Biotechniques for Air Pollution Control : proceedings of the II International Congress on Biotechniques for Air Pollution Control, A Coruña, Spain, October 3-5, 2007

3 Biotechniques for Air Pollution Control Proceedings of the 2nd International Congress on Biotechniques for Air Pollu...

Author: Christian Kennes | María del Carmen Veiga Barbazán

82 downloads 917 Views 9MB Size Report

This content was uploaded by our users and we assume good faith they have the permission to share this book. If you own the copyright to this book and it is wrongfully on our website, we offer a simple DMCA procedure to remove your content from our site. Start by pressing the button below!
Report copyright / DMCA form

DOWNLOAD PDF

3

Biotechniques for Air Pollution Control Proceedings of the 2nd International Congress on Biotechniques for Air Pollution Control, A Coruña, Spain, October 3-5, 2007

Christian Kennes and María C. Veiga (eds.)

A Coruña 2007

Universidade da Coruña Servizo de Publicacións

4

Biotechniques for Air Pollution Control Christian Kennes, María C. Veiga (eds.) A Coruña, 2007 Universidade da Coruña, Servizo de Publicacións Cursos_Congresos_Simposios, nº. 92 616 páxinas. 17 x 24 cm. Índice: páxinas 5-12 ISBN: 978-84-9749-258-4 Depósito legal: C 3157-2007 Materia: 504: Ciencias do medio ambiente. 66: Tecnoloxía química

Edición: Universidade da Coruña, Servizo de Publicacións, http://www.udc.es/publicaciones © Os autores © Universidade da Coruña Distribución: Galicia: CONSORCIO EDITORIAL GALEGO. Estrada da Estación 70-A, 36818, A Portela. Redondela (Pontevedra). Tel. 986 405 051. Fax: 986 404 935. Correo electrónico: [email protected] España: BREOGÁN. C/ Lanuza, 11. 28022, Madrid. Tel. 91 725 90 72. Fax: 91 713 06 31. Correo electrónico: [email protected]. Web: http://www.breogan.org Imprime: Lugami Artes Gráficas

Reservados todos os dereitos. Nin a totalidade nin parte deste libro pode reproducirse ou transmitirse por ningún procedemento electrónico ou mecánico, incluíndo fotocopia, gravación magnética ou calquera almacenamento de información e sistema de recuperación, sen o permiso previo e por escrito das persoas titulares do copyright.

5

Contents

ODOUR CONTROL IMPROVING ODOUR MANAGEMENT AND ABATEMENT PERFORMANCE USING OLFACTORY GC-MS GAVIN PARCSI AND RICHARD M. STUETZ ...........................................................

15

A STATISTICAL PERSPECTIVE ON BIOFILTER PERFORMANCE IN RELATION TO THE MAIN PROCESS PARAMETERS AND CHARACTERISTICS OF UNTREATED FLOWS ANTON PHILIP VAN HARREVELD ........................................................................

29

REFURBISHMENT OF AN ODOUR COLLECTION AND BIOFILTER TREATMENT SYSTEM AT A MUNICIPAL SOLID WASTE COMPOSTING FACILITY IN PERTH, WESTERN AUSTRALIA TERRY J. SCHULZ AND STUART MCALL .............................................................

41

ODOUR AND H2S DEGRADATION IN A FULL SCALE BIOFILTER WITH A MINERAL BASED ORGANIC COATED FILTER MEDIA FRANZ-BERND FRECHEN, WOLFRAM FRANKE AND BJÖRN SCHOLL ......................

53

HIGH THROUGHPUT BIOFILTRATION FOR ODOUR CONTROL AT WATER PURIFICATION PLANT VITALY ZHUKOV, ANDREY VEPRITZKY, LEONID MITIN AND VLADIMIR POPOV ......

59

CONTINUOUS MONITORING OF ODOURS AT A BIOFILTER OUTLET SELENA SIRONI, LAURA CAPELLI, PAOLO CÉNTOLA, RENATO DEL ROSSO AND MASSIMILIANO IL GRANDE ................................................................................

67

REMOVAL OF ODOUR AND AMMONIA IN VENTILATION AIR FROM GROWING-FINISHING PIG UNITS USING VERTICAL BIOFILTERS ANDERS LEEGAARD RIIS ...................................................................................

77

MEMBRANE INLET MASS SPECTROMETRY (MIMS) AS A TOOL FOR EVALUATING BIOLOGICAL AIR FILTERS IN AGRICULTURE ANDERS FEILBERG ............................................................................................

85

6

MODELING AND TECHNOLOGICAL ASPECTS MODELING OF A FUNGAL BIOFILTER FOR THE ABATEMENT OF HYDROPHOBIC VOCs ALBERTO VERGARA-FERNÁNDEZ AND SERGIO REVAH .........................................

95

GASOLINE BIOFILTRATION: AN ANALYTIC MODEL ANDREW M. GERRARD, MARTIN HALECKY AND JAN PACA .................................

107

MODELING OF BIOMASS ACCUMULATION AND FILTER BED STRUCTURE CHANGE IN BIOFILTERS FOR GASEOUS TOLUENE REMOVAL JINYING XI, HONG-YING HU AND CAN WANG ...................................................

115

MODELLING THE ADSORPTION OF STYRENE AND ACETONE ON ACTIVATED CARBON AND PERLITE BEDS ANDREW M. GERRARD, SEBASTIAN MOLLENCAMP, KEHINDE MAKINDE, JAN PACA AND ONDREJ MISIACZEK ...................................................................................

125

CO-TREATMENT OF BENZENE AND TOLUENE VAPOURS IN A BIOFILTER: A FACTORIAL DESIGN APPROACH ELDON R. RENE AND T. SWAMINATHAN .............................................................

133

MATHEMATICAL MODELING AND SIMULATION OF VOLATILE REDUCED SULFUR COMPOUNDS OXIDATION IN BIOTRICKLING FILTERS G. AROCA, M. CÁCERES, S. PRADO, C. SÁNCHEZ AND R. SAN MARTÍN .............

143

ARTIFICIAL NEURAL NETWORK BASED MODEL FOR EVALUATING PERFORMANCE OF IMMOBILIZED CELL BIOFILTER ELDON R. RENE, JUNG HOON KIM AND HUNG SUCK PARK ................................

153

BIODESULPHURISATION AND BIOGAS TREATMENT DEVELOPMENT OF A FAMILY OF LARGE-SCALE BIOTECHNOLOGICAL PROCESSES TO DESULPHURISE INDUSTRIAL GASSES ALBERT J.H. JANSSEN, ROBIN VAN LEERDAM, PIM VAN DEN BOSCH, ERIK VAN ZESSEN, GIJS VAN HEERINGEN AND CEES BUISMAN ...........................................

167

7

STUDY OF A DESULFURIZATION PROCESS TO CONVERT DIBENZOTHIOPHENE TO 2-HYDROXYBIPHENYL BY RHODOCOCCUS RHODOCHROUS NRRL (B-2149) A.B. SOARES JÚNIOR, Y. K. P. GURGEL, B.M.E. CHAGAS, T. B. DOMINGOS, G.R. MACEDO AND E. S. SANTOS .....................................................................

185

CONTROL OF METHANE EMISSIONS ISSUING FROM LANDFILLS: THE CANADIAN CASE JOSIANE NIKIEMA AND MICHÈLE HEITZ .............................................................

193

DESULFURISATION OF BIOGAS BY BIOFILTRATION DIANA RAMÍREZ-SÁENZ AND INÉS GARCÍA PEÑA ...............................................

201

AN INNOVATIVE BIOTRICKLING FILTER FOR H2S REMOVAL FROM BIOGAS LAURA BAILÓN ALLEGUE .................................................................................

215

REMOVAL OF INORGANIC POLLUTANTS REMOVAL OF AMMONIA BY IMMOBILIZED NITROSOMONAS EUROPAEA IN A BIOTRICKLING FILTER PACKED WITH POLYURETHANE FOAM MARTÍN RAMÍREZ, JOSÉ MANUEL GÓMEZ AND DOMINGO CANTERO ...................

227

STUDY OF NH3 REMOVAL BY GAS-PHASE BIOFILTRATION: EFFECTS OF SHOCK LOADS AND WATERING RATE ON BIOFILTER PERFORMANCE GUILLERMO BAQUERIZO, JUAN PEDRO MAESTRE, XAVIER GAMISANS, DAVID GABRIEL AND JAVIER LAFUENTE ........................................................................

239

HIGH H2S CONCENTRATIONS ABATEMENT IN A BIOTRICKLING FILTER: START-UP AT CONTROLLED pH AND EFFECT OF THE EBRT AND O2/H2S SUPPLY RATIO MARC FORTUNY, MARC A. DESHUSSES, XAVIER GAMISANS, CARLES CASAS, DAVID GABRIEL AND JAVIER LAFUENTE .............................................................

251

8

AMMONIA TRANSFORMATION IN A BIOTRICKLING AIR FILTER LARS PETER NIELSEN, MARIE LOUISE NIELSEN, MATHIAS ANDERSEN AND ANDERS M. NIELSEN .....................................................................................................

263

REMOVAL OF HYDROGEN SULFIDE USING UPFLOW AND DOWNFLOW BIOFILTERS WONGPUN LIMPASENI AND NATTAPOL RATTANAMUK ..........................................

273

PROPOSING A NEW BATCH METHOD FOR ASSESSMENT OF BIOLOGICAL ACTIVITY IN H2S DEGRADING BIOTRICKLING FILTERS L. OTEGI AND L. LARREA .................................................................................

283

EFFECT OF VARIOUS FACTORS TO AMMONIA BIODEGRADATION BY TWO STAGE BIOFILTRATION SYSTEM SILVIJA STRIKAUSKA, DZIDRA ZARI A, OLGA MUTERE, ULDIS VIESTURS AND ANDREJS .............................................................................................

293

MEMBRANE BIOREACTORS REMOVAL OF DIMETHYL SULFIDE IN A THERMOPHILIC MEMBRANE BIOREACTOR MUNKHTSETSEG LUVSANJAMBA, AMIT KUMAR AND HERMAN VAN LANGENHOVE ......................................................................

305

BIOLOGICAL WASTE GAS PURIFICATION USING MEMBRANES: OPPORTUNITIES AND CHALLENGES N.J.R. KRAAKMAN, N. VAN RAS, D. LLEWELLYN, D. STARMANS AND P. REBEYRE ......................................................................................................

313

TREATMENT OF THE CONFINED AIR OF A SPACECRAFT CABIN AUDREY RAMIS, CÉCILE HORT, SABINE SOCHARD AND VINCENT PLATEL ............

323

GAS-PHASE TOLUENE BIODEGRADATION BY BURKHOLDERIA VIETNAMIENSIS G4 IN A BIOFILM MEMBRANE REACTOR AMIT KUMAR, JO DEWULF, MUNKHTSETSEG LUVSANJAMBA AND HERMAN VAN LANGENHOVE ......................................................................

335

9

VOC REMOVAL IN CONVENTIONAL AND BIOTRICKLING FILTERS TREATMENT OF GAS PHASE STYRENE IN A BIOFILTER UNDER STEADY-STATE CONDITIONS ELDON R. RENE, MARÍA C. VEIGA AND CHRISTIAN KENNES ..............................

347

DEGRADATION OF SOLVENT MIXTURE VAPORS IN A BIOTRICKLING FILTER REACTOR: IMPACT OF HYDROPHILIC COMPONENTS LOADING AND LOADING RELEASE DYNAMIC JAN PACA, ONDREJ MISIACZEK, MARTIN HALECKY AND KIM JONES ...................

357

PERFORMANCE OF PEAT BIOFILTERS TREATING ETHYL ACETATE AND TOLUENE MIXTURES UNDER NON-STEADY-STATE CONDITIONS F.J. ÁLVAREZ-HORNOS, C. GABALDÓN, V. MARTÍNEZ-SORIA, P. MARZAL AND J.M. PENYA-ROJA ......................................................................................

367

CHARACTERIZATION OF A BIOTRICKLING FILTER TREATING METHANOL VAPOURS A. ÁVALOS RAMÍREZ, J.P. JONES AND M. HEITZ ...............................................

379

PERFORMANCE EVALUATION OF FUNGAL BIOFILTERS PACKED WITH PALL RING, LAVA ROCK, AND PERLITE FOR α-PINENE REMOVAL YAOMIN JIN, MARÍA C. VEIGA AND CHRISTIAN KENNES ....................................

385

STYRENE DEGRADATION IN PERLITE BIOFILTER: THE OVERALL PERFORMANCE CHARACTERISTICS AND DYNAMIC RESPONSE M. HALECKY, J. PACA, A. M. GERRARD AND C.R. SOCCOL ..............................

397

BIODEGRADATION OF METHYL ETHYL KETONE AND METHYL ISOPROPYL KETONE IN A COMPOSITE BEAD BIOFILTER WU-CHUNG CHAN AND KANG-HONG PENG .......................................................

405

REMOVAL OF A MIXTURE OF OXYGENATED VOCs IN A BIOTRICKLING FILTER F. J. ÁLVAREZ-HORNOS, C. GABALDÓN, V. MARTÍNEZ-SORIA, P. MARZAL, J.M. PENYA-ROJA AND F. SEMPERE ............................................................................

415

10

EMERGING BIOREACTOR TECHNOLOGIES SOLID-LIQUID TWO-PHASE PARTITIONING BIOREACTORS FOR THE TREATMENT OF GAS-PHASE VOCs ANDREW J. DAUGULIS, JENNIFER V. LITTLEJOHNS AND NEAL G. BOUDREAU .......

427

MICROBIAL SOLVENT REGENERATION IN BIOTREATMENT OF AIR CONTAMINATED BY STYRENE ERIC DUMONT AND YVES ANDRÈS .....................................................................

439

LONG-TERM STABILITY OF PSEUDOMONAS PUTIDA CULTURES DURING THE OFF-GAS TREATMENT OF TOLUENE RAÚL MUÑOZ, ANTONIA ROJAS, LUIS FELIPE DÍAZ, SERGIO BORDEL AND SANTIAGO VILLAVERDE ..............................................................................

447

DEVELOPMENT OF A NOVEL BIOSCRUBBING PROCESS FOR COMPLETE TREATMENT OF NOX FROM FLUE GASES SANJEEV S.R. ARJUNAGI AND LIGY PHILIP .........................................................

