity
70-90%
Contaminants Heavy metals Type of soil
Not to toxic Total content < 2000 ppm Low clay or silt content
5-10% oil or PAHs of dry weight of soil Total content < 700 ppm Low clay or silt content
References (Paul et al. 1989) (Wilson et al. 1993) (Bouchez et al. 1995) (Cookson 1995) (Breedveld et al. 2000) (Vidali 2001) (Wilson et al. 1993) (Vidali 2001) (Hurst et al. 1996) (Sims et al. 1993) (Barden 1994) (Cookson 1995) (Wilson et al. 1993) (Vidali 2001) (Cerniglia 1984) (Bauer et al. 1985) (Vidali 2001) (Barden 1994) (Sims et al. 1993) (Barden 1994) (Vidali 2001) (Vidali 2001) (Vidali 2001) (Vidali 2001)
Conclusions Microbial degradation is considered to be the major route through which PAHs are removed from contaminated environments and therefore a feasible remediation technology. Currently, in situ and ex situ bioremediation techniques are, however, still very inefficient for removal of PAHs from contaminated soil. Two main problems, however often block the efficient biological treatment of PAH-contaminated soil: (i) the limited bioavailability of these compounds due their hydrophobicity and (ii) the
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Chapter 1
natural soil conditions that reduce the survival and the activity of PAH-degrading bacteria. Bioremediation is a very soil and site-specific process. If the soil shows no or little natural degrading activity, a consortium of PAH-degrading and nutrients could be added in order to boost bioremediation. The successful implementation of one nutrient-bacteria combination at one site for one contaminant, however, does not guarantee success when applied to another soil type. Currently, there are no specific methods for determining the exact nutrient sources and working conditions one should utilize at a polluted site soil to stimulate bacterial activity. Little practical information about the specific needs of PAH-degrading microorganisms in the soil environment is available. Some studies have reported some possible nutrition and environmental requirements for optimal bacterial activity, but only few could draw conclusions towards final field applications. Moreover, non of these studies were specifically directed towards bacteria specialized in PAH-degradation like Mycobacterium spp. and Sphingomonas spp.. In order to successfully stimulate and use PAH-degrading strains for bioaugmentation, it is crucial further research reveals the true specific macro-, micro- and even trace nutrient requirements of these soil bacteria specialized in PAH-degradation. Especially, since application of nutrient supplements on fullscale includes considerable expenses. One challenge is to learn how microorganisms cope with complex mixtures of PAHs and organic substrates in soil and what occurs when mixed groups of microorganisms are put to work at degrading single or mixed contaminants. Secondly understanding the effects of nutrient type and quantity may enable to develop strategies to change inhibitory situations for the better. So, it will become possible to draw comparisons across different sites dependent upon the sitespecific soil characteristics such as nutrient prevalence and other hydro-geological characteristics. In addition, a better understanding of the nutritional needs and degradation kinetics will be useful for the development of mathematical models for carrying out general predictive analysis for bioremediation problems. There are also important engineering constraints affecting practically the entire field of bioremediation such as problems of how to deliver nutrients, the nature of particular sites, the delivery of organisms, the configuration of reactor systems, and cost and time considerations. - 58 -
Literature Review
Many different environmental and antropogeneous factors influence the survival and activity of PAH-degrading bacteria in the soil. As any other technology engineered biological remediation of PAH-contaminated sites can only become successful if the active system is known and relative controllable. This requires that the degrading soil microorganisms are identified, promoted in the creation and maintenance of an active biomass and directed towards optimal PAH-biodegradation.
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“We have a habit in writing articles published in scientific journals to make the work as finished as possible, to cover up all the tracks, to not worry about the blind alleys or describe how you had the wrong idea first, and so on. So there isn't any place to publish, in a dignified manner, what you actually did in order to get to do the work.” - Richard Phillips Feynman (1918-1988) -
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Occurrence and diversity of Mycobacterium species in PAH-contaminated soils.
CHAPTER 2 OCCURRENCE AND DIVERSITY OF FAST-GROWING MYCOBACTERIUM SPECIES IN SOILS CONTAMINATED WITH POLYCYCLIC AROMATIC HYDROCARBONS (PAHs)* * REDRAFTED AFTER: LEYS NATALIE, RYNGAERT ANNEMIE, BASTIAENS LEEN, WATTIAU PIERRE, TOP EVA, VERSTRAETE WILLY, SPRINGAEL DIRK (REVISED) OCCURRENCE GROWING
MYCOBACTERIUM
SPECIES
IN
SOILS
CONTAMINATED
WITH
AND DIVERSITY OF FASTPOLYCYCLIC
AROMATIC
HYDROCARBONS (PAHS), FEMS MICROBIOL. ECOL.
ABSTRACT Mycobacterium species using polycyclic aromatic hydrocarbons (PAH) as sole source of carbon and energy may be essential members of the bacterial populations degrading PAHs in the environment. To study the natural role and diversity of PAHdegrading Mycobacterium communities in contaminated soils, a culture-independent fingerprinting method based on PCR combined with Denaturing Gradient Gel Electrophoresis (DGGE) was developed. As so far all PAH-degrading Mycobacterium isolates could be placed in the phylogenetic branch of the ‘fast-growing’ Mycobacterium species, new PCR primers were selected to specifically target 16S rRNA genes of fast-growing Mycobacterium species. The new primer set proved to be highly selective for the target group in PCR and single-band DGGE profiles were obtained for most Mycobacterium strains tested. Strains belonging to the same species had identical DGGE fingerprints, and in most cases but not all, these fingerprints were typical for one species, allowing partial differentiation between species in a Mycobacterium population. Mycobacterium strains inoculated in soil were detected with a detection limit of 106 CFU g-1 of soil using the new primer set alone or of circa 102 CFU g-1 in a nested PCR approach combining eubacterial and the new Mycobacterium specific primers. The PCR-DGGE detection method was used to rapidly assess the Mycobacterium population structure of several PAH-contaminated soils of diverse origin and different overall contamination profiles, pollution concentrations and chemical-physical soil characteristics. In most PAH-contaminated soils well-known PAH-degrading species like M. frederiksbergense and M.
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Chapter 2
austroafricanum were detected. Interestingly, 16S rRNA genes related to M. tusciae sequences, a Mycobacterium species so far not reported in relation to biodegradation of PAHs, were detected in all soils. INTRODUCTION Polycyclic aromatic hydrocarbons (PAHs) are hazardous environmental pollutants that are found in high concentrations in the surface soil of old gas factories and wood preservation plants (Cerniglia 1992). In spite of the limited bioavailability and poor biodegradability of PAHs, different bacteria, often Mycobacterium strains, have been isolated that are able to use PAHs as sole source of carbon and energy (Guerin et al. 1988; Bastiaens 1998; Churchill et al. 1999; Bastiaens et al. 2000; Solano-Serena et al. 2000; Willumsen et al. 2001a). So far, all PAH-biodegrading Mycobacterium isolates (Guerin et al. 1988; Briglia et al. 1994; Bastiaens 1998; Churchill et al. 1999; Poelarends et al. 1999; Yagi et al. 1999; Bastiaens et al. 2000; Schrader et al. 2000; Solano-Serena et al. 2000; Willumsen et al. 2001a), have been placed in the phylogenetic branch of the ‘fast-growing Mycobacterium species’. In the Mycobacterium genus phylogenetic tree, the ‘fast-growing Mycobacterium species’ form a coherent line of descent, distinct from the more recently evolved slow-growers within which the overt pathogens are clustered (Rogall et al. 1990; Stalh et al. 1990; Pitulle et al. 1992; Tortoli 2003). The ‘fast-growing Mycobacterium species’ are a group of Mycobacterium strains, mostly of environmental origin, that are, based on growth and biochemical characteristics and infectious properties (i.e. Mycobacterium species of Bio Safety Level 1, growth within 7 days) very different from pathogenic and facultative pathogenic more slowly growing species like M. avium, M. tuberculosis, M. leprae or M. ulcerans (i.e. Mycobacterium species of Bio safety level 2 & 3, growth after more than 7 days). The diversity of fast-growing Mycobacterium species in the environment is still greatly unknown, but could be of major interest for bioremediation of PAHcontaminated soils. Therefore, methods for community analysis and monitoring of indigenous and/or inoculated fast-growing PAH-degrading Mycobacterium strains in soil are needed. Direct culture-independent detection methods are more preferable above indirect culture-dependent techniques for detection of Mycobacterium strains because (i) a large fraction of cells in soil is hard to culture or even believed to be - 62 -
Occurrence and diversity of Mycobacterium species in PAH-contaminated soils.
unculturable (Viable But Non Culturable state) (Staley et al. 1985; Amann et al. 1995; Ghezzi et al. 1999), (ii) the hydrophobic Mycobacterium cells are known to adhere strongly to organic soil particles resulting in their difficult recovery (Barry et al. 1998; Draper 1998), and (iii) Mycobacterium species, even the ‘fast-growers’, are relatively slow-growing organisms in comparison to other soil bacteria, which makes them easily overgrown by other bacteria in culture media (Allen 1998). In addition, molecular PCR-based methods had proven to be successful for the diagnosis of Mycobacterium diseases in humans (Böddinghaus et al. 1990; De Beenhouwer et al. 1995; Kox et al. 1995; Kox et al. 1997; van der Heijden et al. 1999) and fish (Talaat et al. 1997) and for the identification of environmental infection sources of Mycobacterium opportunistic pathogens such as M. avium and M. ulcerans in plants, water and soil (Schwartz et al. 1998; Mendum et al. 2000; Stinear et al. 2000). PCR amplification of variable 16S rRNA gene-fragments combined with direct analysis of amplicons by Denaturing Gradient Gel Electrophoresis (DGGE) is a commonly used technique for rapid molecular assessment of the community diversity. In all previous studies, however, the PCR primers were designed to reveal the presence of slowgrowing Mycobacterium species in the tested samples and were never combined with a method for direct community diversity analysis like DGGE. Moreover, most importantly, none of the described primer sets were specific enough to target preferentially fast-growing Mycobacterium species, belonging to non-pathogenic and non-opportunistic species. We describe in this study the development of a new set of non-degenerated primers that annealed as exclusively as possible to 16S rRNA genes of fast-growing Mycobacterium strains and that amplified a short fragment suited for DGGE a cultureindependent method for fingerprinting of only the fast-growing Mycobacterium species. The PCR-DGGE method was theoretically and practically evaluated for detection of fast-growing Mycobacterium species in soil samples and applied to examine the Mycobacterium diversity in PAH-contaminated soils.
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MATERIALS AND METHODS Bacterial strains and growth conditions. The bacterial strains used in this study are described in Table 2-1. For DNA-extraction purposes, strains other than Mycobacterium strains were cultivated in 869-broth (Mergeay et al. 1985) while Mycobacterium strains were cultivated in Middelbrook 7H9 Broth (DIFCO, Kansas City, USA). For inoculation purposes, Mycobacterium strains were cultivated in a phosphate buffered minimal liquid medium previously described (Wick et al. 2001), containing 2 g l-1 of anthracene or pyrene crystals (ACROS Organics, Fisher Scientific, Boston, USA) floating in the medium as the sole carbon and energy source. All cultures were grown in the dark on an orbital horizontal shaker at 200 rpm at a constant temperature of 30 °C. TABLE 2-1 Organism
BACTERIAL STRAINS USED IN THIS STUDY Reported pollutant degrading capacity
PHYLUM OF HIGH G+C GRAM POSITIVE BACTERIA Order of Actinomycetales Actinomyces sp. A1008 NR Actinosynnema mirum 101 NR Arthrobacter sulfureus 8-3 NR Kineospora aurantiaca A/10312 NR Microbispora rosea IMRU37485 NR Micromonosprora chalcea A0919 NR Micromonosprora chalcea A2868 NR Micromonosprora chalcea A2894 NR Planomonospora parontospora B-987 NR Promicromonospora citrea INMI 18 NR Streptomyces albus A0818 NR Streptomyces albus A1893 NR Streptomyces albus A2198 NR Streptomyces albus A3986 NR Streptomyces aureofaciens A-377 NR Streptomyces rutgersensis BJ-608 NR Streptomyces phaeofaciens T-23 NR Streptosporangium album A0958 NR Suborder Corynebacterineae Dietziaceae family Dietzia maris VM0283 diesel Dietzia maris IMV 195 NR Corynebacteriaceae family Corynebacterium glutamicum 2247 NR Tsukamurallaceae family Tsukamurella paurometabola NR Nocardiaceae family Nocardia asteroides N3 NR Nocardia coeliaca AB.4.1.b NR Pseudonocardia hydrocarbonoxydans NR Rhodococcus erythropolis ICPB 4417 NR Gordoniaceae family NR Gordonia hydrophobica 1610/1b Gordonia amarae Se 6 NR Mycobacterinaceae family Mycobacterium aichiense 5545 NR Mycobacterium alvei CR-21 NR Mycobacterium aurum 358 NR Mycobacterium vanbaalenii PYR-1 nap, phe, fan, pyr Mycobacterium austroafricanum E9789 NR Mycobacterium austroafricanum VM0456 phe Mycobacterium austroafricanum VM0450 phe Mycobacterium austroafricanum VM0451 phe, pyr, fan, ant Mycobacterium austroafricanum VM0447 phe Mycobacterium austroafricanum VM0452 phe, fan Mycobacterium austroafricanum VM0573 phe Mycobacterium chlorophenolicum PCP-1 PCP
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16S rRNA gene Accession N°
Reference or origin
NR X84447 X83409 X87110 NR NR NR NR AB028653 X83809 NR NR NR NR NR NR D44381 NR
(Heuer et al. 1997) DSM 43827T DSM 20167T DSM 43858T ATCC 21946 (Heuer et al. 1997) (Heuer et al. 1997) (Heuer et al. 1997) DSM 43869T DSM 43110T (Heuer et al. 1997) (Heuer et al. 1997) (Heuer et al. 1997) (Heuer et al. 1997) DSM 40127T DSM 40830T DSM 40367T (Heuer et al. 1997)
NR X79290
(Wattiau et al. 1999) DSM 43627T
NR
DSM 20411
AF283280
DSM 20162T
NR NR X76955 X81929
(Heuer et al. 1997) DSM 44595T DSM 43281T DSM 43066T
X87340 X80601
DSM 44015T DSM 43392T
X55598 AF023664 X55595 U30662 X93182 AF44622 AF44623 AF44624 AF44625 AF44626 AF44627 X79094
DSM 44147T DSM 44176T DSM 43999T DSM 7251 T DSM 44191T (Springael et al. unpublished) (Springael et al. unpublished) (Springael et al. unpublished) (Springael et al. unpublished) (Springael et al. unpublished) (Springael et al. unpublished) DSM 43826T
Occurrence and diversity of Mycobacterium species in PAH-contaminated soils. Mycobacterium diernhoferi SN1418 NR X55593 DSM 43524T Mycobacterium frederiksbergense FAn9 fan, phe, pyr AJ276274 DSM 44346T Mycobacterium frederiksbergense LB501T ant, phe, diesel AJ245702 (Bastiaens et al. 2000) Mycobacterium frederiksbergense VM0503 fan AF44628 (Springael et al. unpublished) Mycobacterium frederiksbergense VM0531 phe, pyr AF44629 (Springael et al. unpublished) Mycobacterium frederiksbergense VM0458 phe AF44630 (Springael et al. unpublished) Mycobacterium frederiksbergense VM0579 phe AF44631 (Springael et al. unpublished) Mycobacterium frederiksbergense VM0585 phe, fan AF44632 (Springael et al. unpublished) Mycobacterium gilvum SM35 NR X81996 DSM 44503T Mycobacterium gilvum BB1 phe, flu, pyr, fan X81891 DSM 9487 mor, prl, pip AJ012738 DSM 44238 Mycobacterium gilvum HE5 Mycobacterium gilvum LB307T phe, pyr, dib, fan, diesel AJ245703 (Bastiaens et al. 2000) Mycobacterium gilvum VM0505 ant AF44633 (Springael et al. unpublished) Mycobacterium gilvum VM0504 ant AF44634 (Springael et al. unpublished) Mycobacterium gilvum VM0552 phe, pyr AF44635 (Springael et al. unpublished) Mycobacterium gilvum VM0442 phe, fan, ant AF44636 (Springael et al. unpublished) Mycobacterium gilvum LB208 phe, pyr, fan, diesel AJ245704 (Bastiaens et al. 2000) Mycobacterium gilvum VM0583 ant AF44637 (Springael et al. unpublished) Mycobacterium hodleri EM12 fan X93184 DSM 44183T Mycobacterium komossense Ko2 NR X55591 DSM 44078T Mycobacterium neoaurum 3503 NR M29564 DSM 44074T Mycobacterium parafortuitum 311 NR X93183 DSM 43528T Mycobacterium peregrinum 6020 NR AF058712 DSM 43271T Mycobacterium petroleophilum RF002 fan, phe, pyr U90876 (Lloyd-Jones et al. unpublished) Mycobacterium vaccae VM0587 fan AF44638 (Springael et al. unpublished) Mycobacterium vaccae VM0588 fan AF44639 (Springael et al. unpublished) Mycobacterium sp. WF2 fan U90877 (Lloyd-Jones et al. unpublished) Mycobacterium sp. GP1 DBA AJ012626 (Poelarends et al. 1999) PHYLUM OF FLAVOBACTERIA oleoresins NR ATCC 33545T Flavobacterium resinovorum PHYLUM OF PROTEOBACTERIA α-subdivision Agrobacterium luteum A61 NR NR DSM 5889T Brevundimonas diminuta 342 NR AJ227778 DSM 7234T Sphingomonas chlorophenolica PCP X87161 DSM 7098T β-subdivision Ralstonia metallidurans CH34 NR Y10824 DSM 2839T Burkholderia sp. JS150 benzene derivates AF262932 DSM 8530 γ-subdivision Aeromonas enteropelogenes J11 NR X71121 DSM 6394T Acinetobacter calcoaceticus 46 NR AJ247199 DSM 30006T Pseudomonas putida nap, phe, flu, fan NR DSM 8368 δ-subdivision Desulfobacter latus AcRS2 NR AJ441315 DSM 3381T Desulfonema magnum 4be13 NR U45989 DSM 2077T Desulfobulbus rhabdoformis M16 NR U12253 DSM 8777T T = species type strain; NR = Not Reported; nap = naphthalene; fan = fluoranthene; pyr = pyrene; flu = fluorene; phe = phenanthrene; ant = anthracene; dib= dibenzothiophene; mor = morpholine; prl = pyrrolidine; pip = piperidine; PCP = pentachlorophenol; DBA = 1,2-di-bromoethane.
Soils used in this study. Soil samples were taken from different anthropogenic PAH-contaminated sites (Table 2-2). The soil texture, pH (DIN Method 38414, S4), total carbon content (TC), total inorganic carbon (TIC) content (hydrolysis method) and total organic carbon (TOC) content of each soil sample was determined (ISO-CEN EN Method 1484). The soils were chemically analyzed for the 16 PAHs legislated by the U.S. Environmental Protection Agency. PAHs were extracted through an Accelerated Solvent Extraction (ASE 200 Accelerated Solvent Extractor, Dionex Corp., Sunnyval, CA, USA) (EPA Method 3545). ASE-extracts were purified over an internal silica phase in the extraction cell (i.e. in thimble clean up) followed by an alumina column. Quantification was done by capillary gas chromatography (Carlo Erba MFC 500 with split/splitless injector) coupled to a massspectrophotometric detector (quadrupole-type, Fisons QMD 100) (EPA Method 8270). The total concentration of mineral oil present in the soil sample was determined after an ultrasonic tetrachloroethene extraction followed by a FLORISIL clean up (U.S. Silica Company, Berkeley Springs, USA) using an infrared quantification at 2925, 2958 and 3030 cm-1 (NEN Method 5733).