457

DICHLOROMETHANE REMOVAL USING MIXED CULTURES IN A BIOFILTER AND A MODIFIED ROTATING BIOLOGICAL CONTACTOR– START UP STUDIES R. RAVI, LIGY PHILIP AND T. SWAMINATHAN .....................................................

469

BEHAVIOUR AND OPTIMIZATION OF A NOVEL MONOLITH BIOREACTOR FOR WASTE GAS TREATMENT YAOMIN JIN, MARÍA C. VEIGA AND CHRISTIAN KENNES ....................................

477

BIODEGRADATION OF BTXS AND SUBSTRATE INTERACTIONS IN A BIOACTIVE FOAM REACTOR JIHYEON SONG AND SHONG-GYU SHIN .............................................................

487

CHARACTERIZATION AND PERFORMANCE EVALUATION OF A TWOPHASE PARTITIONING BIOREACTOR FOR VOLATILES ORGANIC COMPOUNDS TREATMENT IN OFF-GAS JEAN-MARC ALDRIC AND PHILIPPE THONART .....................................................

495

11

REMOVAL OF DICHLOROMETHANE FROM WASTE GASES USING A FIXED-BED BIOTRICKLING FILTER AND A CONTINUOUS STIRRED TANK BIOREACTOR LAURA BAILÓN, YOLANDA DOPICO, MARCELL NIKOLAUSZ, MATTHIAS KÄSTNER, MARÍA C. VEIGA AND CHRISTIAN KENNES .........................................................

505

MICROBIAL ECOLOGY AND BIOFILMS DEVELOPMENT OF A RELIABLE EXTRACTION METHOD FOR THE RECOVERY OF TOTAL GENOMIC DNA FROM WOODCHIP COLONIZING BIOFILM INVOLVED IN GAS BIOFILTRATION LÉA CABROL, LUC MALHAUTIER, JANICK ROCHER, FRANCK POLY, XAVIER LE ROUX, MARC JOVIC, ANNE-SOPHIE LEPEUPLE AND JEAN-LOUIS FANLO ..............

517

FT-IR CHARACTERIZATION OF BIOFILMS FORMED ON ENGINEERED BIOFILTRATION MEDIA TREATING VOLATILE ORGANIC EMISSIONS FOR THE FOREST PRODUCTS INDUSTRY KIM JONES, MILI KHILNANI, ANAND KARRE, SERGIO SANTOS AND JAN PACA ....

529

MONITORING AND CHARACTERIZATION OF BACTERIAL POPULATIONS OF TWO BIOLOGICAL AIR FILTERS DURING THE START UP PHASE M. JOVIC, L. CABROL, F. DUCRAY, R. GAGNEUX AND A. S. LEPEUPLE ..............

539

BIOFILTER RESPONSE TO BIOMASS REACTIVATION FOR VOC TREATMENT A. ELÍAS, A. BARONA, G. GALLASTEGI, M. LARRAÑAGA AND M. FERNÁNDEZ ...

551

A COMPARATIVE STUDY OF THE CHARACTERISTICS AND PHYSICAL BEHAVIOUR OF DIFFERENT PACKING MATERIALS COMMONLY USED IN BIOFILTRATION ANTONI D. DORADO, XAVIER GAMISANS, DAVID GABRIEL AND JAVIER LAFUENTE ......................................................................................

561

SUITABILITY OF DUST AND BIOAEROSOLS FROM A PIG STABLE AS INOCULUM FOR BIOLOGICAL AIR FILTERS ANJA KRISTIANSEN, PER HALKJÆR NIELSEN AND JEPPE LUND NIELSEN ..............

573

12

PILOT-SCALE AND INDUSTRIAL APPLICATIONS BIOFILTRATION OF BITUMEN VAPOURS–OPERATIONAL ASPECTS MATTHIEU GIRARD, JEAN-LOUIS FANLO, NICOLAS TURGEON, GERARDO BUELNA AND PAUL LESSARD ..........................................................................................

587

COMPARISON OF THREE PILOT PLANTS FILLED WITH ORGANIC MATERIALS FOR THE TREATMENT OF AIR POLLUTANTS FROM A COMPOSTING PLANT SÉBASTIEN BASSIVIÈRE, FLORENCE DUCRAY AND CHRISTOPHE RENNER ..............

597

BIOFILTRATION SYSTEMS FOR THE TREATMENT OF WASTE GAS FROM INDUSTRIAL PLANTS IAN PHILLIPS ....................................................................................................

605

13

ODOUR CONTROL

This page intentionally left blank

15

Improving odour management and abatement performance using Olfactory GC-MS GAVIN PARCSI AND RICHARD M STUETZ Centre for Water and Waste Technology, School of Civil and Environmental Engineering, University of New South Wales, Sydney, NSW, 2033, Australia

ABSTRACT The measurement of odorous emissions is usually assessed either as odour concentrations (OC) by dilution olfactometry or by the chemical analysis of the odorous compounds such as hydrogen sulphide or the separation of complex gas mixture using analytical instrumentation such as gas chromatography. These techniques either provide information on the perceived effect of the emission (olfactory) or characterise the odours in terms of their chemical composition (analytical) but provide limited information on the relationship between odour impact and the chemical composition. The integration of chemical and olfactory techniques using olfactory-gas chromatography allows for the correlation of chemical and sensory measurements via the coupling of an olfactory port to a GC. The incorporation of mass spectrometry (GC-MS-O) enables individual odorants to be separated, identified and characterised according to their intensity and character. GC-MS-O analysis of emissions from poultry sheds has shown that samples vary in terms of their chemical compositions (i.e. different odorants profiles) as well as the different intensities measured and demonstrates the potential benefits that GC-MS-O analysis can offer in identifying key chemical markers for odour management in terms of odorant removal (i.e. receptor impact) and abatement loading due to chemical saturation.

1 INTRODUCTION Complaints due to odour annoyance have become a major issue for intensive livestock, waste management and wastewater treatment operators as the repeated release of unpleasant odours from these facilities can constitute a nuisance to a local population (Gostelow et al., 2003). This impact has become more significant with the expansion

15

16

GAVIN PARCSI AND RICHARD M. STUETZ

of suburbia and the associated rural encroachment, resulting in residential and commercial properties becoming in closer proximity to these facilities than in the past. Traditionally, odour management has been maintained by the use of buffer distances between industry and receptors or by the installation of odour abatement systems that either collect and dispersion the emission or treat the emission to acceptable level to limit receptor impact. Conventional odour abatement systems include chemical scrubbers, biofilters, bioscrubbers and biotrickling filters. Often these systems do not deliver the expected reduction in odour emissions and / or meet their original design specifications in terms of removal efficiency, resulting in the emission of odorous compounds to local receptors leading to odour complaints. The cause of these process failures is often due to inadequate characterisation of the emission source in terms of odour composition and mass loading. A secondary effect of inadequate odour composition information is the ineffective evaluation of odour control systems performance during its operation. The design and optimisation of odour management and abatement systems is based on an understanding of the emissions present in the facilities with background environmental conditions. Typical odours emitted from intensive livestock, waste management and wastewater treatment facilities usually consist of a wide range of odorants; the essential components being hydrogen sulphide (H2S), methanethiol, dimethyl sulfide, aldehydes and some ketones. Most odour abatement designs are based on the use of one or two key odorants such as H2S, reduced sulphur compounds and / or VOC to determine the loading capacity for the system. This approach often doesn’t adequately account for the actual composition and individual concentrations that vary over time and rank the emission differences in terms of odorant removal (i.e. receptor impact) and abatement loading due to chemical saturation. The measurement of odours can either be assessed as odour concentration units (OU) by dilution olfactometry (using the CEN or equivalent national standard for dilution olfactometry) or analytical techniques such as the use of surrogates chemical markers (like H2S) or the chemical analysis of odorous mixtures by chromatographic techniques such as gas chromatography coupled with mass spectrometry (GC-MS) for quantification of individual compounds (Gostelow et al., 2001). Sensory measurements employ human panels (Figure 1) to characterise the odours in terms of their perceived effect but give no information regarding composition, whereas analytical measurements characterise odours in terms of their chemical composition but give little information as to their sensory impact. Current chemical methods for odour monitoring can include field sampling and laboratory analysis (Figure 2) of gaseous emissions such as H2S, volatile organic carbon (VOC), and ammonia measurements and the continuous in-situ monitoring of H2S, volatile organic carbon (VOC), and ammonia.

IMPROVING ODOUR MANAGEMENT AND ABATEMENT PERFORMANCE

Figure 1. Olfactory analysis of odour samples.

Figure 2. Field and continuous monitoring of H2S.

17

18


More recently the integration of chemical and olfactory techniques has been applied to odour analysis to allow the correlation of chemical and sensory measurements via the coupling of an olfactory port to gas chromatograph-mass spectroscopy (GCMS-O). GC-MS-O (Figure 3) allows individual odorants to be separated and identified individually as well as allowing the odour contribution for each compound to be characterised. The olfactory detection port (ODP) consists of a nose cone where panellists perceive the separated odorous compounds by continuously sniffing the GC column effluent and characterises it in terms of intensity and an odour description. The end of GC column is split into two streams via a column splitter (Figure 4) that directs column effluent to the MS and ODP via heated transfer lines.

Figure 3. Olfactory-GC-MS showing odour detection port (ODP) on right.

Olfactory-GC and Olfactory-GC-MS is well established in other science fields such as food aroma’s and taste and odours in drinking water but has limited application to environmental odour analysis until recently. In drinking water taste and odours (or off-flavours) monitoring GC-MS-O analysis has been successfully applied to the characterisation of common off-flavours such as geosmin and MIB (Hochereau and Bruchet, 2004) and has been used to produce odour wheels (Figure 5), which relate the odour descriptors to the chemical composition of odorants (Suffet et al., 1999).


Figure 4. Column splitter directing column effluent to the MS and ODP.

Figure 5. Example of odour wheels for off-flavours in drinking water (Suffet et al., 1999).

19

20


GC-MS-O applications for the assessment of environmental odours has mainly focused on characterising changes in composition of odorous emissions from various agricultural and waste management operations such as swine finishing and poultry sheds and dairy facilities. Studies (Kai and Schäfer, 2004; Wright et al., 2005; Parcsi et al., 2007) have shown that emissions from different intensive livestock operations comprise different chemicals and odorants and that some species that gave an olfactometry response did not always correspond to a response from any other detector, conversely some compounds with large detector responses gave little or no olfactometry response. Additionally speculation is often made as to the identity of the compound based upon it odour characteristic and associated compounds within the matrix. This paper will describe the application of using olfactory-GC-MS for the characterisation of non-methane volatile organic compounds (NMVOC) emissions from tunnel ventilated broiler sheds in Australia and discuss how this technique can be more broadly applied to improve the design and optimisation of odour abatement performance through improved understanding of variations in the composition of odorous emissions in terms of receptor impact (i.e. different odorant profiles) and chemical loading on odour abatement systems.

2 MATERIALS AND METHODS The results that are presented here focus on odorous samples from two tunnel ventilated broiler sheds in Queensland and Victoria, Australia. Samples were collected on sorbent tubes containing either a Tenax TA sorbent (for n-C7 to n-C30 compounds) or a Carbotrap 300 sorbent (a blend of Carbopack C, Carbopack B and Carbosieve SIII for ethane to n-C20) (Markes International, UK), using calibrated sampling pumps. The sample volumes were recorded for each tube to allow for relative quantification. The use of different sorbents ensures that the compounds identified in subsequent analysis accurately represent the suite of compounds that are being emitted from the poultry sheds. The analytes were thermally desorbed from the sorbents and refocused within the cold trap of the thermal desorber (Markes Unity, Markes International, UK). Sample analysis was performed using a GC-MS (Agilent 6890N GC, 5973NMSD, Agilent Technologies) coupled to an Olfactory Detection Port (ODP2 Gerstel GmbH & Co., Germany) (Figure 3). The compounds were identified using gas chromatographic separation and mass selective detection with a HP-5MS capillary column (30m x 0.25mm x 0.25μm Film Thickness, Agilent Technologies). The flow rate of the gas chromatograph was maintained at a constant pressure using helium as the carrier gas. The oven was temperature programmed for a total run time of 44.00min, (50°C for 2 min, 5.00°C/min to 250°C hold for 2 min) this provided adequate separation of the eluting compounds. The mass selective detector was operating in continuous


21

scan mode (50 – 550 m/z) for GC-MS only analysis. The mass spectra were recorded using the Agilent ChemStation software and analysed offline using the Enhanced Data Analysis package (Agilent Technologies). The identification of the volatile organic compounds relied upon the matching of the acquired mass spectra with the ChemStation data bases (NIST02 and Wiley275). Identification of the compounds present within the matrix yielded a large number of different classes of compounds including aromatics, sulphur containing organic species, nitrogen containing species, aldehydes, ketones, alcohols, terpines and other general hydrocarbons. GC-MS-O analysis involved splitting the gas-chromatograph effluent between the mass selective detector and an Olfactory Detection Port. The scan range of the mass selective detector was increased at this stage to provide a more reliable match to the spectral databases (35 – 550 m/z). The mass spectra were recorded using the Agilent ChemStation software and the odour chromatograms were recorded using the Gerstel ODP Recorder software. Analysis was performed offline using the Agilent ChemStation Data Analysis software. To optimise the use of the panellist as an odour detector the split between the MSD and ODP was initially set at 1:1, before being refined to 2:3 (MSD:ODP), these split ratios were calculated using the Gerstal Column Calculator (Gerstel GmbH & Co., Germany.) These calculations were based on a column flow of 1.6mL.min-1 for the carrier gas Helium with an initial temperature of 50°C with the flow programmed to be constant flow as the temperature increases. In addition to the collection and analysis of NMVOCs, odour bags were collected onsite and analysed at local laboratories (as determined by dynamic dilution olfactometry as per CEN standards), this allows for the comparison to be drawn between the NMVOC emissions and the odour concentrations.