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Chapter 2
TABLE 2-2
Soil
Origin
SOIL SAMPLES USED IN THIS STUDY Soil type
pH
TOC (%)
PAH conc. (mg kg-1)
Oil conc. (mg kg-1)
DNA conc.* (µg g-1)
MYCO‡
Nested †
S587 Corn field (Belgium) sand 5.5 2.15 0.289 < 50 31.00 + S588 Horse pasture (Belgium) sand 6.0 2.46 0.391 < 50 18.00 + S585 Pine tree forest (Belgium) sand 5.8 3.19 0.673 < 50 31.75 + S589 Ditch in agricultural area (Belgium) sand 5.8 4.24 0.721 < 50 49.50 + S592 Vegetable garden (Belgium) sand 7.0 3.16 1.011 < 50 38.25 + S584 Compost heap (Belgium) sand 7.3 7.04 1.063 < 50 27.75 + S591 Non-paved land road (Belgium) sand 9.0 0.76 3.357 < 50 6.25 NP TB3 Coal gasification plant (Belgium) sand 8.23 1.52 14 < 50 2.65 + K3840 Gasoline station site (Denmark) Sand 8.20 0.50 20 98 2.75 + B101 Coal gasification plant (Belgium) Sand 7.00 2.63 107 70 27.25 + E6068 Gasoline station site (Denmark) Sand 7.96 9.94 258 300 5.40 + TM Coal gasification plant (Belgium) Sand 8.00 3.85 506 4600 4.75 + Barl Coal gasification plant (Germany) Gravel 8.90 4.63 1029 109 6.15 NP AndE Railway station site (Spain) Clay 8.10 2.35 3022 2700 3.40 NP * DNA recovery per g soil, mean value of 2 parallel extractions of 1 gr of soil. ‡ result of direct PCR with Mycobacterium specific primers MYCO66f and GC40-MYCO600r on soil DNA extract: + = detectable PCR product, NP = no detectable PCR product, ND = not determined. † result of nested PCR with eubacterial primers 27f and 1492r followed by Mycobacterium specific primers MYCO66f and GC40-MYCO600r on soil DNA extract: + = detectable PCR product, NP = no detectable PCR product, ND = not determined.
Design of 16S rRNA gene primer set specific for fast-growing Mycobacterium strains. Primer sequences were selected from a multiple alignment of circa 200 16S rRNA genes (Genbank, NCBI) of circa 100 fast- and 100 slow-growing Mycobacterium species constructed with the Bionumerics software (Applied Maths, Version 2.50). The alignment was further analyzed with the PLOTCON program (EMBOSS, Version 1.9.1) to identify conserved and variable gene regions. Based on the alignment, the new Mycobacterium specific primers were selected in gene regions that are conserved within the group of fast-growing Mycobacterium species but as variable as possible within slowgrowing Mycobacterium species. In addition, for optimal species differentiation in DGGE-analysis of the PCR-products (see below), the primers were selected so that they amplified a region between 200 bp and 600 bp long with high variability. The selectivity of the selected primers was evaluated by visual analysis of the primer region within the constructed alignment of Mycobacterium rrn genes, by the Sequence Match program (RDP II) (Cole et al. 2003) and by the Advanced Blast Search program (Genbank, NCBI) (Altschul et al. 1990). The best primer combination consisted of forward primer MYCO66f (5'-CATGCAAGTCGAACGGAAA-3', E. coli position 66 to 84) and reverse primer MYCO600r (5'-TGTGAGTTTTCACGAACA-3', E. coli position 600 to 583). A 40 basepair long GCclamp (Muyzer et al. 1993; Muyzer et al. 1996) was attached to the 5' end of the MYCO600r primer for DGGE analysis of the Mycobacterium amplicons. This new primer couple MYCO66f and GC40MYCO600r amplified a 538 bp sequence of the 16S rRNA gene resulting in a PCR-product of 578 bp. DNA-extraction. Genomic DNA from pure bacterial cultures was obtained as described by Belisle et al. (Belisle et al. 1998). The DNA recovery was approximately 2.7 to 27.3 µg DNA g-1 soil. For PCR purposes, the DNA-concentration was adjusted to a final concentration of 100 ng µl-1. For fast-growing Mycobacterium strains, 100 ng of template DNA corresponds to 1.2-1.9×107 cell equivalents of genomic DNA and 2.4-3.8×107 copies of PCR targets assuming a genomic molecular weight of 3.135.20×109 Daltons per cell and two 16S rRNA gene copies per genome (Bercovier et al. 1986). DNA was extracted from 1 g soil using a protocol described by Boon et al. (Boon et al. 2000). After
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ND ND ND ND ND ND ND + + + + + NP +
Occurrence and diversity of Mycobacterium species in PAH-contaminated soils.
purification over a Wizard column, the DNA concentration in the 50µl soil extract was measured photospectroscopically. To assure that the soil DNA was of PCR quality, dilution series of soil DNA extracts were tested by PCR with universal eubacterial 16S rRNA gene primer pair GC40-63f and 518r as previously described (Marchesi et al. 1998). PCR amplification of pure strain and soil DNA. The PCR protocol used with the MYCO66f and MYCO600r primer pair consisted of a short denaturation of 15 s at 95°C, followed by 50 cycles of denaturation for 3 s at 94°C, annealing for 10 s at 50°C, elongation for 30 s at 74°C, and a final extension for 2 min at 74°C. PCR was performed on Biometra or PerkinElmer Thermalcylcers. PCR mixtures contained 100 ng of pure strain DNA or dilutions of soil DNA as templates, 1 U Taq polymerase, 25 pmol of the forward primer, 25 pmol of the reverse primer, 10 nmol of each dNTP, and 1 × PCR buffer in a final volume of 50 µl. Primers designed by Cheung and Kinkle (MycF and MycR) (Cheung et al. 2001) were used in PCR as described. All primers were synthesized by Westburg (Westburg BV, Leusden, The Netherlands). The Taq polymerase, dNTPs and PCR buffer were purchased from TaKaRa (TaKaRa Ex TaqTM, TaKaRa Shuzo Co., Ltd., Biomedical Group, Japan). DGGE analysis. The PCR-products were examined on 1.5 % agarose gels (MetaPhor, BioWhittaker, Labtrade Inc., Miami, Florida, USA) and directly used for DGGE analysis on polyacrylamide gels as described previously (Muyzer et al. 1998a). Optimal denaturing conditions were defined based on the theoretical melting temperatures of amplification fragments calculated with the Melt Analysis Software (Version 1.0.1, INGENY). A 6 % polyacrylamide gel with a denaturing gradient of 40 % to 75 % (100% denaturant gels contain 7M urea and 40 % formamide) was used for DGGE-apparatus. Electrophoresis was performed at a constant voltage of 130 V for 16 h 40 m in 1 × TAE running buffer at 60 °C in the DGGE-machine (INGENYphorU-2, INGENY International BV, The Netherlands). After electrophoresis, the gels were stained with 1 × SYBR Gold nucleic acid gel stain (Molecular Probes Europe BV, Leiden, The Netherlands) and photographed under U.V. light using a Pharmacia digital camera system with Liscap Image Capture software (Image Master VDS; Liscap Image Capture, Version 1.0, Pharmacia Biotech, Cambridge, England). Photofiles were processed and analyzed with the Bionumerics software (Version 2.50, Applied Maths, Kortrijk, Belgium). Sensitivity of PCR-DGGE method. To study the sensitivity of the PCR-DGGE method, a known amount of viable Mycobacterium cells were added to white sand (a model soil matrix) or natural soil samples at different final cell densities (i.e. a 10-fold dilution series of approximately 108 to 101 cells g1
) prior to DNA-extraction. Cells were harvested from liquid cultures, washed twice and added in 100
µl aqueous suspensions to 1 g of soil. One, two or three different Mycobacterium strain (LB501T, VM0552 and DSM 43524T) were separately or simultaneously added in different cell densities to assess the effect of cell ratios on the detection sensitivity for each single strain within a Mycobacterium population. The total soil DNA was subsequently used in PCR with the MYCO-primers and PCRproducts were analyzed by DGGE. Sequence analysis of amplified 16S rRNA gene fragments. PCR products of Mycobacterium 16S rRNA genes were cloned into plasmid vector pCR2.1-TOPO using the TOPO Cloning Kit (N.V. Invitrogen SA, Merelbeke, Belgium) as described by the manufacturer. DGGE patterns of cloned fragments were compared with the fingerprints of the parent soil Mycobacterium community to identify
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Chapter 2
which signals from the community fingerprint were cloned. A 500 bp long fragment was sequenced (Westburg BV, Leusden, The Netherlands) from a selection of clone inserts with different DGGEpatterns. The sequences were analyzed with the 'Chimera Check' program (RDPII) (Cole et al. 2003) to detect possible chimeras and with ‘Blast Search’ program (Genbank, NCBI) (Altschul et al. 1990). Cloned sequences were imported into an alignment of Mycobacterium 16S rRNA genes and edited manually to remove nucleotide positions of ambiguous alignment and gaps. Sequence similarities were calculated over the 16S rRNA gene fragment between the MYCO-primers, corrected using Kimura's two-parameter algorithm to compensate for multiple nucleotide exchange and used to construct a distance-based evolutionary tree with the Neighbor-Joining algorithm (Saitou et al. 1987). The topography of the branching order within the dendrogram was evaluated by using the MaximumLikelihood and the Maximum-Parsimony character-based algorithms in parallel combined with bootstrap analysis with a round of 500 reassemblings. An out-group of the closely related genera Rhodococcus and Dietzia was included to root the tree. Nucleotide sequence accession numbers. The partial 16S rRNA gene sequences of Mycobacterium clones reported in this study are available from GenBank under accession numbers AY148197 to AY148217.
RESULTS Design of specific primers for fast-growing PAH-degrading Mycobacterium strains. A new specific primer set was designed to amplify the 16S rRNA genes of fastgrowing Mycobacterium species. Based on an alignment of approximately 200 sequences, the 16S rRNA genes of fast- and slow-growing Mycobacterium species appeared highly conserved in comparison to other bacteria, i.e., only a few short welldefined regions within the gene were found to be highly variable (data not shown). A minimum similarity of 94% over the total length of the 16S RNA gene was found for all fast-growing Mycobacterium species, making it is very difficult to select strictly group- or species- specific primers. The best possible primer combination was selected from the alignment taking into account the amplicon length and amplicon variability and the Blast and sequence Match results of both primers. The sequence of the forward primer MYCO66f (E. coli locations 66-84) was conserved in 300 rrn gene sequences of mainly fast- but also some slow-growing Mycobacterium strains of the approximately 900 Mycobacterium sequences currently available in the Genbank database (NCBI) (Table 2-3). The MYCO66f primer also
- 68 -
Occurrence and diversity of Mycobacterium species in PAH-contaminated soils.
aligned 100% with Corynebacterium, Phytoplasma, Gordonia and Propionibacterium 16S rRNA genes. The sequence of primer MYCO600r (E. coli locations 600-583), however, was 100% conserved in only 165 sequences of exclusively fast-growing environmental Mycobacterium strains, including all known PAH degrading species (Table 2-3). It clearly differed from most other 16S rRNA gene sequences from slow-growing Mycobacterium strains and non-Mycobacterium strains with 1 to 7 mismatches of the 18bp long primer region (Table 2-3). As for some slow-growing mycobacteria the mismatches with the MYCO660r primer were more concentrated to the 5’ in stead of the 3’ primer end, amplification of the 16S rRNA genes may be possible. Nevertheless, the MYCO660r primer was our best possible choice and showed more mismatches with sequences from slow-growing Mycobacterium than any other primer decribed so far. The high number of mismatches will hinder or prevent the amplification of template from most slow-growing Mycobacterium species and most species not belonging to the Mycobacterium genus. The primer couple MYCO66f and MYCO600r produced products of the appropriate size with the DNA obtained from the 40 tested Mycobacterium strains representing different fast-growing environmental and PAH degrading species (Table 2-1). Due to the risks associated with most slow-growing Mycobacterium species classified as 'Biosafety Level 2 & 3’-agents (Anonymous 1999) only fast-growing reference strains were tested. Nevertheless, PCR reactions conditions were optimized and always performed under very stringent reaction conditions (very short annealing times) to minimize amplification of 16S rRNA genes with only few mismatches located at the 5’ primer end as found in some slow-growing Mycobacterium species. No PCRproducts were obtained with DNA from strains belonging to other related genera such as Actinomyces, Arthrobacter, Dietzia, Corynebacterium, Nocardia, Sphingomonas, Burkolderia, Acinetobacter, Desulfobacter and more (Table 2-1). Moreover, non of the cloned sequences obtained from PCR products with soil DNA extracts as template, showed close relationships to any slow-growing Mycobacterium species or species not belonging to the Mycobacterium genus.
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Chapter 2
TABLE 2-3
Organism
DNA-SEQUENCE
HOMOLOGY BETWEEN THE MYCOBACTERIUM GENUS SPECIFIC PRIMERS AND THE 16S RRNA GENE SEQUENCE OF SOME REFERENCE MYCOBACTERIUM SPECIES 16S rRNA gene *
Primers† MYCO66f
MYCO600r
(E.coli 66-84) (E. coli 600-583) 5' - CATGCAAGTCGAACGGAAA - 3' 5' - TGTGAGTTTTCACGAACA - 3' Fast-growing Mycobacterium species M. aichiense X55598 M. alvei AF023664 M. aurum X55595 M. austroafricanum X93182 M. chlorophenolicum X79094 M. diernhoferi X55593 M. frederiksbergense AJ276274 M. gilvum X81996 M. hodleri X93184 M. komossense X55591 M. neoaurum M29564 M. parafortuitum X93183 M. peregrinum AF058712 M. petroleophilum RF002 U90876 M. vaccae VM0587 AF44638 M. sp. AJ012738 M. sp. WF2 U90877 Slow-growing Mycobacterium species M. gordonae X52923 M. genavense X60070 M. branderi X82234 M. leprae X55587 M. tuberculosis H37Rv NC_000962 M. ulcerans X88926 Non-Mycobacterium bacteria Gordonia terrae AB111113 AF397061 Corynebacterium sp. AF262996 Phytoplasma sp. AF500334 Rhodococcus globerulus X77779 Dietzia maris X79290 * †
----------------------------------------------------------------------------------------------------------------------------------------------------------------------------------T-region unsequenced -------------------------------------------------------------------------------------------------------------
-
-
-
-
-
------A ------------
-
-
-
-
-
-
-
-
-
-
-
----------------T--------------------------------------------------------------------------------------------
C- - - - - A - - - - - - - - - - CCCCCGA - - - - - - - - - - --------------G--C - - - - - A - - - - - - - - -NN C-----A----------C-----A-----------
----------------------------------------------------------------------T-- - - - - - - - - - - -G - - - T - -
- - C - - A - - - CACA - - CGCCG - TA - - - CACA - - CG-T G A- A - - AAACTAG -G- - - - - - - - - - - A - - - -GC -A - GA - - - - - AG- - - G-
Accession No of 16S rRNA gene sequence in the Genbank (NCBI, Rockville Pike Bethesda, USA) N = A or T or G or C, Dashes indicate homologous sequences.
Differentiation of fast-growing Mycobacterium species by DGGE-analysis. In order to examine the potential of the PCR-DGGE method based on the new primer set to differentiate between species within a Mycobacterium community, pure strain DGGE-patterns of a variety of fast-growing Mycobacterium strains were compared. Different Mycobacterium isolates belonging to the same species showed identical DGGE-patterns. This was observed for 8 tested M. austroafricanum related strains, 7 tested M. frederiksbergense related strains and 10 tested M. gilvum related strains (data not shown). Usually different species showed different DGGE fingerprints (Figure 2-1). For example, the PCR-products obtained for the M. frederiksbergense strain (lane 16) migrated clearly differently from the products from M. austroafricanum (lanes 3 & 4) or M. gilvum (lane 9) strains.
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1, M. vaccae VM0587; 2, M. diernhoferi DSM43524T; 3, M. austroafricanum DSM7251; 4, M. austroafricanum DSM44191T; 5, M. aurum DSM43999T; 6, M. chlorophenolicum DSM43826T; 7, Mycobacterium sp. DSM44238; 8, M. hodleri DSM44183T; 9, M. gilvum DSM44503T; 10, M. petroleophilum RF002; 11, Mycobacterium sp. WF2; 12, Mycobacterium sp. GP1; 13, M. peregrinum DSM43271T; 14, Mycobacterium sp. VM0585; 15, Mycobacterium sp. VM0579; 16, M. frederiksbergense DSM44346T; 17, M. neoaurum DSM44074T; 18, M. alvei DSM44176T; 19, M. komossense DSM44078T; 20, M. parafortuitum DSM43528T; 21, M. aichiense DSM44147T.
Occurrence and diversity of Mycobacterium species in PAH-contaminated soils.
*
**
*
FIGURE 2-1 MYCOBACTERIUM SPECIES DIFFERENTIATION BY DGGE-ANALYSIS OF MYCOBACTERIUM DNA-FRAGMENTS AMPLIFIED WITH THE MYCOBACTERIUM GENUS SPECIFIC PRIMER PAIR MYCO66F AND GC40-MYCO600R. DGGE-fingerprints of strains were compared by Bionumerics software based on co-running standard (not shown). The symbol * indicates multiple band DGGE patterns for single strains with arrows indicating cloned and sequenced bands.
However, in some cases the differences in migration between different species were minor, due to the strong conservation of the Mycobacterium 16S rRNA genes. The amplified rrn gene fragments of species such as M. spaghni (not shown), M. hodleri (lane 8) and M. gilvum (lane 9) with a similarity of 99 to 100% and could not be differentiated by DGGE. For comparison, the amplicons of primer pairs, i.e., MYCO66f and GC40-MYCO600r and MycF and GC40-MycR (Cheung et al. 2001), were analyzed under identical DGGE-conditions. Although a different fragment of the 16S rRNA gene was amplified, PCR-DGGE with both primer sets resulted in the same species differentiation degrees for all tested Mycobacterium strains (data not shown). The species specific DGGE fingerprints usually displayed one single band. However, some strains revealed extra bands in comparison to other members of the same species (Figure 2-1). For example strains DSM 7251 (lane 3) and VM0450 (data not shown) both showed a similar pattern with 5 bands, strains VM0579 (lane 15), VM0531 (data not shown) and VM0503 (data not shown) displayed 3 bands and - 71 -
Chapter 2
VM0585 (lane 14) even 10 bands. The same number of DGGE-bands per strain were obtained when using an other Mycobacterium specific primer the MycF and MycR primers designed by Cheung and Kinkle on the test strains (Cheung et al. 2001) (data not shown). The sequences of these multiple bands for one strain showed only very limited variation, i.e., 98-99% similarity or max. 4 point mutations and deletions over the 500 bp sequenced fragment, and chimera analysis (Cole et al. 2003) of sequenced bands was negative. Multiple random cloning attempts of these PCR products in total or of single excised bands to selectively clone more fainter or higher located additional bands to further investigate their sequence divergence were not successful. Sensitivity of the Mycobacterium specific PCR-DGGE protocol. To examine the sensitivity of the PCR-DGGE protocol to detect Mycobacterium strains in soil, a known decreasing amount of Mycobacterium sp. LB501T cells were added prior to DNA-extraction to white sand and different PAH-contaminated soil samples (Table 2-2). For some contaminated soils there was a clear inhibitory effect of the soil matrix on the PCR amplification, so the soil DNA template was diluted 1:10 or 1:100 prior to PCR. In addition, white sand was used as a model soil matrix. In a PCR reaction with primer pair MYCO66f and GC40-MYCO600r, LB501T cells could generally be detected until a minimum cell concentration of 106-108 cells per of soil. The same order of detection limit was obtained when using an other Mycobacterium specific primer the MycF and MycR primers designed by Cheung and Kinkle on the same DNA-extracts (Cheung et al. 2001) (data not shown). Similar results were also obtained when two or three different Mycobacterium strains (LB501T and VM0552 or LB501T, VM0552 and DSM 43524T) were simultaneously added to white sand in equal cell concentrations (ratios 1:1 or 1:1:1 ratio), i.e., all strains were detected equally well until a concentration of 106 to 107 CFU g-1 soil (Figure 2-2). Attempts to improve significantly the detection limit by optimizing the DNA extraction and purification protocol or reducing length of the GC-clamp were not successful (data not shown). Only via a nested PCR approach, combining a first PCR with universal eubacterial primers and a second PCR with the MYCO-primers, the detection limit could be lowered to circa 102 cells per gram of soil (data not shown).