3 RESULTS A range of odour samples were collected during four sampling programs from two tunnel ventilated broiler sheds in Queensland and Victoria, Australia in order to characterisation of NMVOC emissions over the chicken growing out cycle (typically 9 weeks). 3.1 GC-MS ANALYSIS GC-MS analysis revealed that there was a marked variation in not only the abundance of species that were present during the grow-out cycle, but also the species that were present varied throughout the cycle. Figure 6 shows two typical total ion chromatograms (TIC’s) from one of the sampling locations. Both samples were collected under identical conditions, on the same day, from the same duty fan on the same shed at the same ventilation rate. The only difference was the sample volume,

22


the Carbotrap300 was 2.91L and the Tenax TA was 3L. The compounds labelled are A – 1-butanol, B – dimethyl disulphide, C – toluene, D – styrene, E – N-butyl-1butanamine, F – 4-ethyl-decane, G – butylated hydroxytoluene (BHT). Table 1 shows a list of predominant NMVOC compounds that were isolated and identified within the matrix of the exhaust emissions from the poultry sheds.

Figure 6. GC-MS analysis of sorbent tubes: Carbotrap300 (top spectra) and Tenax TA (lower spectra). (A – 1-butanol, B – dimethyl disulphide, C – toluene, D – styrene, E – N-butyl1-butanamine, F – 4-ethyl-decane, G – butylated hydroxytoluene).


23

3.2 GC-MS-O ANALYSIS GC-MS-O analysis allows the simultaneous collection of olfactory and mass spectral data from GC analysis. Figure 7 shows a typical total ion chromatogram with the odour chromatogram overlayed to identify the odorants within the matrix. The results shows that only a small number of the compounds present are identified by the operator as odorous, and therefore could be potentially responsible for the odorous emissions from the poultry shed samples. Figure 7 also shows that the intensity of odorous compounds can be scaled from 0-3 thereby identifying the most odorous compounds and the one’s that are more likely to cause offensive to local receptors. Table 2 lists the NMVOCs that were isolated and identified by the ODP operator as being odorous. The most predominant odorants in the poultry emission matrix was determined to be dimethyl disulphide and 2, 3-butanedione (diacetyl). The ODP operator can also include voice activated odour descriptors to describe the character of odorants (Figure 8). 3.3 VARIATIONS IN ODORANT PROFILES The correlation of dominant odorants from the poultry shed emissions (Table 2) with the results of dilution olfactometry has shown that odour emission trends can be strongly linked to the abundance of these specific compounds. Figure 9 illustrates the relationship over the grow-out cycle between the abundance of dimethyl disulphide as acquired by mass spectral data and odour concentrations (determined by dilution olfactometry). The results have been normalised to the volume of air that was being exhausted from the shed at the time of sampling and shows that the variations in odour and NMVOC emissions can be linked to the either the bird age or bird mass. Figure 10 supports these observations and shows that the emissions of two key odorants (dimethyl disulphide and 2, 3-butanedione) are also subject to diurnal variations which is most likely the result of bird activities within the shed over the 24 hours due to feeding and lighting cycles.

24


Table 1 Non-methane volatile organic compounds identified using GC-MS. Compound Family Aromatics

Sulphur

Aldehydes

Ketones

Nitrogen Alcohols Carboxylic Acids Terpines

Other Hydrocarbons

Compounds Isolated Toluene o-Xylene p-Xylene Benzene 1-ethyl-4-methyl-benzene 1-ethyl-2-methyl-benzene Acetophenone Benzaladehyde Phenol Styrene Dimethyl Sulphide Dimethyl Disulphide Dimethyl Trisulphide Butanal 3-methyl-butanal Cyclohexanal Hexanal 2-ethyl-1-hexanal 2-butanone Diacetyl 3-methyl-2-butanone 3-hydroxy-2-butanone Trimethylamine 1-butanol Cyclohexanol Acetic Acid α-pinene β-pinene Limonene Camphene Camphor Carene Eucolyptol Tetradecane Hexadecane Tetrahydrofuran


25

Table 2 Odorants identified using olfactory detection port (from Figure 7). Compound family Sulphur Ketones

Compound Dimethyl Disulphide Dimethyl Trisulphide 2,3-butanedione (diacetyl) 2-butanone Acetophenone 3-hydroxy-2-butanone

Odour Threshold Value (ppb)1 0.16 – 12 0.005 – 0.10 2.3 – 6.5 50,000 65 800

Figure 7. GC-MS-O analysis showing total ion chromatogram and odour chromatogram (A – 2-butanone, B – 2, 3-butanedione, C – dimethyl disulphide D – 3-hydroxy-2-butanone E – dimethyl trisulphide and F – acetophenone).

1

Odour Detection Values reported by Leffingwell & Associates http://www.leffingwell.com/ odorthre.htm

26


Figure 8. GC-MS-O analysis showing the additions of odour descriptors on the odour chromatogram.

Figure 9. Variations of odour and dimethyl disulphide at different stages of a typical chicken grow-out cycle.


27

Figure 10. Diurnal variations of odour and two key odorants.

4 CONCLUSIONS The GC-MS analysis of samples from different poultry sheds revealed that that there is a complex matrix of non-methane volatile organic compounds that form the emissions from these facilities. The simultaneous collection of olfactory and mass spectral data via GC-MS-O analysis demonstrated that only a small number of the NMVOC’s present in the matrix are responsible for the resulting odorous emissions. Olfactory-GC-MS analysis was able to identify the key odorants in the poultry emissions samples as dimethyl disulphide and 2, 3-butanedione. These compounds were determined to be the most odorous over the chicken grow-out cycle and showed that distinct odorant profiles occur due the different growth stage during poultry shed production (i.e. the age of the bird or the total mass of birds within the shed). The GC-MS-O analysis also showed that diurnal variations in odorants compositions where also influenced by chicken activity within the poultry sheds. As odour abatement process failure is often due to inadequate characterisation of the emission source in terms of odour composition. The application of olfactoryGC-MS analysis offer a potential approach to identify key odorous markers from different emission sources as demonstrated with the analysis of poultry shed emissions.

28


The ability to identify compounds that have greater receptor impact will enable improved design of odour abatement systems to remove specific odorous compounds. Improved characterisation of odorous emissions will also enable more effective evaluation of odour control systems performance during its operation.

5 ACKNOWLEDGEMENTS The authors acknowledge the financial support of the Australian Poultry CRC (Project 04-45 – Dust and Odour Emissions from Poultry Sheds) and thanks the Queensland Department of Primary Industries and Fisheries (Erin Gallagher, Neale Hudson, JaeHo Sohn and Mark Dunlop), the Victorian Department of Primary Industries (Maurie Miles) and UNSW (Xinguang Wang and Gautam Chattopadhyay) for their assistance in VOC sampling and analysis. Gavin Parcsi was financially supported through a PhD scholarship from the Australian Poultry CRC.

REFERENCES Gostelow, P., Longhurst P.J., Parsons, S.A. and Stuetz, R.M. (2003) Sampling for Measurement of Odours. IWA Scientific and Technical Report No.17, IWA Publishing, London. Gostelow, P., Parsons, S.A. and Stuetz, R.M. (2001) Odour measrement in sewate treatment – a review. Water Res. 35(3): 579-597. Hochereau, C. and Bruchet, A. (2004) Design and application of a GC-Sniff/MS system for solving taste and odour episodes in drinking water. Water Sci. Technol. 49(9): 81-87 Kai, P. and Schäfer, A. (2004) Identification of key odour components in Pig House Air using hyphenated gas chromatography olfactometry. Agricultural Engineering International: the CIGR Journal of Scientific Research and Development. VI(04 006). Suffet, I.H., Khiari, D. and Bruchet, A. (1999) The drinking water and odour wheel for the millennium: beyond geosmin and 2-methylisoborneol. Wat. Sci. Technol. 40(6): 1-13. Wright, D.W., eaton, D.K., Nielsen, L.T., Kuhrt, F.W., Koziel, J.A., Spinhirne, J.P. and Parker, D.B. (2005) Multidimensional gas chromatography-olfactometry for the identification and prioritization of malodors from confined animal feeding operations. J. Agric. Food Chem. 53: 8663-8672.

29

A statistical perspective on biofilter performance in relation to the main process parameters and characteristics of untreated flows ANTON PHILIP VAN HARREVELD Odournet S.L., Ctra. de l’Esglèsia 60 bis, Barcelona 08017, Spain

ABSTRACT A large number of olfactometric measurements of odour removal efficiency of municipal waste organic fraction and green waste composting installations were compiled and analysed graphically. The number of measurements and installations is >50 for treated gas characteristics and >15 for filter media characteristics. All installations concerned were located in the Netherlands and Belgium. All untreated and treated gas odour concentrations were measured in duplicate or triplicate, according to olfactometry standard EN13725 or its predecessor NVN2820. The data were then analysed using graphical methods to identify trends and relation between effectiveness of performance and a large number of operational parameters and characteristics, including: • Area flow loading (m3·m-2 filter area·hour-1) • Contact time • Temperature of untreated flow • Odour concentration of untreated flow • Ammonia concentration of untreated flow • Dry matter content of filter media • pH of filter media

1 INTRODUCTION Biofilters are widely used to treat odorous process flows from composting installations used for processing the organic fraction of municipal waste and green wastes (separately collected vegetable, fruit and garden waste). The main objective of treatment is reduction of odour emissions and their related impacts. The criteria for satisfactory treatment result vary from country to country. For example:

29

30

ANTON PHILIP VAN HARREVELD

• The Netherlands Emissions Guideline (Infomil, 2000) considers an odour concentration in the treated gas of < 2500 ouE·m-3 indicative for an efficient biofilter. • In Germany (30.BimSchV, 2001) a biofilter is expected to produce treated gas flows with a concentration of < 500 ouE·m-3. • In Spain Odournet consultants frequently see environmental license conditions and contract conditions for waste management installations that require the concentration from biofilters not to exceed 1000 ouE·m-3. These widely differing performance criteria beg the question: What is a consistently attainable endpoint for treatment of organic fraction and green waste composting process flows in biofilters? This question leads to the related question: Which are the parameters that determine the effectiveness of treatment of a biofilter? A large number of supposedly relevant parameters have been identified in literature (IPPC BREF, 2003). Some of these factors relate to the design parameters of the biofilter, and others to the characteristics of the raw gas to be treated. 1. Biofilter design parameters: a. Area flow loading. The volume flow of gas treated by 1 square meter of biofilter surface, typically expressed as m 3·m-2·hour -1. This design parameter may vary over a range of 50 to 600 m3·m-2·hour-1. In practical applications for composting installations a range of 100 to 200 m3·m-2· hour-1 is more typical. b. Time of residence. The time that the gases remain in contact with the filter medium is a function of the area flow loading and the filling height of the media material. The parameter is expressed in seconds. As the actual biological oxidation occurs in the liquid adhered to the media, the time permitted for transfer of odorous components from the gas to the liquid phase is of obvious importance. Typically, the gas retention time in the biofilter media is between 30-60 seconds. c. Temperature of filter media. The temperature of the filter media must be suitable for the biomass to thrive. At low temperatures, the metabolism will occur at a slower rate. At higher temperature, above the mesophile range of 10 to 45ºC, the population will shift from mesophile to thermophile species, while at even higher temperatures biomass will not be able to exist. Typically, the optimum temperature in a biofilter is considered to be within the range of 30 to 38ºC.

A STATISTICAL PERSPECTIVE ON BIOFILTER PERFORMANCE

31

2. Biofilter media characteristics: a. Type of filter media. There is a wide range of filter media, most of organic origin, e.g. peat, heather, wood chips, root chips, bark chips, compost, etc. Inorganic media are used in some cases, e.g. extruded clay pellets, lava rock, typically mixed with some organic material. The media provides for the basic life support of the biomass, and can contribute by providing: physical support surface, moisture, nutrient supply and pH control. b. Nutrient availability. The biomass should satisfy part of its nutrient requirements by using the components provided by the raw gas flow. However, these may not provide a balanced diet and deficiencies in nutrients such as phosphorus, nitrogen and trace elements may limit biomass development. In that case nutrient availability must be supplemented from the media of by irrigation with a nutrient solution. c. Dry matter. The dry matter content should be such that the filter structure has a homogeneous structure, while not becoming soggy, with the risk of creating anaerobic zones. An optimum dry matter content is typically in the range of 25 to 40%, in the experience of Odournet, while the literature indicates a wider range of 40-60% dry matter (IPPC BREF, 2003). d. Irrigation water. The quality of the water used for irrigation can have an impact on the availability of nutrients. Well water usually contains high iron levels, which can bind phosphate and cause a nutrient deficiency in the biofilter. e. pH. The optimum range is between pH 6 and 8. When the acidity moves outside this optimum range the filter can still function, but the population will become more specialised and hence limited in its biodiversity, which may influence its effectiveness. f. pH buffer capacity. Some materials have a larger capacity to buffer pH in situations where acids are being formed by the biological oxidation of reduced sulphur compounds (e.g. H2S) and ammonia (NH3). g. Nitrate. The total content of ammonium (NH4+), nitrite (NO2-) and nitrate (NO3-), expressed as NH4+NOx-N in g per kg wet material, can impact on the microbial activity. Literature values indicate that at values in excess of 4 g·kg-1 wet material the nitrification in the filter is completely inhibited. An alternative indicator is that a concentration in excess of 6 g·litre-1 of nitrate in the liquid fraction cause a reduction in microbial activity. A range of between 0.25 and 3.5 g of NH4+NOx-N per kg wet material should be maintained.