- 72 -
Occurrence and diversity of Mycobacterium species in PAH-contaminated soils.
To assess the impact of unequal cell concentration ratios on the detection sensitivity, different concentrations of VM0552 or VM0552 and DSM 43524T (1:1) cells were added to white sand in the presence of a constant concentration of circa 108 CFU g-1 of LB501T. Declining cell amounts of VM0552 and DSM 43524T could be detected in the presence of 108 CFU g-1 LB501T cells until a concentration of 106 CFU g-1 (Figure 2-2).
A
VM0552:LB501T in 1:1
10
8
10
6
10
4
10
DSM43524: VM0552:LB501T in 1:1:1
2
107 105 103 101
*
*
LB501T VM0552 DSM43524
B
VM0552:LB501T
DSM43524: VM0552:LB501T
1:1 1:2 1:4 1:10 1:100 1:10000
1:1:1 1:1:2 1:1:4 1:1:10 1:1:100 1:1:10000
*10
108 108 108 108
108 108
*
107
107
108 5x107 2x107 107
106 104
107
5x106 2x106
106
105
103
107
5x106 2x106
106
105
103
7
107
107
107
LB501T VM0552 DSM43524
FIGURE 2-2 DETECTION EFFICIENCY OF THE PCR-DGGE METHOD USING GENUS SPECIFIC PRIMER PAIR MYCO66F AND GC40-MYCO600R
MYCOBACTERIUM
(A) PCR-DGGE detection of the simultaneously added M. frederiksbergense LB501T, M. gilvum VM0552 and M. diernhoferi DSM 43524T at cell concentration of approximately 108, 107, 106, 105, 104, 103 and 102 CFU g-1 in white sand. (B) PCR-DGGE detection of M. gilvum VM0552 and M. diernhoferi DSM 43524T added in declining cell concentration together with a constant cell density of M. frederiksbergense LB501T of 108 CFU g-1. The symbol * indicates the detection limit on the figure.
- 73 -
Chapter 2
Diversity analysis of fast-growing Mycobacterium populations in PAH contaminated soil. The MYCO-primer PCR-DGGE method was used to assess the Mycobacterium diversity in a set of PAH-contaminated soil samples with different contamination records (Table 2-2). Indigenous Mycobacterium cells could be detected in 6 of the 7 tested non-contaminated soils and in 6 of the 7 tested PAH-contaminated soils (Figure 2-3). Despite the high concentrations of PAHs, PCR-DGGE fingerprinting with universal eubacterial primers revealed, however, the presence of a heterogeneous bacterial soil community in soils negative in PCR with the MYCO-primer set. Moreover, DNA-extracts from parallel samples with added Mycobacterium cells produced good PCR products with the MYCO-primer set, omitting PCR inhibition as possible cause for the negative PCR results with the MYCO-primer set. The DGGE fingerprints of the Mycobacterium community in the positive soil samples were complex, comprising several bands for each soil (Figure 2-3). The 16S rRNA gene PCR products from 4 samples were randomly cloned and clones representing different bands from one soil fingerprint were sequenced. All sequences exhibited high levels of similarity to 16S rRNA gene sequences of Mycobacterium strains (Table 2-4) and could be placed within the phylogenetic group of fast-growing Mycobacterium species (Figure 2-4), confirming the specificity of the primer set. All positive soils revealed clone sequences most similar to the 16S rRNA genes of a variety of exclusively fast-growing Mycobacterium species. Sequences that were closely related to 16S rRNA gene sequences of known PAH- and oil-degrading species such as M. frederiksbergense, M. austroafricanum and M. petroleophilum were detected in soils K3840 and B101 but not in soil TM. However, the dominant number of 16S rRNA gene soil clone sequences from all 3 PAH-contaminated soils showed highest sequence similarity with the 16S rRNA gene of the relatively unknown M. tusciae. M. tusciae related sequences were not retrieved from the noncontaminated soil. Interestingly, the different soil fingerprints revealed bands closely related to M. tusciae but with different migration profiles. In addition, the cloned sequences showed a relatively high variation in similarity scores (from 99 to 95 %). The M. tusciae sequences isolated in this study grouped with other unidentified Mycobacterium sequences cloned from DNA from petroleum contaminated soils found by Cheung and Kinkle using the MycF and MycR primer pair (Cheung et al. 2001) (Figure 2-4). Besides the M. tuscia strains, only strains of the M. monacence - 74 -
Occurrence and diversity of Mycobacterium species in PAH-contaminated soils.
(AF107039) species, a fast-growing species represented by a type strain of clinical origin and an atypical isolate (U46146), seem to be closely linked to this cluster.
Low PAH con. S589
Mix
TB3
K3840
S589/4 S589/3
B101
E6068
TM
K3840/3
S589/6 S589/1
M. frederiksb. M. sphagni M. sp.WF2 M. gilvum
High PAH con.
K3840/4 K3840/8 K3840/2 K3840/1 K3840/6
S589/7 S589/5
M. vaccae M. diernhoferi
K3840/7
TM/8 TM/9 TM/2 TM/1 B101/2 B101/4 B101/7 B101/6 B101/3 B101/5
TM/3 TM/6 TM/4 TM/5 TM/7
B101/1
S589/2
FIGURE 2-3 PCR-DGGE FINGERPRINT OF INDIGENOUS MYCOBACTERIUM CELLS IN SOIL SAMPLES USING MYCOBACTERIUM SPECIFIC PRIMERS MYCO66F AND GC40-MYCO600R Cloned ‘bands’ are indicated within the soil fingerprint based on the comparison of migration profiles of pure clones and the soil profile.
TABLE 2-4
ORIGIN Soil S589
CLONED 16S RRNA GENE SEQUENCES RETRIEVED FROM SAMPLES CLONES
ACCESSION N°
S589/1 / S589/2 / S589/3 / S589/4 / S589/5 / S589/6 / S589/7 / Soil K3840 K3840/1 AY148216 K3840/2 AY148200 K3840/3 AY148207 K3840/4 AY148201 K3840/6 AY148214 K3840/7 AY148197 K3840/8 AY148210 Soil B101 B101/1 AY148202 B101/2 AY148208 B101/3 AY148198 B101/4 AY148204 B101/5 AY148212 B101/6 AY148217 B101/7 AY148215 Soil TM TM/1 AY148211 TM/2 AY148209 TM/3 AY148203 TM/4 AY148205 TM/5 AY148206 TM/6 AY148196 TM/7 AY148213 TM/8 AY148199 * known oil or PAH degrading bacterium
PAH-POLLUTED
NEAREST MATCH IN BLAST ANALYSIS (ACCESSION N°) M. alvei DSM 44176T (AF023664) M. moriokaense DSM 44221T (AJ429044) M. lacus (AF406783) M. lacus (AF406783) (Y15709) M. anthracenicum * M. margeritense 1336 (AJ011335) (X93184) M. hodleri DSM 44183T Mycobacterium sp. M0183 (AF055332) * Mycobacterium sp. HXN1500 (AJ457057) M. tusciae DSM 44338T (AF058299) (AJ245702) M. frederiksbergense LB501T * T* (X93182) M. austroafricanum DSM 44191 M. gadium ATCC 27726 (X55594) (AF058299) M. tusciae DSM 44338T Mycobacterium sp. JKD2385 (AF221088) M. isoniacini INA-I (X80768) M. holsaticum1406 (AJ310467) (AF058299) M. tusciae DSM 44338T M. septicum HX1900 (AJ457056) (U90876) M. petroleophilum RF002 * Mycobacterium. sp. WF2 * (U90877) M. tusciae DSM 44338T (AF058299) M. tusciae DSM 44338T (AF058299) (AF058299) M. tusciae DSM 44338T T (AF058299) M. tusciae DSM 44338 M. tusciae DSM 44338T (AF058299) M. moriokaense MCR07 (AF058299) M. septicum HX1900 (AJ457056) (AF130308) M. tusciae DSM 44338T
- 75 -
SOIL
SIMILARITY 97 % 94 % 97 % 97 % 98 % 95 % 97 % 99 % 98 % 95 % 98 % 98 % 98 % 99 % 98 % 97 % 97 % 96 % 99 % 98 % 97 % 99 % 99 % 97 % 97 % 98 % 97 % 98 % 99 %
0
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
Chapter 2
X 77 77 9
R . gl ob eru lu s NC IMB12 3 15
X 79 29 0
D i. ma ris D SM4 36 72
U 90 87 6
M. pe trol eo ph ilu m RF0 02
A Y14 82 17
M. sp . cl on e B1 01 /6
A J01 27 38
M. sp . D SM4 42 38
A F0 55 33 2
M. sp . M0 18 3
A Y14 82 16
M. sp . cl on e K3 84 0/1
A Y14 82 13
M. sp . cl on e TM/7
A J45 70 56
M. se pticu m HX N1 90 0
A Y14 82 12
M. sp . cl on e B1 01 /5
A J45 70 55
M. se pticu m HX N5 00
A F0 23 66 4
M. al ve i DSM 44 17 6T
A F2 21 08 8
M. sp . JD K2 38 55
A Y14 82 02
M. sp . cl on e B1 01 /1
X 55 59 1
M. kom oss en se D SM4 40 78 T
X 55 59 8
M. ai chi en se D SM4 41 47 T
A J01 26 26
M. sp . GP1
A J45 70 57
M. fred eri ksb erg en se H XN 15 00
A J27 62 74
M. fred eri ksb erg en se D SM4 43 46 T
A F2 94 74 9
M. sp . cl on e KT-2 6
A F2 94 75 0