32


h. Sulphate. High concentrations of sulphate (SO42-), formed by biological oxidation of reduced sulphur compounds can cause low pH and an excess of conductivity, which in turn indicated suboptimal osmotic conditions for the biomass. i. Conductivity. The electric conductivity of the media is indicative for the salinity of the liquid in the media. If the salinity, or the concentration of dissolved ions, becomes too high, the osmotic properties of the liquid become unfavourable for microbial activity. As a rule of thumb the electrical conductivity should be less than 1000 μS·cm-1. 3. Raw gas characteristics: a. Temperature. The raw gas should enter the biofilter at a temperature that is suitable to maintain the optimum media temperature as discussed above. b. Relative humidity. The raw gas should enter the biofilter media fully saturated with humidity to avoid any drying out of the filter. Drying out of the filter media is the most common cause of biofilter failure, in the experience of Odournet consultants. c. Odour concentration. The odour concentration of the raw gas will determine the job at hand of the biofilter. As the range of odorants can be extremely variable it is difficult to provide a full discussion. Generally speaking, biofilters typically achieve higher efficiencies when used at higher concentrations (100000 to 1000000 ou E·m -3) than at lower concentrations. Compounds that are less soluble are treated with less efficiency, for obvious reasons. d. Absence of toxic or inhibiting compounds. Toxic compounds in the raw gas should be avoided (CO, chlorinated compounds). Large concentrations of less odorous compounds, such as CH4, may reduce the effectiveness of the filter in treating the more odorous compounds. This paper explores the data obtained from a large number of measurements conducted in the Netherlands and Belgium in a variety of biofiltration units servicing composting processes. Some of the units were assessed repeatedly over a period of approx. 5 years. The measurements were all conducted by dynamic olfactometry according to the EN13725:2003 standard, in laboratories with an accredited quality system according to ISO17025. The concentration of the treated gas in relation to some of the parameters discussed above is explored in a number of graphs, to provide an insight based on statistics obtained under real-world conditions. As these observations were obtained from actual performance evaluations, and not from a controlled experimental setup, it was not possible to keep all variables constant and vary one parameter of interest.


33

Covariance between non-independent variables is therefore a factor to be kept in mind.

2 MATERIALS AND METHODS 2.1 OLFACTOMETRY The odour concentrations in ouE·m-3 of raw gas and treated gas flows, collected before and after the biological treatments, were measured in strict adherence to the NVN2820:1996 and later the EN13725:2003 standard for olfactometry (Van Harreveld, 1998). All measurements were conducted in laboratories with a quality system according to ISO17025 in place, and accredited on this basis by the Accreditation Council of the Netherlands, under the umbrella of European Accreditation. All measurement results are based on a minimum of duplicate samples, and typically on triplicate samples of raw gas and treated gas. The geometric mean of the individual measurements is presented in the graphs. The measurements were conducted during the period between 1995 and 2000.

3 RESULTS AND DISCUSSION 3.1 TREATED GAS ODOUR CONCENTRATION IN RELATION TO AREA FLOW LOADING Figure 1 shows the relation of outgoing odour concentration at different levels of area loading, in the range between 40 and 350 m3·m-2·hour-1.

Figure 1. Relation between the odour concentration in the treated gas and area flow loading.

34


The graph in Figure 1 does not show a clear correlation between area flow loading and treated gas odour concentration in the range considered. However, concentrations of the treated gas below 1000 ouE·m-3 are only seen to occur at area flow loadings below approximately 180 m3·m-2·hour-1. 3.2 TREATED GAS ODOUR CONCENTRATION IN RELATION TO RESIDENCE TIME Figure 2 shows the relation of outgoing odour concentration at different residence times, in the range between 20 and 90 seconds.

Figure 2. Relation between the odour concentration in the treated gas Cod in ouE·m-3 and residence time in seconds.

The graph does not show a clear correlation between residence time and treated gas odour concentration in the range considered. 3.3 TREATED GAS ODOUR CONCENTRATION IN RELATION TO THE TEMPERATURE OF THE RAW GAS

Figure 3 shows the relation of outgoing odour concentration at different temperatures of the raw gas, in the range between 24 and 8ºC.


35

Figure 3. Relation between the odour concentration in the treated gas and the raw gas temperature.

The graph does not show a clear correlation between raw gas temperature and treated gas odour concentration in the range considered. It should be emphasized that the large majority of the observations are below 40ºC. Treated gas concentrations below 1.000 were not observed at temperatures in excess of 37ºC. 3.4 TREATED GAS ODOUR CONCENTRATION IN RELATION TO DRY MATTER CONTENT OF THE MEDIA

Figure 4 shows the relation of outgoing odour concentration at different levels of dry matter content of the media, in the range between 25% and 40%. The graph does not show a clear correlation between dry matter content and treated gas odour concentration in the range considered, although less favourable treated gas values appear to occur at low dry matter content. It should be noted that the range considered coincides with the range of advisable values.

36


Figure 4. Relation between the odour concentration in the treated gas and dry matter content of the media.

3.5 TREATED GAS ODOUR CONCENTRATION IN RELATION TO PH OF THE MEDIA Figure 5 shows the relation of outgoing odour concentration at different levels of pH of the media, in the range between 5.0 and 7.5.

Figure 5. Relation between the odour concentration in the treated gas and pH of the media.


37

The graph does not show a clear correlation between pH and treated gas odour concentration in the range considered, although less favourable treated gas values appear to occur at the extremes of the range considered. It should be noted that the range considered almost coincides with the range of advisable values of pH between 6 and 8. 3.6 T REATED

GAS ODOUR CONCENTRATION IN RELATION TO RAW GAS AMMONIA

CONCENTRATION

Figure 6 shows the relation of outgoing odour concentration at different levels of raw gas ammonia concentration, in the range between 0 and 120 ppm, with one observation at approximately 200 ppm.

Figure 6. Relation between the odour concentration in the treated gas and ammonia concentration.

The graph does not show a clear correlation between the ammonia concentration in the raw gas and the treated gas odour concentration in the range considered. Even at concentrations in excess of 50 ppm, that are generally considered undesirable, the treated gas odour concentration appears to remain at fairly typical levels. However, it should be noted that these are snapshot values. The detrimental effect of sustained high ammonia loads will show itself with time, when the electrical conductivity is increased due to increased levels of NH4+NOx-N, leading to reduced

38


microbial metabolism and hence treatment capacity. For the observations presented in the figure above information on the standing time of the filter medium is not available. 3.7 TREATED

GAS ODOUR CONCENTRATION IN RELATION TO THE CONCENTRATION OF THE

RAW GAS

Figure 7 shows the relation of outgoing odour concentration at different concentrations of raw gas odour concentration, in the range between a few thousands and 500000 ouE·m-3.

Figure 7. Relation between the odour concentration in the treated gas and the raw gas.

Even at very high raw gas concentrations, of several hundreds of thousands of ouE·m-3, the treated gas will have a concentration rarely exceeding 10000 ouE·m-3. Of the measured odour concentration in the treated gas (n = 123) of biofilters treating organic green waste, 4% of the odour concentrations in the treated gas were found to be less than 500 ouE·m-3, a requirement often derived from German regulations. Only 14% of treated gas measurements was < 1000 ouE·m-3, while 52% was < 2500 ouE·m-3, a criterion used in the Netherlands for a as indicative for a well functioning biofilter. Based on the same data, the removal efficiency of the biofilters is plotted in relation to the odour concentration of the raw gas in Figure 8.


39

Figure 8. Relation between the odour removal efficiency an the odour concentration in the raw gas.

The graphs show that biofilters have a capability to treat a wide range of raw gas odour concentrations with a remarkably consistent efficiency, typically at >80%. The lower efficiencies are presumably an indication for suboptimal operational conditions, as they typically occur at lower odour loads.

4 CONCLUSIONS The observations presented in this study lead to the following conclusions: 1. Biofilters appear to be a very robust treatment method suitable for reducing odour emissions in raw gas from green waste and organic fraction composting processes, ranging from a few tens of thousands of ouE·m-3 to half a million of ouE·m-3 with odour removal efficiencies in the majority of cases in excess of 90%. 2. For the parameters considered in this study, which were generally within the suggested range of values in the literature, no clear relation between the parameter value and the treated gas odour concentration could be observed. 3. The values of residual odour concentration are often in excess of the targets that are applied in various countries as emission criteria for composting odours after treatment with biofiltration. Of the data considered:

40


a. 4 % satisfied the German criterion of Cod, treated < 500 ouE·m-3 b. 14 % complied with the criterion of Cod, treated < 1000 ouE·m-3 that is often seen in Spanish waste management contracts and environmental licences. c. 52 % of the observations was found to satisfy the criterion of Cod, treated < 2500 ouE·m-3 that is mentioned in the Netherlands Emission Guideline as indicative for a well functioning biofilter. 4. A performance criterion more restrictive than a treated gas concentration of Cod, treated < 2500 ouE·m-3 appears to be too restrictive and not viable as a value that operators need to attain consistently.

5 ACKNOWLEDGEMENTS The data for this study were provided by anonymous clients of the Odournet group and compiled by Margrethe Bongers, formerly a senior consultant at Odournet.

REFERENCES 30.BimSchV, Bundesimmissionsschutzgesetz, Dreißigste Verordnung zur Durchführung des BundesImmissionsschutzgesetzes paragraph 6, Emissionsgrenzwerte, 2001, http://bundesrecht. juris.de/bimschv_30/index.html EN13725:2003. Air Quality – Determination of Odor Concentration by Dynamic Olfactometry, Brussels: CEN. Infomil. (2000) NeR Nederlandse Emissie Richtlijn. (Netherlands Emissions Guideline), ISBN 90 76323 01 1, 2000 (English version available via www.infomil.nl). IPPC BREF 2003, Reference Document on Best Available Techniques in Common Waste Water and Waste Gas Treatment / Management Systems in the Chemical Sector, European Commission, subclause 3.5.2.1., Seville, http://eippcb.jrc.es NVN2820:1990 Luchtkwaliteit. Sensorische geurmetingen met olfactometer (Air Quality – Sensory measurement of odour using an olfactometer), NEN Netherlands Normalisation Institute, Delft, 1990. www.nen.nl Van Harreveld, A.Ph. (1998) A review of 20 years of standardization of odour concentration measurement by dynamic olfactometery in Europe. J. Air Waste Manage. Assoc. 49: 705-715.

41

Refurbishment of an odour collection and biofilter treatment system at a municipal solid waste composting facility in Perth, Western Australia TERRY SCHULZ AND STUART MCALL Principal of the Odour Unit Pty Limited, Sydney, Australia and CEO of Southern Metropolitan Regional Council, Perth, Australia

ABSTRACT This paper discusses a case study at a large, regional Municipal Solid Waste composting plant in Perth, Western Australia, where odour emissions from the facility have resulted in adverse odour impacts in the surrounding community. The plant was designed as a fully-enclosed operation, where all ventilation air from two major processing buildings was to be collected and treated in four large biofilters. The paper describes a detailed investigation into the collection and treatment systems at the plant, which identified deficiencies in the design of both systems. It contains the results of the physical and olfactory investigations, and documents the design for an upgraded collection and biofilter-treatment system that has subsequently been installed.

1 INTRODUCTION Southern Metropolitan Regional Council owns and operates a large Municipal Solid Waste Composting Facility (WCF) in Canning Vale, Perth, Western Australia. The WCF has a capacity of 150,000 tonnes/year of MSW, using a fully-enclosed, ‘invessel’ composting process. It is one of four large MSW composting plants in Australia. The location of the Canning Vale WCF is within a larger Regional Resource Recovery Centre, in the suburbs of Perth, the capital of Western Australia. It has been operating since late-2002.

41

42

TERRY J. SCHULZ AND STUART McALL

Perth is a city of approximately 1 million people and is the capital of Western Australia. It has a very hot and dry climate, with summer temperatures regularly above 350 Celcius, and ambient relative humidities that can stay below 40 per cent for several months at a time. As will be explained, these weather conditions can have a strong influence on the design and performance of biofilter-based odour control systems in the area. Since the commissioning of the Canning Vale WCF a high level of adverse odour impacts has been experienced in the residential areas up to 400 metres from the plant. This situation lead to the development of a strong and well-organised community opposition to the continued operation of the facility, and resulted in increasing pressure from the WA environmental regulatory authority to mitigate the odour emissions. In February 2006, following a number of earlier attempts to identify the reasons for the unacceptable odour emissions from the plant, investigations commenced on a comprehensive assessment of the design and performance of the entire odour control collection and treatment system at the plant. The methodology used, the findings and the implementation of the findings of that study are described in this paper. The outcome of the study has been a comprehensive re-design of the odour collection system and a total refurbishment of the four biofilters that treat all odour emissions to atmosphere. A full report (The Odour Unit, 2006) documenting the full investigation and results is available on the SMRC website.

2 DESCRIPTION OF THE FACILITY The Canning Vale WCF was designed using the proprietary ‘Bedminster’ composting system. This system currently receives 120,000 tonnes per year of raw MSW from six local government areas in Perth, and processes it into a fully composted product, in the process separating plastics and metals for subsequent recycling. While there are several different types of MSW composting processes in use around the world, the features of this particular WCF composting process may be summarised as follows: • A separate MSW receivals hall (the Tipping Building), having a volume of approximately 20,000 cubic metres; • Four, rotating tunnel digesters, in two banks of two digesters, each bank receiving MSW in the Tipping Building and unloading into different ends of the Aeration Building; • The Aeration Building, in which the partially digested material is processed further using a forced aeration technique for a minimum of 45 days. The volume of this building is 200 metres long, 50 metres wide and has a volume

REFURBISHMENT OF AN ODOUR COLLECTION AND BIOFILTER TREATMENT SYSTEM

43

of approximately 100,000 cubic metres. This paper will focus predominantly on the capture and treatment of the odour emissions from this building; • The Load-out wing of the Aeration Building, where fully composted material is stored and loaded into trucks; and • The biofilters. There are four biofilters in total. Two smaller units service the Tipping Building, while two larger biofilters receive and treat air drawn from the Aeration Building. The two Tipping Building biofilters are each fed by two identical fans, each having a capacity of 55,000 m3/hr. One of these fans extracts air from above the inlet to the digester vessels, while the other draws air from within the body of the building. The discharge from each fan enters a packed-column scrubbing unit, designed to remove particulates from the air stream and provide some humidification. The design of the building is symmetrical, having two digesters, two fans, two scrubbers and a biofilter on each side. The original design of the Tipping Building had a single roller-shutter doorway for truck access. The Tipping Building biofilters (#3 and #4), like the Aeration Building biofilters (#1and #2), are open bed, roofed units, featuring a foul air distribution system based on the use of multiple, slotted PVC pipes buried in gravel, and fed from a longitudinal distribution header duct. Each fan feeds to its own header duct and biofilter section. As a result, each of the two Tipping Building biofilters behaves as two separate units, although visually appearing to be a single biofilter. There is no capability to direct the output from one fan to another section of the biofilter. The Aeration Building foul air extraction system uses six 72,000 m3/hr fans mounted outside the longest side of the building. Fans 1 to 4 direct air to Biofilter #1, (the largest biofilter – 1,500 m2), while fans 5 and 6 service Biofilter #2 (1,000 m2). The original design of the extraction system involved the six fan-suction ducts terminating inside the wall, approximately five metres above ground level. A number of wall-mounted inlet vents were located along the wall opposite the fans, with the intention of affecting a flow of air laterally across the building. The two ends of the building, containing the odorous digester outlets, the compost conveyors and trommel screens, were not fitted with any form of collection system. The roof height at the ends of the building is higher than in the aeration/maturation section, rising to a maximum height immediately adjacent to the roller shutter doors used to load-out screenings. As with the Tipping Building biofilters, the two Aeration Building biofilters behave as six independent units, each fed by its own fan, with no cross connection between these units. A summary of the odour collection and treatment system ‘as installed’ at the WCF is shown in Table 1.

44


Table 1. Canning Vale WCF odour control system summary. Tipping Building Ventilation

Tipping Building Biofilters

Aeration Building Ventilation

Aeration Building Biofilters

20,000 m3 building volume 4 fans, each 55,000 m3/hr 10.5 air changes/hour 2 biofilters, 800 m2 bed area per biofilter plenum pipe-in-gravel air distribution wood chip/compost bed medium 120,000 m3 building volume 6 fans, each 72,000 m3/hr fan suction ducts in wall cross flow air movement 4.3 air changes/hour 2 biofilters, 1,500 m2 and 1,000 m2 bed area plenum pipe-in-gravel air distribution wood chip/compost bed medium

3 PROBLEMS IDENTIFIED IN PREVIOUS STUDIES As mentioned in Section 1, the operation of the WCF had resulted in the detection of nuisance odours identified as having originated from the facility, up to 400 metres from the plant. Unfortunately the odour complaint records retained by the regulator and SMRC contained little odour descriptor information and therefore did little to identify the source activity within the WCF responsible for the nuisance. Typical descriptors used by complainants were ‘garbage’, ‘waste’, ‘landfill’ and ‘rotting waste’. A previous odour impact assessment study carried out in 2005 identified three key odours that were judged to be responsible for the bulk of the odour complaints. These odours, and their sources within the WCF, are summarised in Table 2. The clear conclusion from this earlier investigation was that the biofilters and Aeration Building were the most problematical odours, and that the odours from the Tipping Building, were not likely to be detectable in the community. At that time the biofilters were known to be operating sub-optimally, due to bed moisture control problems and the resulting tendency for foul air short-circuiting. In an effort to counter this problem SMRC elected to decrease the fan speed, and hence output, of each of the ten biofilter fans. This decision was taken in early 2005.


45

At the time of this study there appeared to be little understanding of the magnitude of any fugitive odour releases from the Aeration Building. Table 2. Odour types within the Canning Vale WCF (2005 Study). Odour character Garbage, rotting vegetative material Fruit cake, fermented Fruit Stale air or water, rubber tyre Cheesy, baby vomit Earthy, damp, mild garbage Forest floor, mouldy, medicinal,

Likely odour source Tipping Floor, Digester discharge area, Aeration floor Aeration floor Aeration floor Biofilters 1 and 2 Biofilters 3 and 4

4 INVESTIGATION METHODS The investigation was carried out from February to April 2006. In the course of this investigation a number of methodologies were used to assess both odour impacts and the performance of the odour collection and treatment systems. These are summarised below. 4.1 ODOUR IMPACT ASSESSMENT 4.1.1 OLFACTORY ASSESSMENT AT THE SOURCE This procedure involved The Odour Unit staff assessing the character, intensity and relative emission rate of odours generated within the facility. While subjective in nature, olfactory assessment is able to differentiate between different odour characters. If a clearly defined odour can be detected at a processing source, this enables this same odour and source to be identified if detected downwind of the facility during an ambient odour assessment. 4.1.2 AMBIENT ODOUR SURVEYS The use of quantitative ambient odour assessments is now commonplace in odour impact assessment worldwide. In the absence of an Australian Standard method for this type of assessment a simple but robust procedure was developed, involving a trained and calibrated assessor systematically traversing the downwind areas from an odour source, logging wind data and the incidence, character and intensity of odours present. In this project the ambient surveys were carried out in the neighbouring down wind residential areas.

46


4.1.3 COMMUNITY FEEDBACK The scope of work for this project required TOU to interact with community members, as a means of identifying the most problematical odours from the facility. On two occasions samples of odours from a range of processing sources within the WCF were presented to community representatives. The sources assessed were: • • • •

Tipping Floor odour; Aeration Floor odour; Tipping Floor biofilter-treated air; and Digestion odour (collected during a digester unloading event).

On the second occasion a community representative was presented with samples collected at source and therefore at ‘full strength’, as well as the same samples diluted to a level likely to be experienced in the community (10 to 20 odour units). This was done to cover the possibility that odour character could change with dilution. 4.2 ODOUR CONTROL SYSTEM ASSESSMENT 4.2.1 DUCT AIRFLOW MEASUREMENT Airflow rates in the various ducts in the collection system and into the biofilters were measured using a hot-wire anemometer that measured gas velocity. Where possible the velocity measurements were taken in a straight section of ducting to maximise accuracy of the measurement. All flow readings presented in this report are expressed at the temperature prevailing in the duct. 4.2.2 DUCT PRESSURE MEASUREMENT Pressure or vacuum was measured simultaneously with velocity measurement in each of the ducts. A simple water-filled manometer was used. The pressure readings were used to determine the pressure duty on each of the 10 fans tested, as well as the back-pressures upstream of each of the cells in the biofilters. Relative humidity was measured into each of the biofilters, and in the Tipping Floor and Aeration Buildings. Ambient readings were also taken. A TSI VelociCalc 8386A instrument was used for this purpose. 4.2.3 BIOFILTER SURFACE FLOW MEASUREMENTS A technique was developed to assess the extent to which spatial variation in airflow rates occurred across the surface of the biofilter beds. This technique involved the use of a cylindrical, open-bottomed ‘hood’ (0.8 m diameter) that was placed on the surface of the biofilters. A 100 mm PVC pipe mounted on the lid of the hood accelerated the low velocity of air leaving the bed, enabling a velocity reading to be accurately taken.


47

This technique has been found to accurately assess relative changes in airflow leaving the surface of a biofilter bed. In this respect it is useful in assessing whether short-circuiting or blockages exist in biofilter beds. A variation in velocity readings of greater than 30 percent would indicate an air distribution problem. However the technique is not able to measure the true airflows leaving a biofilter bed, due to the intrusive nature of the device. 4.2.4 BIOFILTER PRESSURE LOSS MEASUREMENTS A means of quantifying the pressure losses across Biofilter 1 was needed in order to differentiate between the losses through the plenum distribution pipes and the biofilter medium. Two holes were dug in Biofilter 1, one in Cell 1 serviced by Fan 1, and the second at Cell 4. Cell 4 is serviced by Fan 4, which in turn draws particulateladen air into the biofilter from the final product screening area. Two sets of manometer tubing were inserted into the bed at these two locations. The first tube at each location was located immediately above the plenum pipe and beneath the shade cloth mesh layer, in the gravel layer, beneath the biofilter medium. The second tube was positioned above the shade cloth. The holes were backfilled and compacted back to as near their original condition as was possible, and pressure readings taken 4.2.5 BULK AIR MOVEMENT ASSESSMENT This study found it imperative that the general airflow patterns be assessed within the Aeration Building. A portable smoke generating machine was used for this purpose. The machine is able to generate copious quantities of smoke for 15-20 seconds at a time, and was sufficient to fill a section of the building and determine airflow patterns. This method was used with great success and was easily able to identify deficiencies in the odour collection system design. 4.2.6 ODOUR TESTING A decision was taken early in this study not to use olfactometry testing. While such testing would have quantified the odour removal performance of the biofilters it was felt that the performance of the biofilters was already known to be sub-optimal.

5 RESULTS 5.1 ODOUR IMPACT ASSESSMENT The results of the three methods used to assess the WCF processing sources emitting the odours being detected in the community showed clearly that two odours present in the Aeration Building were responsible. These sources were the aerated

48


compost piles (a compost odour) and the digester unloading and screening operation (a fermenting/fruity odour). The source of both of these odours was determined to be fugitive odour emissions from the Aeration Building. It was also determined that incomplete odour destruction in the biofilters was also causing odour impacts, suggesting that the biofilters needed refurbishment. 5.2 DUCT AIRFLOW AND PRESSURE MEASUREMENTS The results of airflow and pressure testing are shown in Table 3. Table 3. Airflow and pressure measurements WCF biofilters (16 March 2006). Location Biofilter 1 – Fan 1 Biofilter 1 – Fan 2 Biofilter 1 – Fan 3 Biofilter 1 – Fan 4 Biofilter 2 – Fan 5 Biofilter 2 – Fan 6 Fan specifications Biofilter 3 – Fan 11 Biofilter 3 – Fan 12 Biofilter 4 – Fan 9 Biofilter 4 – Fan 10 Fan specifications

Airflow (m3/hr) 45,600 45,200 43,900 40,700 * 49,600 72,000 35,700 46,500 30,700 34,500 55,000

Biofilter delivery pressure (kPa) 3.9 3.9 3.6 3.9 * 4.0 4.5 (suction plus delivery) 1.6 0.9 ** 1.8 1.6 4.5 (suction plus delivery)

* not accessible for measurement ** biofilter bed was dry, scrubber out of service

5.3 FOUL AIR HUMIDITY MEASUREMENTS The results of the humidity testing on the Tipping Floor Biofilters 3 and 4, which were equipped with scrubber/humidifier units for each fan, indicated that acceptable relative humidity (RH) readings above 90% could only be achieved when the ambient RH was above 60%. The overall finding was that the scrubber/humidifiers were only able to achieve an RH net increase of 31.3 percentage units. Given that the WCF operates in Perth’s very dry climate, where summer RH values in the range 2030% are normal, it was determined that the Tipping Floor Biofilters 3 and 4 were operating under considerable stress. The problem was found to be even worse for the Aeration Building Biofilters 1 and 2 where no form of foul air humidification was provided.


49

5.4 BIOFILTER SURFACE FLOW MEASUREMENTS These results showed erratic and unacceptable airflows across the surface Biofilters 1 and 2. A five-fold variation in surface flows was measured, indicating poor distribution of foul air beneath the biofilter beds. 5.5 BIOFILTER PRESSURE LOSS MEASUREMENTS Table 4 contains the results of pressure loss testing in two sections of Biofilter 1. The section serviced by Fan 1 was receiving air typical of that contained in the bulk of the Aeration Building. The section serviced by Fan 4 was receiving air laden with a high particulate loading, drawn from the product screening area. Table 4. Pressure losses across Biofilter 1 (5 April 2006). Location Southern end of Biofilter 1 (Fan 1) Pressure into Plenum Pipe Outside pipe and below shade cloth mesh (pressure loss through slotted pipe) Above shade cloth mesh, below medium At biofilter medium surface Northern end of Biofilter 1 (Fan 4) Pressure into Plenum Pipe Outside pipe and below shade cloth mesh (pressure loss through slotted pipe) Above shade cloth mesh, below medium At biofilter medium surface

Pressure (kPa) +3.80 +2.00 (1.80) +1.80 0.00 +4.30 +1.10 (3.20) +1.10 0.00

Note: these readings were taken at the reduced airflows shown in Table 3

5.6 BULK AIR MOVEMENT ASSESSMENT The smoke testing carried out inside the Aeration Building revealed highly significant adverse airflows were occurring, as follows: • There was no consistent airflow pattern evident; • There was no cross-flow of air from the inlet louvres to the fan suction ducts; • Thermal buoyancy of the air above the compost piles was dominating the air movement within the building; and • The internal, auxilliary ‘push’ fans, designed to assist the biofilter fans in the cross-flow of air from the inlets to the fan suction ducts were having a negative effect on this desired air movement.

50


It was concluded from the smoke testing that a thermally induced, positive air pressure was occurring under the roof of the Aeration Building, despite the action of the fans. The pressure at floor level was found to be neutral (neither positive nor negative).

6 DISCUSSION The technical component of the study were able to provide evidence that confirmed the findings of the odour impact assessment investigations, namely that fugitive odour emissions from the Aeration Building at the WCF were likely to be causing adverse odour impacts in the community. The cause of these fugitive emissions was determined to be an inability to achieve and maintain negative pressure conditions inside the building, due primarily to two factors - the reduced outputs from the six Aeration Building fans, and the thermally buoyant plumes rising from the compost piles. It was also determined that this situation was being exacerbated by the failure of the existing odour collection system within the building to capture the highly odorous emissions from the digester load-out and screening operations, at each end of the building. The pressure and flow measurements for the Aeration Building biofilters revealed a serious problem and justified the decision to decrease the fan outputs, in the short term. The results showed excessive pressure losses through the plenum pipe air distribution system, due to a combination of high design air velocities and large loses through the slotted orifices in the pipes. While the pressure loses through the biofilter beds were acceptable, the medium was found to be at the end of its useful life.

7. REFURBISHMENT DESIGN AND PERFORMANCE As a result of the findings of this study the odour collection system within the Aeration Building and all four biofilters have been refurbished, as follows: • The fans have been returned to their design airflows; • A longitudinal header duct has been installed along the length of the building, immediately under the roof apex. All six fans draw from this header duct. Secondary ducts have been installed at each end of the building to capture the air from the digester load-out and screening emissions; • The biofilter plenum pipes have been fitted with 30 to 40 mm holes at 500 mm spacing, to reduce pressure losses into the biofilter medium. Approxi-


51

mately 50,000 holes have been installed in 560 plenum pipes in the four biofilters; • The medium in all four biofilters has been replaced; and • A high-pressure misting system has been installed in the Aeration Building. This system is able to raise the ambient humidity in the building to at least 90%. This system was commissioned in early 2007 and has been found to achieve strong biofilter performance and measurable negative pressures at all levels and locations within the Aeration Building. While this represents an effective technical solution to the odour emissions problem and has resulted in a sharp decrease in odour complaints, the local community opposition to the continued operation of the WCF is still significant. SMRC is continuing to work with the community to allay its residual concerns.

REFERENCE The Odour Unit (2006) Canning Vale waste composting facility Odour control system review. (www.smrc.com.au/pdf/SMRC_FinalReportForCommunity18052006.pdf).