M. sp . cl on e KT-2 7
A Y14 82 00
M. sp . cl on e K3 84 0/2
A Y14 82 01
M. sp . cl on e K3 84 0/4
X 55 59 3
M. di ern ho feri D SM4 35 24 T
M 29 56 4
M. ne oa ur um D SM44 07 4 T
A J31 04 67
M. ho lsa ticu m 14 06
A Y14 81 98
M. sp . cl on e B1 01 /3
X 55 59 4
M. ga di um ATC C2 77 26
A Y14 81 97
M. sp . cl on e K3 84 0/7
A F0 58 71 2
M. pe re gri nu m D SM4 32 7 1T
X 93 18 4
M. ho dl eri DSM 44 18 3T
X 80 76 8
M. iso ni ac ini IN A-I
A Y14 82 08
M. sp . cl on e B1 01 /2
A Y14 81 96
M. sp . cl on e TM/6
X 93 03 3
M. mo rio kae nse MCR O 7
A B02 84 83
M. sp . TA5
A F3 30 69 5
M. sp . cl on e C C-4
A F2 94 74 2
M. sp . cl on e AC -1
X 93 18 3
M. pa ra fo rtui tu m DSM 43 52 8T
X 81 99 6
M. gi lvu m DS M44 50 3T
X 55 59 5
M. au ru m D SM4 39 99 T
A F2 20 43 1
M. sp . cl on e KT-1 9
X 79 09 4
M. ch lor op he no lic um D SM4 38 26 T
A Y14 82 14
M. sp . cl on e K3 84 0/6
X 93 18 2
M. au stro afric an um D SM44 19 1T
U 90 87 7
M. sp . WF2
A Y14 82 15
M. sp . cl on e B1 01 /7
A Y14 82 06
M. sp . cl on e TM/5
A Y14 82 04
M. sp . cl on e B1 01 /4
A Y14 82 05
M. sp . cl on e TM/4
A Y14 82 07
M. sp . cl on e K3 84 0/3
A Y14 82 03
M. sp . cl on e TM/3
A F0 58 29 9
M. tusc iae DSM 44 33 8T
13
A Y14 82 11
M. sp . cl on e TM/1
7
A Y14 82 09
M. sp . cl on e TM/2
A Y14 81 99
M. sp . cl on e TM/8
A Y14 82 10
M. sp . cl on e K3 84 0/8
A F2 20 42 8
M. sp . cl on e C T- 11
95
A F2 20 43 0
M. sp . cl on e JT-1 5
96
A F1 07 03 9
M. mo na cen se B9 -2 1-1 78
A F2 20 42 7
M. sp . cl on e AT-3
U 46 14 6
M. sp .
A F2 20 42 9
M. sp . cl on e C T- 12
A F2 20 43 2
M. sp . cl on e KT-2 2
A F2 20 43 3
M. sp . cl on e KT-2 3
A F2 94 74 5
M. sp . cl on e C T- 24
A F2 94 74 6
M. sp . cl on e C T- 25
A F2 94 74 8
M. sp . cl on e JC -14
A F2 94 74 4
M. sp . cl on e C T- 10
A F2 94 74 7
M. sp . cl on e JC -13
A F2 94 74 3
M. sp . cl on e C T- 9
100
100
100
77
10 0 10 0 93 76
99
92
100
100
96 75
75
100
85
100 85
100 8498 59
85
100 99 83
10 0
96
82
100 100
90
91
91
100 99
77
83
100
87
88
100
90
93
100 93
96 89
94
94
100
94
97
100 98
92
100
10 0
66
48
100
45
100
- 76 -
Occurrence and diversity of Mycobacterium species in PAH-contaminated soils.
FIGURE 2-4
PHYLOGENETIC ANALYSIS OF 16S RRNA GENE SEQUENCES DETECTED IN SOIL SAMPLES USING MYCOBACTERIUM SPECIFIC PRIMERS MYCO66F AND GC40-MYCO600R.
Phylogenetic positioning of Mycobacterium 16S rRNA gene sequences detected in soil within the Mycobacterium genus. The evolutionary tree was generated by the Neighbor-joining method based on Kimara 2-parameter corrected similarity percentages of 538 bp 16S rRNA gene fragments between the MYCO primers and branching orders were evaluated using the Maximum-Parsimony algorithm. The topology was also evaluated by bootstrap analysis (500 reassemblings) and percentages of bootstrap support are indicated at the branch points, with values above 70% indicating reliable branches. An outgroup of the closely related genera Rhodococcus and Dietzia was included to root the tree. The bar at the top indicates the estimated evolutionary distance, i.e., 1% indicating an average of 1 nucleotide substitution at any nucleotide position per 100 nucleotide positions. The evolutionary distance between two strains is the sum of the branch lengths between them.
DISCUSSION We developed a specific PCR-DGGE method to rapidly assess the diversity of or monitor fast-growing Mycobacterium species in PAH-contaminated soil. From our test results, it could be concluded that the new MYCO66f and MYCO600r primer set, although theoretically also targeting a few slow-growing species, amplifies preferentially 16S rRNA gene of fast-growing Mycobacterium species from environmental samples. It is the first primer set that was specifically developed for the detection fast-growing Mycobacterium species only. All previously described primer combinations were specific for the total Mycobacterium genus (Böddinghaus et al. 1990; Kox et al. 1995; Kox et al. 1997; Talaat et al. 1997; Schwartz et al. 1998; van der Heijden et al. 1999; Mendum et al. 2000; Stinear et al. 2000; Cheung et al. 2001) or
exclusively
for
pathogenic
and
facultative
pathogenic
slow-growing
Mycobacterium species (De Beenhouwer et al. 1995; Schwartz et al. 1998). Even the Mycobacterium genus specific 16S rRNA gene primer set that was parallel designed during the course of this work and used in combination with Temperature Gradient Gel Electrophoresis (TGGE) for diversity analysis of indigenous Mycobacterium populations in petroleum-contaminated soil, was targeting both fast- and slowgrowing species (Cheung et al. 2001). Forward primers MYCO66f (this study) and MycF (Cheung et al. 2001) were similarly conserved in 16S rRNA genes of Mycobacterium strains but reverse primer MYCO600r (this study) was far more specific for fast-growing Mycobacterium species than the MycR primer (Cheung et al. 2001).
- 77 -
Chapter 2
DGGE analysis of gene fragments amplified with the MYCO-primer set could differentiate between most fast-growing Mycobacterium species including all important PAH-degrading species. For some very closely related species the DGGE fingerprints overlapped. For comparison, DGGE-differentiation of the same species was also limited when using PCR-products obtained with the Mycobacterium specific primer set of Cheung and Kinkle (Cheung et al. 2001). Similarly, TGGE-analysis of 16S rRNA gene fragments could not discriminate between several species of Burkolderia (Falcão Salles et al. 2002) and Bifidobacterium (Satokari et al. 2001) or other Gram-positive coryneform soil bacteria such as Arthrobacter and Nocardoides (Felske et al. 1999), due to the high levels of conservation of the amplified 16S rRNA gene fragments. It is clear that the practical resolution limit of the DGGE-technique is at the species or genus level or intermediate between the two, depending on the gene conservation level within the taxonomic group that is under investigation. Most Mycobacterium strains were characterized by a single band DGGE-fingerprint, only a few showed satellite bands. The sequences of multiple bands for one strain displayed very high similarity. Others also reported multiple-band DGGE-patterns for pure strains of species such as Paenibacillus polymyxa (Nübel et al. 1996), Burkholderia cepacia (Falcão Salles et al. 2002) and Bifidobacterium adolescentis (Satokari et al. 2001) due to the presence of multiple 16S rRNA gene copies with sequence heterogeneity. Southern blotting and DNA-DNA hybridization revealed, however, that the indicated Mycobacterium strains contained a maximum of two copies of the rrn operon (Chapter 3), theoretically producing maximum 2 different bands on a DGGE-gel. This is consistent with other reports and 3 total genome sequences from Mycobacterium strains from clinical origin showing that slowgrowing and fast-growing Mycobacterium strains possess only 1 or 2 copies of rrn genes respectively (Bercovier et al. 1986; Klappenbach et al. 2001). It is therefore unlikely that the obtained multiple band DGGE-fingerprints are due to the presence of more than 2 rRNA gene copies. Reaction conditions such as these used for the MYCO-primers, i.e., mixtures with high concentrations of very homologous 16S rRNA gene templates, may generate chimeras (Cole et al. 2003). However, it is unlikely that random chimera formation caused these multiple-band DGGE patterns since different DNA-preparations of the same strain always produced the same multiple-band DGGE-fingerprints. Moreover, cloned sequences of single strain - 78 -
Occurrence and diversity of Mycobacterium species in PAH-contaminated soils.
multiple bands were negative in chimera analysis and showed limited variations. The slower migrating fainter satellite bands (located higher in the gel) may have been due to heteroduplex formation between 2 copies of 16S rRNA genes with 1 species. It is known that specific heteroduplexes can be detected in denaturing gels when 16S rRNA genes of Mycobacterium strains with at least 1 to 2 % difference in nucleotide sequence anneal (Waléria-Aleixo et al. 2000). In a mix of PCR products of 2 different genes, one or two heteroduplex bands additional to the one or two homoduplex bands may appear in the denaturing gel fingerprint. Alternatively, culture impurities could be the cause of multiple band patterns, although several culture DNA-extracts were tested and severe precautions for culture or template cross contamination were taken. Nevertheless, the possible appearance of multiple-band DGGE-fingerprints for single Mycobacterium strains may indicate that the number of different Mycobacterium species actually present in the soil community may well be lower than the number of bands in a soil fingerprint. With the one-step PCR-DGGE method using only the MYCO66f and GC40MYCO600r primer pair, we could detect Mycobacterium cells in white sand and several PAH-contaminated soils until a minimum cell concentration of circa 106 cells per gram soil. None of the other Mycobacterium genus specific primers sets developed in the past, report on the Mycobacterium abundance or detection limit in environmental samples for comparison (Schwartz et al. 1998; Mendum et al. 2000; Stinear et al. 2000; Cheung et al. 2001) but we found a similar detection limit when using the primer set developed by Cheung and Kinkle (Cheung et al. 2001). The value reported for a similar direct PCR-DGGE detection method for Burkholderia species in soil was only slightly lower (detection limit 5 × 105 CFU g-1) (Falcão Salles et al. 2002), although more copies of the rrn genes are present in the target bacterium (6 rrn copies in Burkholderia while 2 in Mycobacterium strains). Nested PCR using eubacterial primers in the first round and the MYCO-primers with GC-clamp in the second round could drastically lower the detection limit to circa 102 cells per gram soil. Another approach could be the use of the more abundant rRNA molecules instead of the rRNA gene as targets for the MYCO-primers in a reverse transcription PCR (RT-PCR) protocol. In a RT-PCR set up using the MYCO-primer set, Mycobacterium sp. LB501T was detected at a concentration as low as 102 active cells gr-1 soil (Hendrickx et al., unpublished data). Based on our results, all fast-growing - 79 -
Chapter 2
Mycobacterium strains are also expected to be detected equally well in a mixed Mycobacterium community. No differences in lysis or in preferential amplification based on primer homology would be expected between the different fast-growing PAH-degrading Mycobacterium species and there are indications that all fast-growing Mycobacterium species contain the same amount (2 copies) of 16S rRNA gene copies (see above). Finally, the new specific PCR-DGGE method was successfully used to assess the diversity of fast-growing Mycobacterium species in a set of different contaminated and non-contaminated soils. Fast-growing Mycobacterium species could be detected in 6 out of the 7 tested PAH-contaminated soil samples and in 6 out of 7 tested noncontaminated soil samples. These results clearly suggest a wide distribution of fastgrowing Mycobacterium species in the environment. For the 5 positive soils containing PAH-concentrations of 500 mg kg-1 or lower, there was no clear correlation between Mycobacterium biodiversity (assessed by the number of bands in the Mycobacterium DGGE fingerprints) and the PAH-concentration of the soils. Cheung and Kinkle (Cheung et al. 2001), however, observed a clear negative correlation between the number of Mycobacterium bands in the TGGE fingerprint (2 to 18) and the PAH-content (0.07 to 473 mg kg-1) of a soil. They suggested that the Mycobacterium diversity in more contaminated soil samples may be reduced by the toxicity of the PAHs or by a natural selection process in which the Mycobacterium community was enriched and finally dominated by one or a few populations more adapted to the higher PAH concentrations. The fact that there was no clear reduction in Mycobacterium diversity with increasing PAH-concentrations in our test soils could thus be due to the lack of toxicity or the lack of community adaptation in our soil samples with PAH-concentrations of 500 mg kg-1. With the one-step PCR-DGGE method, no Mycobacterium PCR signal could be detected in the 2 soils with highest PAH-concentrations of circa 1000 and 3000 mg kg-1. Although we did not perform toxicity tests, we have no indications that the very low or undetectable concentration of Mycobacterium species in heavily contaminated soils was due to toxicity as all tested soils were characterized by a rather complex total bacterial community DGGEfingerprints obtained with universal primers and contained a total concentration of 106 to 108 cultivatable bacterial cells independent of the PAH-concentration (Vanbroekhoven et al. unpublished). The absence or lower concentrations of fast- 80 -
Occurrence and diversity of Mycobacterium species in PAH-contaminated soils.
growing Mycobacterium species in very heavily PAH-contaminated soils (1000 – 3000 mg kg-1) may indicate the natural selection of fast-growing Mycobacterium species in PAH-polluted soil enriched in poorly bioavailable and highly recalcitrant higher molecular PAHs. In soils containing high concentration of more easily degradable 3-ring PAHs Mycobacterium species might be out competed by more quickly growing PAH-degrading bacteria such as Sphingomonas or Pseudomonas species. Mycobacterium species are maybe better adapted to harsh oligotrophic soil conditions as they have a low maintenance energy demand and make use of several PAH bioavailability-enhancing mechanisms such as high-affinity uptake systems and adhesion to the substrate (Wick et al. 2001; Wick et al. 2002a). Sequence analysis of the indigenous soil DGGE-fingerprints of 4 of the soils revealed the presence of strains closely related to known fast-growing PAH-degrading isolates from the M. frederiksbergense and M. austroafricanum species in 2 of the PAHcontaminated soils. None of the detected sequences seemed to originate from strains related to M. gilvum, another well-known PAH-degrading species (Boldrin et al. 1993; Bastiaens 1998; Bastiaens et al. 2000). However, sequences related to the M. tusciae species were repeatedly detected in all PAH-contaminated soils, originating from different countries and different industrial sites, but not in the non-contaminated soil. These results may indicate an important role for M. tusciae and/or related species in PAH-degradation processes in soil. The type strain of this species, M. tusciae strain DSM 44338T, is a facultative pathogenic clinical isolate from a sick child (Tortoli et al. 1999), but also 2 unpublished vinyl chloride degrading soil isolates (Coleman et al., unpublished) were recently identified as members of the M. tusciae species. Based on the different DGGE-bands and the varying similarity of the clones in our study to the type strain, we may have detected still unknown species relatively closely related to M. tusciae. Although, the M. tusciae species has never been isolated or detected before in PAH-contaminated soil, the M. tusciae sequences isolated in this study grouped with other unidentified Mycobacterium sequences cloned from DNA from petroleum contaminated soils found by Cheung and Kinkle using the MycF and MycR primer pair (Cheung et al. 2001).
- 81 -
Chapter 2
The repeated detection of Mycobacterium cells in soils with low PAH-concentrations and low in organic carbon support the natural importance of fast-growing Mycobacterium species in PAH-polluted soil. The developed PCR-DGGE detection system is an important tool to specifically monitor the natural abundance, the diversity and the dynamics of these bacteria in soil for optimization of bioremediation. The developed primer pair may also be useful in a RT-PCR approach with ribosomal RNA soil extracts to analyze the diversity of the actively PAH-degrading population of fastgrowing Mycobacterium species. Eventually these primers could be combined with primers developed for the detection of messenger RNA of the well conserved PAH catabolic genes in Mycobacterium strains (Khan et al. 2001; Krivobok et al. 2003). This will add to a better understanding of the role of Mycobacterium species in the biodegradation of PAHs in the environment. ACKNOWLEDGEMENTS This work was supported by the European Commission, through the contracts BIO4CT97-2015 and QLRT-1999-00326. We thank E.M.H. Wellington for providing bacterial strains and S. Schioetz-Hansen, J. Amor and J. Vandenberghe for providing soil samples.
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Occurrence of M. frederiksbergense in PAH-contaminated soils.
CHAPTER 3 MYCOBACTERIUM FREDERIKSBERGENSE, A MYCOBACTERIUM SPECIES SPECIALISED IN POLYCYCLIC AROMATIC HYDROCARBON (PAH) DEGRADATION, IS UBIQUITOUS IN PAH-CONTAMINATED SOILS
* REDRAFTED
AFTER:
LEYS NATALIE, RYNGAERT ANNEMIE, BASTIAENS LEEN,
VAN
CANNEYT MARK,
SWINGS JEAN, TOP EVA, VERSTRAETE WILLY, SPRINGAEL DIRK (SUBMITTTED)
MYCOBACTERIUM
FREDERIKSBERGENSE, HYDROCARBON
A
(PAH)
MYCOBACTERIUM
SPECIES
SPECIALISED
IN
POLYCYCLIC
DEGRADATION, IS UBIQUITOUS IN PAH-CONTAMINATED SOILS,
AROMATIC
ENVIRON.
MICROBIOL.