53

Odour and H2S degradation in a full scale biofilter with a mineral based organic coated filter media FRANZ-BERND FRECHEN1, WOLFRAM FRANKE1

AND

BJÖRN SCHOLL2

1

Department of Sanitary and Environmental Engineering, University of Kassel, Kurt-Wolters-Strasse 3, 34125 Kassel, Germany 2 Magistrate of the city of Frankfurt/Main, Goldsteinstrasse 160, 60528, Frankfurt/Main, Germany

ABSTRACT In order to minimize the odorous emissions from sludge storage tanks on the waste water treatment plant «Niederrad» of the city of Frankfurt/Main, about 12.000 m3/h of foul air has to be treated. Due to high Hydrogen-Sulphide (H2S) loads, the installed standard biofilter systems failed operation after one year. Thus, one of the existing filter beds was filled with a mineral based organic coated material; the other one was re-filled at the same time with a standard biofilter media to allow a comparative study. In a long term monitoring program from May 2006 to June 2007, both media were compared regarding degradation of H2S and odour. The one-year measurement program revealed that the mineral high performance media performs much better then the standard organic media.

1 INTRODUCTION In order to minimize the odorous emissions from sludge storage tanks on the wastewater treatment plant «Niederrad» of the city of Frankfurt/Main, about 12.000 m3/h of foul air has to be treated. A first step was the installation of a biofilter with a humidifier, designed in 2004 according to estimated loads. Shortly after start-up of the biofilter operation, severe odorous emissions from the biofilter were recognized. In 2005, a first filter monitoring campaign, performed by the operator, showed that the H2S concentration in the raw gas was much higher than estimated during design and was

53

54

FRANZ-BERND FRECHEN, WOLFRAM FRANKE AND BJÖRN SCHOLL

also much higher than recommended in the VDI guideline 3477 (2004) for biofilter design. As a replacement of the biofilter with a chemical scrubber would have caused serious problems in terms of space restrictions and economic aspects, the first choice was to undertake a test with a new, high performance biofilter media, which, according to the manufacturer, would be able to deal with high H2S loads. This so called IHCS (inert hydrophilic compound structure) media was filled in one of the two existing filter beds in spring 2006; the other one was re-filled at the same time with a standard biofilter media to allow a comparative study. In a long term monitoring program from May 2006 to June 2007, performed by the University of Kassel, both media were compared regarding degradation of odour and H2S.

2 MATERIALS AND METHODS 2.1 TREATMENT PLANT LAYOUT The technical specifications of the air treatments system is given in Table 1. Table 1. Technical specification of air treatment. Aspect Air stream Filter size Filter volume Filter media

Unit m3/h m2 m3 -

line 1 6.000 85 150 Mineral based organic coated IHCS media by Otto Industries, Bad Berleburg (D)

line 2 6.000 85 150 Wood mix (root wood, chipped wood and bark mulch)

The IHCS media was developed by Otto Industries, Bad Berleburg (D), and it was the first time that its performance was comparatively tested in a large scale application on a wastewater treatment plant. The core of the IHCS media consists of clay granulate and the coating is a mixture of cement, fertilizer, activated carbon and chipped wood (OTTO, 2005a). The material could be recycled by the producer (OTTO, 2005b). Figure 1. shows a sketch of the installed air treatment system.

ODOUR AND H2S DEGRADATION IN A FULL SCALE BIOFILTER WITH A MINERAL

55

Figure 1. Scheme of the air treatment system with sampling points.

All sampling points (behind ventilator line 1 [bv1], behind humidifier line 1 [bh1], behind filter line 1 [bf1], bv2, bh2 and bf2) are marked in the scheme. For comparison of the filter performance, this paper focuses on sampling points bh1, bh2, bf1 and bf2. 2.2 ODOUR MEASUREMENT The quantification of odour is the key to describe the odorous performance of an air treatment system. In order to fulfil European standards, the odour concentration was measured according to DIN EN 13725:2003, updated with DIN EN 13725:2006. The basis of this measurement method is that odour could be recognized by a person if the individual odour threshold is exceeded. For example, if an odorous air needs to be diluted 2n times for just not exceeding and n times for exceeding the odour threshold, the odour concentration is the geometric mean of n and 2n, which is 1.41n. For improving statistical accuracy on the whole, twelve or more single

56


measurement results are needed to calculate the dilution number. Although a dilution number has no unit, the odour concentration cod is denoted with European odour units per cubic meter (ouE/m3). 2.3 H2S-MEASUREMENT H2S-Measurement was performed with Odalog measurement devices by AppTek (AUS). An OdaLog detects H2S-concentrations from 0 ppm to 200 ppm with a resolution of 0,1 ppm. A data logger is included, which can store 32.000 sets of values (date/time, concentration of H2S and temperature). The device has an IR-connector for periodical data acquisition and optionally (OdaTrak) a fibre optics connector for online data transfer.

3 RESULTS AND DISCUSSION 3.1 ODOUR MEASUREMENT The result of the odour measurement program is shown in Figure 2.

Figure 2. Results of the odour measurement program.

On the whole, eleven samples were taken and analysed during one year. The IHCS-filter (bf1) always showed a higher efficiency and most of the time reached

ODOUR AND H2S DEGRADATION IN A FULL SCALE BIOFILTER WITH A MINERAL

57

significant lower clean gas concentrations than the standard woodmix biofilter (bf2). The difference between the two materials tended to become bigger near the end of the measurement campaign. The deviation in concentration values in raw gas between both air treatment lines was caused by the disadvantageous connection of the pumping station exhaust air outlet behind the air splitter into line 1 pipe (Figure 1) that made this highly polluted air stream move only into line 1 and not into line 2. 3.2 H2S MEASUREMENT The online monitoring delivered 350,000 to 430,000 sets of values for each sampling point during one year. The difference was caused by availability of devices and some necessary service stops. In Figure 3, the frequencies of the H2S-concentrations is presented. The raw gas concentration of H2S showed an average value of 25.7 ppm for the IHCS filter (bh1) and 23.9 ppm for the wood mix biofilter (bh2). It can be seen clearly that H2S degradation was performed successfully in both filter units, as H2S values of less than 1 ppm were detected in the clean gas. The effect of connecting the pumping station exhaust air to the pipe system behind the air splitter can be seen clearly. Only line 1 employing the IHCS filter material is loaded with this highly polluted air.

Figure 3. Results of H2S-Monitoring.

58


An estimation of the total H2S load for both filters can give an impression of the possible load capacities. Other sulphurous gases like thiols or dimethyl sulfide (DMS) were disregarded as well as the sulphur runoff from the two biofilter line effluents that will definitively contain sulphuric acid; because no data were available on this streams. Keeping this in mind, the estimated H2S loads are 1,919 kg H2S/a for the IHCS filter and – slightly less – 1,786 kg H2S/a for the wood mix biofilter (6,000 m3/h, mean values of H2S concentrations). With regard to filter volumes of 150 m3, the specific H2S load capacity of the IHCS was 12.8 kg/m3 without filter breakdown and for the standard wood mix biofilter 11.9 kg/m3 with operation failure in the end.

4 CONCLUSIONS Both air treatment systems operated successfully during most of the time. Although the IHCS filter was loaded with more H2S and odour than the standard filter, it was more advantageous concerning degradation performance, even increasing towards the end of the one year operation. Hence, the mineral based organic coated biofilter media seems to be an interesting way of improving standard biofilters. This material allows operation of biofilters with average H2S concentrations of 20 ppm, which is much higher than VDI guideline 3477 (2004) recommends. Economic aspects depend on the alternatives. The IHCS media is more expensive (specific material cost 270 €/m3) than a standard biofilter media (specific material cost 60 €/m3). In this case the upgrade of the biofilter with the new IHCS media was more economic than the installation of a chemical scrubber regarding investment as well as annual cost. REFERENCES DIN EN 13725:2003: Air quality. Determination of odour concentration by dynamic olfactometry; German version EN 13725:2003. Beuth Verlag. Berlin, Germany. 2003. DIN EN 13725:2006: Air quality. Determination of odour concentration by dynamic olfactometry; German version EN 13725:2003. Corrigenda to DIN EN 13725:2003-07. German version EN 13725:2003/AC:2006. Beuth Verlag. Berlin, Germany. 2006. Otto. (2005a) Safety data sheet of IHCS Media. Otto Luft und Klimatechnik GmbH & Co. KG. Bad Berleburg, Germany. 28.02.2005. Otto. (2005b) Biological air pollution control. Product information. Otto Luft und Klimatechnik GmbH & Co. KG. Bad Berleburg, Germany. 2005. VDI Guideline 3477. (2004): Biological waste gas purification - Biofilters. Beuth Verlag. Berlin, Germany. 2004.

59

High throughput biofiltration for odour control at water purification plant VITALY ZHUKOV1, ANDREY VEPRITZKY2, LEONID MITIN2, VLADIMIR POPOV1 1 A.N.Bakh Institute of Biochemistry, Russian Academy of Sciences, Leninskiy pr., 33 119071 Moscow, Russia 2 Innovational Biotechnologies Ltd., Leninskiy pr., 33 119071 Moscow, Russia

ABSTRACT A high throughput trickling biofilter for odour control was designed basing on the principles of biotrickling filter technology developed in Moscow Bakh Institute of Biochemistry. All the necessary blocks except a fan: temperature and humidity control unit, a biofilter bed, an irrigation system, a control block and display unit are combined within one compact biofiltration module – a standard container 6000x2400x2400 mm. The plant is thermo-insulated that enables outdoor installation. The biofilter is easily scaled up by adding extra filtration beds. A typical biofiltration module rated for 5,000-10,000 m3/h has a contact time of 3-6 s (biofilter bed total volume – 10.5 m3) and a maximum footprint of 14.5 m2. After extensive pilot plant studies the first 5000 m3/h trickling biofilter easily scalable to 20000 m3/h was installed at Moscow Water Works in spring 2007 to control odour emissions - hydrogen sulfide, mercaptanes and other malodorous volatile organic compounds in up to 60 mg/m3 concentration. The performance results of the industrial biofilter are discussed.

1 INTRODUCTION Biofiltration – is an established technique to control odours (Devinny et al., 1999). Various types of biofilters have been suggested and successfully applied varying from the most simple open air units with soil/compost beds (www.bohnbiofilter.com) to sophisticated fully automated enclosed plants using proprietary artificial media and enabling full process control (Popov and Zhukov, 2005). The area of odour control is rather competitive and only most economic solutions have a chance of surviving on the market. Moreover deodouration is regarded less technically challenging than VOC 59

60

VITALY ZHUKOV, ANDREY VEPRITZKY, LEONID MITIN AND VLADIMIR POPOV

control enabling simple biofiltration plants to perform quite satisfactory and comply with existing legislation and end-user expectations. Trickling bed biofilters offer a number of advantages over conventional biological methods to treat off-gases. In some cases the contact time between the VOC laden air and the biocatalyst in trickling bed biofilters may be reduced to below 10 s range (Gabriel and Deshusses, 2003) thus minimizing the overall system footprint, dimensions and power requirements. For a number of years we are perfecting the design of the trickling biofilters (Zhukov et al., 1998; Popov and Bezborodov, 1999) and explore their potential in various applications: e.g. water soluble VOCs used in flexographic printing (Popov et al., 2004) , chlorinated compounds (Popov et al., 2005), formaldehyde (Popov et al., 2000), BTEX (Bezborodov et al., 1998), etc. Here we report an All-in-One high-throughput biotrickling filter for odour control applications and present its preliminary performance characteristics at Moscow Water Works.

2 MATERIALS AND METHODS 2.1 MICROORGANISMS As a base of the microbiological consortium used to populate the carrier in the biofilter the thionic bacterium Thiobacillus novellas has been used which efficiently degraded hydrogen sulfide. It was complimented by other strains from the in-house collection capable to utilise mercaptanes and volatile compounds present in the emissions of the water purification plants. All the strains used have been tested in a specialised certified laboratory properly authorised to perform such studies and were proved to be non-pathogenic, non-virulent and non-toxic for mammals. 2.2 CARRIER An inert polymer carrier with open-pore foam-like structure was used to immobilize the microbial consortium. 2.3 ANALYTICAL METHODS AND MONITORING Biofilter performance was monitored organoleptically and instrumentally. The personnel present on site evaluated the intensity of the smell at the outlet of the biofilter. The inlet and outlet concentrations of H2S (main source of the obnoxious odour) were routinely measured on-site electrochemically by a portable electrochemical analyzer Colion-1. The electrochemical cell analyzer was from time to time calibrated against a standard laboratory calorimetric procedure for quantifying SH-compounds. Prior to standard laboratory assay the SH-containing compounds were trapped by passing the predetermined volume of the air through the cartridges filled with glass spheres covered with solution of lead or mercury acetate.

HIGH THROUGHPUT BIOFILTRATION FOR ODOUR CONTROL AT WATER PURIFICATION PLANT

61

The assay of the gas mixtures was performed via GC/MS with Shimadzu GC2010 instrument equipped with GCMS-QP2010 Thermal Desorber (Markes International Ltd., UK) and capillary columns. The sample for GC/MS was obtained by adsorbing the volatiles by aspirating the controlled volume of air through a Tenax© column.

3 RESULTS AND DISCUSSION 3.1 CONSTRUCTION OF THE BIOFILTER To make biofilter more user-friendly and versatile a novel layout of the plant was used. An All-in-One principle was followed. All the main functional units of the plant: inlet air distribution and conditioning system; biofilter bed; irrigation and pHcontrol system; air transport system and droplet remover; control unit; etc. are enclosed within a standard 20-feet container (6000x2400x24000 mm) (Figure 1). The plant is properly thermo-insulated that enables outdoor installation. This is most important for the regions with a severe climate to which Russia evidently belongs to. When not in operation the plant can run in a stand-by mode maintaining its internal temperature regimes and thus precluding freezing or overheating.

Figure 1. Schematic outline of the All-in-One biotrickling filter.