ABSTRACT The fast-growing Mycobacterium species M. frederiksbergense seems to be specialized in colonizing PAH-contaminated soils and to play a role in PAH biodegradation in those soils. This species contained up to now only PAH-degrading isolates derived from PAH-contaminated soil. Based on 16S rRNA gene sequence similarity,
FAME
analysis
and
ribotyping,
7
additional
PAH-degrading
Mycobacterium isolates were identified as M. frederiksbergense strains. A cultureindependent PCR based protocol was developed to detect and monitor specifically M. frederiksbergense strains in contaminated soils. A specific primer pair, MYCOFf and MYCOFr, was developed targeting exclusively the 16S rRNA gene sequence of M. frederiksbergense strains. PCR reactions using the new primer sets on Mycobacterium and non-Mycobacterium template DNA demonstrated that the MYCOFf and MYCOFr primer set was highly selective for M. frederiksbergense. Using the new primer set, M. frederiksbergense strains inoculated in soil could be detected at a cell concentration of 104 cells per g soil via direct PCR and subsequent DNA-DNA hybridization of the PCR products or at a cell concentration of maximum 102 cells per g soil via a nested PCR approach consisting of a first PCR reaction using universal primers and a second reaction using the MYCOFf/MYCOFr primer set. PCR on soil DNA extracts revealed the presence of M. frederiksbergens-like strains in PAH-
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Chapter 3
contaminated soils of diverse origin and different overall contamination profiles, PAH-concentrations and physico-chemical soil characteristics. These results indicate a role for M. frederiksbergense strains in the biological degradation of PAHs in PAHcontaminated environments. INTRODUCTION Polycyclic aromatic hydrocarbons (PAHs) are hazardous environmental pollutants that are found in high concentrations in the surface soil of sites housing former gas production plants and wood preservation plants through spills of coal tar and/or creosote (Cerniglia 1992). Biodegradation is considered as the main route of natural PAH removal in soil. Therefore, bioremediation can be used as an alternative for chemical or physical remediation techniques to clean up PAH contaminated sites. Many of the PAH-degrading bacteria isolated from PAH-contaminated soils have been identified as Mycobacterium (Bastiaens 1998; Bastiaens et al. 2000; Springael et al. unpublished). Those PAH-degrading Mycobacterium isolates were, based on 16S rRNA gene sequence, often assigned to the species M. frederiksbergense (Willumsen et al. 2001a), M. gilvum (Boldrin et al. 1993; Bastiaens et al. 2000; Vila et al. 2001; Gauthier et al. 2003), M. austroafricanum (Bogan et al. 2003), M. vanbaalenii (Heitkamp et al. 1988a; Godvidaswami et al. 1995; Wang et al. 1995; Khan et al. 2001; Moody et al. 2001; Khan et al. 2002), M. hodleri (Kleespies et al. 1996), M. flavescens (Dean-Ross et al. 1996), M. anthracenicum (Wang et al. unpublished) and M. chelonae (Kanaly et al. 2000a; Kanaly et al. 2000b; Kanaly et al. 2002). M. frederiksbergense is of particular interest for PAH-bioremediation purposes as up to now, this species seems to be specialized in PAH-degradation. The M. frederiksbergense species is taxonomically represented by only one strain, i.e., PAHdegrading type strain M. frederiksbergense FAn9T (DSM 44346T) (Willumsen et al. 2001a), but several PAH-degraders Mycobacterium isolates from diverse origin including Mycobacterium sp. strain LB501T, seem to be closely related to strain FAn9T based on 16S rRNA gene sequence (Bastiaens 1998; Bastiaens et al. 2000; Springael et al. unpublished). Strain LB501T displays several adaptations to the low bioavailability of PAH-compounds, i.e., it makes close contact with and strongly adhere to the PAH crystal surface, uses high-affinity uptake systems and has very low maintenance energy requirements (Wick et al. 2001; Wick et al. 2002a). Moreover, it
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Occurrence of M. frederiksbergense in PAH-contaminated soils.
has a unique cell wall with an extreme high negative surface charge and high hydrophobicity, possibly playing a role in the modes of interaction of the strain with PAH-compounds (Bastiaens 1998; Bastiaens et al. 2000; Wick et al. 2001; Wick et al. 2002a; Wick et al. 2002b; Wick et al. 2003b). Furthermore, in contrast with other PAH-degrading Mycobacterium strains, several of those apparently FAn9T related strains are able to degrade PAHs at low temperature (e.g. 4-12°C) and in the presence of high salt concentrations (I = 0.500) (Leys et al. unpublished data). Direct detection of species such as M. frederiksbergense by PCR or other molecular techniques is therefore of importance for further study of the distribution and role of this species in PAH-contaminated environments. Primer sets exist that target all Mycobacterium species (Schwartz et al. 1998; Mendum et al. 2000; Cheung et al. 2001) or the group of the fast-growing Mycobacterium species (Chapter 2) for their detection by PCR in environmental samples. In 1996, 16S rRNA gene based primer set have been reported for the PCR detection of M. chlorphenolicum PCP1 (DSM 43826T) (Briglia et al. 1996) and PAH-degrading strains M. vanbaalenii PYR1 (DSM 7251T) and Mycobacterium sp. strain PAH135 (Wang et al. 1996). Specific detection methods for the PAH-degrading M. frederiksbergense FAn9T (DSM 44346T), however, do not exist. In this study, we report the development of a new species-specific detection method for selective monitoring of PAH-degrading strains of the M. frederiksbergense species. The new developed 16S rRNA gene primer set was used to assess the natural presence of M. frederiksbergense strains in a set of PAH-contaminated soils.
MATERIALS AND METHODS Bacterial strains and growth conditions. The bacterial strains used in this study are described in Table 3-1. For DNA-extraction purposes, Mycobacterium strains were cultivated in Middelbrook 7H9 Broth (DIFCO) while all other strains were cultivated in 869-broth (Mergeay et al. 1985). All cultures were grown in the dark on an orbital horizontal shaker at 200 rpm at a constant temperature of 30 °C.
TABLE 3-1 Strain
STRAINS USED IN THIS STUDY 16S rRNA gene accession N°
Reference or origin
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MYCOF†
Chapter 3
MYCOBACTERIUM STRAINS M. aichiense 5545 T X55598 DSM 44147T M. alvei CR-21 T AF023664 DSM 44176T T T M. aurum 358 X55595 DSM 43999 M. austroafricanum E9789 T X93182 DSM 44191T M. chlorophenolicum PCP-1 T X79094 DSM 43826T T T M. diernhoferi SN1418 X55593 DSM 43524 M. frederiksbergense Fan9 T * AJ276274 DSM 44346T + M. gilvum BB1* X81891 DSM 9487 M. gilvum SM35 T X81996 DSM 44503T M. hodleri EM12 T * X93184 DSM 44183T M. komossense Ko2 T X55591 DSM 44078T M. neoaurum 3503 T M29564 DSM 44074T T T M. parafortuitum 311 X93183 DSM 43528 M. peregrinum 6020 T AF058712 DSM 43271T M. vanbaalenii PYR-1 T * U30662 DSM 7251T Mycobacterium sp. GP1 AJ012626 (Poelarends et al. 1999) Mycobacterium sp. HE5 AJ012738 DSM 44238 Mycobacterium sp. LB208* AJ245704 (Bastiaens et al. 2000) Mycobacterium sp. LB307T* AJ245703 (Bastiaens et al. 2000) Mycobacterium sp. LB501T* AJ245702 (Bastiaens et al. 2000) + Mycobacterium sp. RF002* U90876 (Lloyd-Jones et al. unpublished) Mycobacterium sp. VM0442* AF44636 (Springael et al. unpublished) Mycobacterium sp. VM0447* AF44625 (Springael et al. unpublished) Mycobacterium sp. VM0450* AF44623 (Springael et al. unpublished) Mycobacterium sp. VM0451* AF44624 (Springael et al. unpublished) Mycobacterium sp. VM0452* AF44626 (Springael et al. unpublished) Mycobacterium sp. VM0456* AF44622 (Springael et al. unpublished) Mycobacterium sp. VM0458* AF44630 (Springael et al. unpublished) + Mycobacterium sp. VM0503* AF44628 (Springael et al. unpublished) + Mycobacterium sp. VM0504* AF44634 (Springael et al. unpublished) Mycobacterium sp. VM0505* AF44633 (Springael et al. unpublished) Mycobacterium sp. VM0531* AF44629 (Springael et al. unpublished) + Mycobacterium sp. VM0552* AF44635 (Springael et al. unpublished) Mycobacterium sp. VM0573* AF44627 (Springael et al. unpublished) Mycobacterium sp. VM0579* AF44631 (Springael et al. unpublished) + Mycobacterium sp. VM0583* AF44637 (Springael et al. unpublished) Mycobacterium sp. VM0585* AF44632 (Springael et al. unpublished) + Mycobacterium sp. VM0587* AF44638 (Springael et al. unpublished) Mycobacterium sp. VM0588* AF44639 (Springael et al. unpublished) Mycobacterium sp. WF2* U90877 (Lloyd-Jones et al. unpublished) OTHER GENERA Rhodococcus erythropolis ICPB 4417 T X81929 DSM 43066T Nocardia asteroides N3 NR (Springael et al. unpublished) Dietzia maris VM0283 NR (Wattiau et al. 1999) Dietzia maris IMV 195 T X79290 DSM 43627T Actinosynnema mirum 101 T X84447 DSM 43827T Arthrobacter sulfureus 8-3 T X83409 DSM 20167T Planomonospora parontospora B-987 T AB028653 DSM 43869T T T Promicromonospora citrea INMI 18 X83809 DSM 43110 Streptomyces aureofaciens A-377 T NR DSM 40127T Streptomyces rutgersensis BJ-608 T NR DSM 40830T T T Streptomyces phaeofaciens T-23 D44381 DSM 40367 Brevundimonas diminuta 342 T AJ227778 DSM 7234T Sphingomonas chlorophenolicum T X87161 DSM 7098T Pseudomonas putida NR DSM 8368 * PAH degrading strains † result of PCR with primers MYCOFf and MYCOFr: + = detectable PCR product, - = no detectable PCR product.
Soils used in this study. The soil samples used in this study were taken from different anthropogenic PAH-contaminated sites (Table 3-2). Chemical properties of the soil samples were analyzed as described previously (Chapter 2).
TABLE 3-2
SOIL SAMPLES USED IN THIS STUDY
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Occurrence of M. frederiksbergense in PAH-contaminated soils.
Soil
Origin
Soil type
pH
TOC (%)
PAH conc. (mg kg-1)
Oil conc. (mg kg-1)
DNA conc.* (µg g-1)
MYCO† PCR, cells g-1
K3840 Gasoline station site (Denmark) Sand 8.20 0.50 20 98 2.75 + , 109 B101 Coal gasification plant (Belgium) Sand 7.00 2.63 107 70 27.25 + , 108 E6068 Gasoline station site (Denmark) Sand 7.96 9.94 258 300 5.40 +, 107 TM Coal gasification plant (Belgium) Sand 8.00 3.85 506 4600 4.75 + , 108 Barl Coal gasification plant (Germany) Gravel 8.90 4.63 1029 109 6.15 NP AndE Railway station site (Spain) Clay 8.10 2.35 3022 2700 3.40 +,