62


The plant is very simple to start and operate. It requires just proper tap water and power supply and connection to the on-site existing drainage/sewage system. The air ducts are easily fitted to the inlet flange of the system. The system can be regarded of the type «install-and-run». The main parameters of the plant are shown in Table 1, while its general view and location on site appear in Figure 2. Table 1. Parameters of the biofiltration plant. Parameter and unit Nominal air flow, m3/h EBRT, s Optimal inlet concentrations of volatiles, mg/m3 Pressure drop, Pa Installed power, kWt Irrigation flow, m3/h Maximal air linear velocity, m/s Consumption of tap water (including evaporation), l/h Power consumption for outdoor installation (summer / winter), kWt*h Temperature of the incoming air, °C Allowed ambient temperature, °C Dimensions (HxLxW), mm Operating foot-print, m2 Dry weight, kg Wet weight (operating), kg

Value 4,000 – 11,000 3-10 < 250 350 – 1,250 < 18,5 < 12,5 < 0,70 < 110 < 3.5 / 15.5 0 – 50 (-20) – (+40) 2,475x6,300x2,400 14,5 < 3,700 < 10,500

The compact arrangement of all the components of the biofilter within one unit provides a number of technological advantages: easy transportation, erection and mounting; easy interfacing with the infrastructure existing on site (power, water, sewage, etc.); easy maintenance and service. The plant can be easily scaled up by adding either additional filtration beds or by adding extra complete units. 3.2 OPERATION OF THE BIOFILTER Main operation facilities of the sewage department of the Moscow Water Works are located rather close to the housing area. It is not feasible to relocate them, thus management of Moscow Water is looking for efficient and cost-effective technology that will enable to secure that neighbouring households do not complain about the irritating odours. An extensive programme of testing and pilot runs was launched that


63

enabled Moscow Water to select a trickling biofilter technology as the best available technology to control odours. The final phase of tests with a full-scale biofiltration plant, Figure 2, is currently in progress.

Figure 2. Biofilter at the premises of Moscow Water Works.

The biofilter was mounted close to the workshop for the residue sedimenting and dewatering that produces concentrated malodourous emissions. More than 70 components that varied in concentration quite considerably over time were identified in the emission by GC/MS technique. The dominant one was methane (1500-1600 mg/m3), while hydrogen sulfide – the major irritating pollutant was present in 12-16 mg/ m3 concentration with peaks up to 60-90 mg/ m3. The profile of the inlet and outlet H2S concentrations measured on-line with electrochemical sensor is presented in Figure 3. The results of the start-up and several months monitoring confirmed reliability of the technology. After about one month acclimation period the plant reached its target performance efficiency and is able to remove >98-99 % of the hydrogen sulfide coping with the spike emissions of about 30-80 mg/m3.

64


The instrumental GC – GC/MS assays showed that such pollutants as mercaptanes, amines, limonene, aromatic compounds were degraded with the efficiency of >99 % and could not be determined at the biofilter outlet (Table 3). Thioglycolic acid – one of the major pollutants with an average concentration of 5.5-6.0 mg/ m3 and peaks up to 10 mg/ m3 was usually degraded with efficiency of 80-99 %. Considerable depletion of the methane content - 40-70 % was also noted.

Figure 3. Start-up and performance of the trickling biofilter for odour control operating at Moscow Water Works. EBRT = 6 s.

The long-term (up to one year period) monitoring of the biotrickling filter plant is required to draw sound conclusions on the efficiency and the running costs. However already preliminary results show that the new All-in-One biotrickling filter could provide a viable solution to odour control at municipal water treatment plants.


65

Table 2. Efficiency of VOC degradation in the biofilter. Compound 1. Methane 2. Isobutyl amine 3. Ethyl sulfide 4. Thioacetic acid 5. Thioglycolic acid 6. Dioxane 7. Toluene 8. Ethylene diamine 9. Methyl mercaptane 10. Butyric acid 11. Dibutyl amine 12. Ethanol amine 13. Limonene 14. Dibutyl sulfide

Retention time, min 00:55 01:09 01:43 01:57 02:19 02:37 02:47 02:53 02:59 03:54 05:27 05:38 08:26 08:50

Peak area, Inlet 11226.2 72.32 5.89 211.80 86.32 22.84 32.53 27.11 21.65 6.67 8.50 11.24 13.19 3.61

arb.un. Outlet 2557.85 – – 47.54 – – – – – – – – – –

Conversion, % 77 >99 >99 78 >99 >99 >99 >99 >99 >99 >99 >99 >99 >99

4 ACKNOWLEDGEMENTS The project was supported in part by a grant from the Russian Federal Agency of Science and Innovations 02.447.11.3001. The authors thank management of Moscow Water, in particular Dr. Dmitry Danilovich and Mr. Michael Kozlov, for co-operation and good will.

REFERENCES Bezborodov, A.M., Rogojin, I.S., Ushakova, N.A., Kurlovitch, A.E. and Popov, V.O. (1998) Optimization of the microbiological technology for purification of air emissions from the mixture of benzene, toluene and xylene. Appl. Biochem. Microbiol.(Russia) 34: 265-269. Devinny, J.S., Deshusses, M.A. and Webster, T.S. (1999) Biofiltration for air pollution control. Boca Raton: Lewis Publishers. 299 p. Gabriel, D. and Deshusses, M.A. (2003) Retrofitting existing chemical scrubbers to biotrickling filters for H2S emission control. Proc. Natl. Acad. Sci. USA. 100: 6308-6312.

66


Popov, V, and Bezborodov, A. (1999) Industrial technology for microbiological abatement of air emissions. Appl. Biochem. Microbiol. (Russia) 35: 570-577. Popov, V.O., Bezborodov, A.M., Cross, P. and Murphy, A. (2000) Industrial trickling bed biofilters for abatement of VOCs from air emissions. In (ed. Ed.D.Wise e.a.) Remediation of Hazardous Waste Contaminated Soils, Marcel Decker Inc., 2nd ed., pp 449-473. Popov, V.O., Bezborodov, A.M., Cavanagh, M. and Cross, P. (2004) Evaluation of industrial biotrickling filter at the flexographic printing facility. Env. Prog. 100: 39-44. Popov, V., Khomenkov, V., Zhukov, V., Cavanagh, M. and Cross, P. (2005) Design, construction and long-term performance of novel type of industrial biotrickling filters for VOC and odour control. In: (Kennes C. and Veiga M.C., Eds), International Congress on Biotechniques for Air Pollution Control, La Coruòa, Spain, pp. 257–262. Popov, V.O. and Zhukov, V. (2005) Odour removal in industrial facilities. In (Shareefdeen Z. and Singh A., Eds.), Biotechnology for Odour and Air Pollution Control. Springer-Verlag, Heidelberg-New-York, Germany, pp. 305-326. Zhukov, V.G., Rogojin, I.S., Ushakova, N.A., Zagustina, N.A., Popov, V.O. and Bezborodov, A.M. (1998) Development of the technology of air deodouration and its field testing with the use of a pilot plant. Appl. Biochem. Microbiol. (Russia) 34: 370-376.

67

Continuous monitoring of odours at a biofilter outlet S ELENA S IRONI 1 , L AURA C APELLI 1 , P AOLO C ÉNTOLA 1 , R ENATO D EL R OSSO 1 , MASSIMILIANO IL GRANDE2 1

Olfactometric Laboratory, Department CMIC «Giulio Natta», Politecnico di Milano, Piazza Leonardo da Vinci 32, 20133 Milan, Italy 2 Progress S.r.l., Via N.A. Porpora 147, 20131 Milan, Italy

ABSTRACT The experience matured in the field of biofiltration applied to odour removal enabled to study the correlation between some of the characteristic parameters of the biofilter bed (e.g. T, RH, ΔP) with the emitted odour concentration. Today odour measurement can be performed only by classical olfactometry. Classical olfactometry is an expensive and time-consuming method that is not suitable for continuous monitoring as needed by operators of compost facilities. This paper describes the experimental approaches adopted for the development of a system for the continuous monitoring of odour emissions, i.e. an instrument for the repeated air analysis, capable of qualitatively and quantitatively recognizing odours. This work shows the results of the first experiments carried out with the purpose of developing an electronic nose to be applied at a biofilter outlet for the real-time odour concentration measurement and for the detection of the exceeding of a given odour «alarm threshold».

1 INTRODUCTION Biofiltration is a process commonly used for the removal of odorous compounds (Devinny et al., 1999). The biofilter efficiency depends on several key parameters such as the moisture content of the medium (Van Lith et al., 1997; Leson and Winer, 1991; Quiniam et al., 1999; Krailas and Pham, 2002; Morales et al., 2003), temperature (Yang and Allen, 1994) and pH (Devinny et al., 1999). Typical organic media may be 40-80% water (by weight) when saturated. The recommended water content is commonly evaluated at 50% of the water-holding

67

68

SELENA SIRONI, LAURA CAPELLI, PAOLO CÉNTOLA, RENATO DEL ROSSO

AND

M. IL GRANDE

capacity of the material. The rough surface and porosity provide extensive microbial habitats that may improve colonisation and bacterial diversity (Delhomenie et al., 2002). Most biofilters are colonised with mesophilic microbial communities, that is to say microbes that thrive at temperatures between 20 and 45°C (Darlington et al., 2001). In general for this temperature range, enzymatic activity increases as a function of temperature. Finally most biofilters are designed to operate in a near-neutral pH range (6-8) (Mac Nevin and Barford, 2001). Also the nature of the packing material is of great importance for the service life of the biofilter, microbial growth, removal performance and operational cost. Thus, the choice of biofilter medium (size of particles and their organic or inorganic nature) is a fundamental factor in the successful running of biofilters (Krailis et al., 2000; Tawil, 2001). Biofilters are widely used as odour abatement systems at different kinds of industrial plants. Odour control at the outlet of biofilters is very important, especially in the case of municipal solid waste (MSW) treatment plants, because odour emissions from this kind of plants are subject to limits fixed by the competent authority. The efficiency of biofilters in odour removal can be evaluated by monitoring the above mentioned parameters or by the direct analysis of odour concentration. Dynamic olfactometry is the technique commonly used for odour measurements at biofilters inlets and outlets, but olfactometric analyses are expensive and time consuming, and they can’t be carried out continuously. The aim of this paper is to describe the approaches adopted for developing an electronic nose suitable for the continuous monitoring of the odour concentration at a biofilter outlet, with the purpose of getting over the discontinuity bound to dynamic olfactometry. Nonetheless, the development of such a system requires a validation, which might be achieved by the execution of periodical olfactometric analyses for verifying the instrument accuracy in odour concentration determination.

2 MATERIALS AND METHODS 2.1 OLFACTOMETRIC ANALYSES The collection of samples to be analyzed by dynamic olfactometry on the biofilter outlet is carried out by means f a static hood, which has the function of isolating the sampling point from the external conditions, and to channel the air stream in a stack from where the sample is collected with a depression pump inside a NalophanTM bag with a TeflonTM inlet tube (Bockreis and Steinberg, 2005) (Figure 1).

CONTINUOUS MONITORING OF ODOURS AT A BIOFILTER OUTLET

69

The odour concentration of the samples collected at the biofilter outlet is determined by dynamic olfactometry, in conformity with the European Norm EN 13725:2003. The odour concentration is expressed in European odour units per cubic metre (ouE/m3), and it represents the number of dilutions with neutral air that are necessary to bring the concentration of an odorous sample to its odour perception threshold concentration. The analysis is carried out by presenting the sample to a group of selected panellists at increasing concentrations by decreasing serial dilutions, until the panel members perceive an odour that is different from the reference neutral air. The odour concentration is then calculated as geometric mean of the odour threshold values of each panellist, multiplied by a factor that depends on the olfactometer dilution step factor. An olfactometer ECOMA model TO8, based on the «yes/no» method, was used as dilution device (Figure 2). This instrument with aluminium casing has 4 panellists places in separate open boxes. Each box is equipped with a stainless steel sniffing port and a push-button for «yes» (odour threshold). The measuring range of the TO8 olfactometer goes from a minimum dilution factor of 4(=22) to a maximum dilution factor of 65536(=216), with a dilution step factor 2. All the measurements were conducted within 30 hours after sampling, relying on a panel composed by 4 panellists. The odour concentration was calculated as geometric mean of the odour threshold values of each panellist, multiplied by 2 .

Figure 1. Equipment for odour sampling at a biofilter outlet.

70


AND

M. IL GRANDE

Figure 2. The ECOMA TO8 olfactometer.

2.2 D EVELOPMENT

OF A SYSTEM FOR THE CONTINUOUS ODOUR CONCENTRATION

DETERMINATION

2.2.1 GENERAL PRINCIPLES The development of this system is based on the use of an already existing technology, i.e. the electronic nose. The first prototype of electronic nose was described by Persaud and Dodd in 1982, and it consists of an instrument which comprises an array of electrochemical sensors with partial specificity and an appropriate pattern recognition system, capable of recognizing simple or complex odours. Even though electronic noses are studied since several years, most studies concern applications that are very different from the one proposed in this work. For these reasons, a complete re-design of the instrument is required in order to make it suitable for the specificities associated with the application in the environmental sector. The work required for the development of a similar system is composed of two fundamental activities, which are strictly interconnected, i.e. the instrument design and the definition of its utilization procedures. The instrument design comprises the following aspects: the choice of the gas sensors; the implementation of a suitable software for the instrument operation and the data acquisition and processing, and finally the study of technical features which are needed in order to make the instrument usable not only in laboratory but also in the field. Parallel to the instrument design activity it is extremely important to define the electronic nose utilization procedures.


71

The use of an electronic nose provides to relate an unknown «match data set» to a «training data set» acquired by the instrument in the so-called «training» phase, during which a number of odorous samples are analyzed in order to create the database of odour concentration values used as a reference by the instrument for further odour concentration estimation. The principal utilization procedures to be defined concern therefore the instrument settings, the training methods and the data processing methods. 2.2.2 ELECTRONIC NOSE DESCRIPTION The instruments used for this study have been developed in collaboration with Sacmi s.c.a.r.l. and the Sensor Laboratory of the University of Brescia (Falasconi et al., 2005) (Figure 3).

Figure 3. Electronic nose used for the study.

The system includes a pneumatic assembly for dynamic sampling (pump, electro-valve, and electronic flow meter), a thermally controlled sensor chamber with 35 cm3 of internal volume and an electronic board for controlling the sensor operational conditions. The electronic noses have been equipped with an array of six thin film MOS (Metal Oxide Semiconductor) sensors, which make the system sensitive to a large spectrum of volatile compounds, and a humidity sensor. For the analyses, the carrier flow rate was 150 cm3 min-1 and the temperature of the sensor chamber was kept constant at 50°C.

72


AND

M. IL GRANDE

2.2.3 TRAINING FOR ODOUR QUANTIFICATION The quantification of odour concentration by means of an electronic nose requires a particular training: odorous gas samples must be diluted at different odour concentration values and analyzed in order to create a database that can be used for the estimation of the odour concentration of unknown air samples by interpolation of the training points. For this reason, after collection, the odour concentration of the samples is measured by dynamic olfactometry. Once the odour concentration of the samples is known, each sample is diluted at different odour concentration values, by means of the same dilution device used for the olfactometric analyses (olfactometer), in order to obtain samples with odour concentration values included in the typical odour concentration range of odorous ambient air (20-1000 ouE m-3) (Sironi et al., 2007). 2.2.4 INSTRUMENTAL SENSITIVITY TOWARDS ODOUR CONCENTRATION VARIATIONS In order to develop a system for the continuous monitoring of odours at a biofilter outlet it is necessary to evaluate its sensitivity towards odour concentration variations. For this reason, we are currently carrying out a set of experimental tests, in order to evaluate how different sensors respond at the analysis of odorous gas samples at different odour concentration values. The instrumental sensitivity is tested using different typologies of odours: pure compounds and gas samples collected at biofilter outlets. As pure compounds, we decided to consider ammonia (NH3) and hydrogen sulphide (H2S), because these are the compounds for which an concentration limit is fixed by the competent authority at the outlet of biofilters installed at plants for the mechanical biological treatment of MSW.

3 RESULTS AND DISCUSSION This paragraph reports the results of some of the studies conducted with the aim of developing an electronic nose suitable for the continuous monitoring of odours at a biofilter outlet. More in detail, the results reported concern the studies relevant to the experimental verification of the instrumental sensitivity towards odour concentration variations. It is worth to remember that this results are preliminary and therefore partial, because our work in this field is still in progress. As mentioned in the previous paragraph, the sensitivity of different sensors is tested using samples of pure compounds and of gaseous mixtures collected at biofilter outlets. As an example, we consider the tests conducted with a set of samples at different odour concentration values (22, 34, 44 and 110 ouE/m3) obtained by dilution of a


73

sample collected at the outlet of a biofilter installed at a plant for the mechanical and biological treatment of MSW. Figure 4 shows the responses of a Nickel Oxide sensor (p-type) relevant to the analyses of the above mentioned samples. It is possible to observe that the amplitude of the response curves increases with the odour concentration values. Based on this observations it is possible to affirm that the studied sensor is sensitive even to very small variations of the odour concentration (i.e. 10 ouE/m3).

Figure 4. Response curves relevant to the analysis of gas samples at different odour concentration values.

Figure 5 shows a PCA relevant to the same test. It is possible to observe that the odour concentration values grow along the dircetion of the first principal component.

74


AND

M. IL GRANDE

Figure 5. PCA relevant to the analysis of gas samples at different odour concentration values.

The data relevant to the analyses of the samples at different odour concentration values obtained by dilution of a sample collected at a biofilter outlet can be splitted in two, a test training set composed by 5 measures, and a test match set composed by the remaining 5 measures. The obtained test training set is used as a reference for the estimation of the odour concentration values of the test match set, in order to evaluate the instrument accuracy in odour quantification. Table 1 shows the results of this test, i.e. the odour concentration values attributed by the instrument to the measures that form that test match set. The extrapolation of statistical data from the test results, e.g. mean, maximum and minimum error, and correlation index between true and estimated values, allows to gain some information about the estimation accuracy (Table 2). The calculated correlation index is equal to 0.99923, indicating a very good correlation between estimated and true and estimated values.

75


Table 1. Results of the odour concentration estimation test. N. 1 2 3 4 5

Match Value 19.46 35.25 32.31 43.41 109.48

True Value 22 34 34 44 110

Error -2.54 1.25 -1.69 -0.59 -0.52

Table 2. Statistical parameters relevant to the odour concentration estimation test. Mean absolute error Max error Min error Correlation index

1.3173 2.5398 0.51962 0.99923

4 CONCLUSIONS Our current work consists in the execution of a large number of tests with the aim of studying the optimal combination of sensors for the development of an electronic nose for the continuous monitoring of odours at a biofilter outlet. The most important characteristics for such a system are high sensitivity and accuracy. The studies we already conducted allowed to exclude some sensors, with too low sensitivity or stability, considering that sensors to be applied for real-time odour concentration measurements should be very sensitive even to very small odour concentration variations ( 1 month) use of hexadecane as a benzene delivery phase for A. xylosoxidans Y234 has been amply demonstrated (Nielsen et al., 2005a). Styrenebutadiene (28% styrene) ABA co-polymer beads (with dimensions of approximately L=4.25 mm, D=3.75 mm and density of 0.94 g/mL) were used as toluene-sequestering phase in the solid-liquid TPPB. For oxygen transfer experiments nylon 6,6, glass beads and silicone rubber beads were used in addition to the styrene-butadiene (SB) beads. 2.2 MICROORGANISM, MEDIUM, AND CULTURE CONDITIONS Achromobacter xylosoxidans Y234 is known to have the ability to degrade toluene, benzene and m-xylene. The growth medium (Davidson and Daugulis, 2003) was: 7 g/L (NH4)2SO4, 0.75 g/L MgSO4•7H2O, 6.6 g/L K2HPO4, 8.42 g/L KH2PO4, 2 g/L sodium benzoate, and 1 mL/L trace elements. Eight 125 mL Erlenmeyer shake flasks containing 50 mL of medium were inoculated from frozen stock prior to incubation at 30oC and 150 rpm for 24 hours in preparation for their inoculation in the bioreactor. 2.3 REACTOR SET-UP AND OPERATION A 5 L New Brunswick Scientific BioFlo III was set to operate at 30oC, a pH of 6.6, an agitation speed of 800 rpm and a total working volume of 3 liters. The aqueous medium consisted of 14 g/L (NH4)2SO4, 1.5 g/L MgSO4×7H20, 13.2 g/L K2HPO4, 16.84 g/L KH2PO4 and 0.16mL/L trace elements. For the fermentations conducted using liquid n-hexadecane as the second phase, an organic fraction of 0.33 was used with the remainder being the aqueous, cell-containing phase. The two phases were maintained as a dispersion through agitation. For the fermentations conducted using SB beads as a second phase, 500g of polymer beads were used with 2.518L of aqueous medium added for a final total volume of 3L. Higher bead fractions were found to result in excessive build up of beads behind baffles and other reactor internals, due to the relatively large size of the beads. The toluene delivery system consisted of an Erlenmeyer flask with 2L of toluene and a regulated amount of compressed air being sparged through it that continued into the reactor. A water bath kept the flask at the 30 oC. This air stream was mixed with air for bioreactor aeration and this combined stream was delivered into the reactor through a sparger. Dissolved oxygen levels were measured with a polarographic-membrane electrode. Concentrated nutrient boluses were added periodically (every 2-3 days) to

430

ANDREW J. DAUGULIS, JENNIFER V. LITTLEJOHNS

AND

NEAL G. BOUDREAU

the reactor to ensure that the system was not nutrient limited. A small amount of Sigma Antifoam 289 (~0.5mL) was added as required. 2.4 ANALYTICS Liquid samples, to measure biomass concentration, were centrifuged, the liquid supernatant was discarded and the biomass was then re-suspended in deionized water. Appropriate dilutions were then performed and measured at 650 nm and compared to a previously determined calibration curve. Inlet and outlet gas samples were taken by means of a gas tight 250 μL syringe. A Perkin Elmer AutoSystem Gas Chromatograph fitted with a flame-ionizing detector and a fused silica capillary column (DB-5, 0.53mm I.D., 30m length, 1mm film thickness, Model 125-503J, J & W Scientific, Inc., Folsom, CA) was used to analyze toluene concentrations. The aqueous phase toluene concentration was calculated based on Henry’s Law between air and the aqueous medium previously found to be 0.247 (mg/L)gas/(mg/L) aq. Toluene concentrations in n-hexadecane and SB beads were determined based on partition coefficients relative to the aqueous medium as determined previously. 2.5 STEADY STATE AND TRANSIENT OPERATION Immediately after inoculation a total flow rate of 1.71L/min air (0.57vvm) at a toluene concentration of 10 mg/L was established for a loading rate of 343 g/m3.h. This loading rate was maintained during the biomass growth phase and between dynamic step experiments. The cell growth reached a steady state in each case within 5 to 7 days at a cell mass in the bioreactor of between 20 and 25 grams (CDW). Achieving a steady state biomass concentration even with continued addition of substrate is due to the use of the consumed substrate for cell maintenance purposes only, rather than cell growth. All transient experiments were performed once the biomass levels had stabilized after the initial 5 to 7 day growth period. Inlet toluene steps were introduced for periods of 60 minutes by varying the proportions of air passing through the toluene flask and the aeration air, after which the toluene loading was reduced to its initial level. The size of the step was normalized with respect to total cell mass present to ensure that the performance of each system was not affected by the amount of biomass present. The step was imposed for each bioreactor configuration above the steady state feeding condition at a loading of approximately 110 (gToluene/m3reactor.h)/(g-cells). Alternatively, from a stable loading of 343 g/m3.h, steps of approximately 2400 g/m3.h were performed, after which the flows were readjusted to their original set-points and the inlet and outlet toluene levels were monitored until the instantaneous removal efficiency of the system returned to its original steady state value.

SOLID-LIQUID TWO-PHASE PARTITIONING BIOREACTORS

431

2.6 OXYGEN TRANSFER EXPERIMENTS The same New Brunswick Bioflo III operating at 30 °C was used for oxygen transfer experiments. All systems, which consisted of either tap water or 500 g of either glass beads, nylon 6,6, styrene–butadiene copolymer (SB) or silicone rubber in tap water to a total working volume of 3 L, were operated at aeration rates of 0.5 L/min, 0.75 L/min and 1 L/min and agitation rates of 100-800 rpm. Mass transfer measurements were taken in duplicate with the average value being reported, and the effect of probe response was also incorporated in the analysis. All solid particles used were approximately spherical in shape. The unsteady-state method described by Shuler and Kargi (2002) was used to determine mass transfer coefficients that reflected either the presence of inert particles (e.g. glass beads and nylon 6,6) alone, or the presence of particles that also possessed oxygen affinity (e.g styrene–butadiene or silicone rubber). The effect of the presence of inert particles was «separated» from the combined effect of «presence + affinity» using mathematical analysis as previously described (Littlejohns and Daugulis, 2007) to isolate the oxygen transfer enhancement that can be obtained in the presence of polymer beads possessing oxygen affinity.

3 RESULTS AND DISCUSSION 3.1 TPPB OPERATION Shortly (< two days) after inoculation, the removal efficiencies increased to greater than 95% and at a toluene loading of 343 g/m3.h, the cells reached steady state masses ranging between 19.4 and 26.2 grams (CDW) within 7 days. The removal efficiencies of the systems remained greater than 95% for the entirety of the experiments except during transients. Recent work by Nielsen et al. (2005a) has shown that a constant cell concentration will eventually be established due to cellular maintenance requirements, which are responsible for all of the substrate consumed. Figures 1-3 show the transient responses when the loading (~110 (g/m3.h)/ (g-cells)) was increased from its nominal rate of 343 g/m3.h to approximately 2400 g/ m3.h for a period of 60 minutes. The instantaneous removal efficiencies of the singlephase system and the polymer phase system both dropped immediately upon the onset of the step reaching minimum values of 57 and 87%, respectively before the end of the 60 minute step. This is also reflected in the outlet toluene concentrations which reached 20 and 10 mg/L, respectively. The instantaneous removal efficiencies of the n-hexadecane as a second phase system remained above 95% for the entirety of the 60 minute step, and in fact increased at the initial stages of the transient, reflecting absorption of the higher toluene loading. Outlet toluene concentrations remained low for this system, reaching just 2 mg/L, or about one-tenth of the single-phase system.

432


AND

NEAL G. BOUDREAU

Figure 1. Removal efficiency (triangles) and exit toluene concentration (squares) for single-phase system.

The solvent as a second phase system removed 97% of the toluene fed to the system over the course of the 200 minute experiment (Table 1), with the polymer phase system removing 90% of the toluene, and the single phase clearly performing the worst of the three systems removing only 69%. The performance comparison between the systems with second phases may not be entirely fair, however, given the different masses of second phases that were used. It can be anticipated that as more polymer phase is used (approaching the mass of n-hexadecane) the performance of this system would be closer to the two-liquid phase system. A comparison of the DO traces of the three systems (Figure 4) shows that DO for the solvent system remained the highest reaching a minimum DO value of 48% of saturation, while that of the polymer system was intermediate (33% of saturation), and the DO for the system with no second phase dropped to the lowest level (10% of saturation). The higher level of oxygen in the n-hexadecane case may be expected due to the greater capacity for oxygen by this solvent (Nielsen et al., 2003), and it is also interesting to see that the SB beads had a similar effect, albeit to a lesser degree with the mass of beads used in this case. Thus the presence of a second phase, originally intended to absorb and sequester toxic VOC substrates, has the added beneficial effect of enhanced oxygen absorption and release.


Figure 2. Removal efficiency (triangles) and exit toluene concentration (squares) for solvent-phase system.

Figure 3. Removal efficiency (triangles) and exit toluene concentration (squares) for polymer-phase system.

433

434


AND

NEAL G. BOUDREAU

Figure 4. DO profiles for single (circles), polymer (triangles), and solvent (squares) systems.

Table 1. Performance summary during imposed step transients. Reactor Type

95% Recovery Time (Minutes)

DO Recovery Time (Minutes)

Toluene Released During Step (mg)

Toluene Released After Step (mg)

Total Toluene Released (mg)

Overall Removal Efficiency (%)

Single -phase Solvent Polymer

30

60

3002

292

3294

69

63 48

115 162

116 864

198 165

314 1030

97 90

3.2 OXYGEN TRANSFER In light of the enhanced oxygen transfer seen in the polymer TPPB, the effect of the presence of polymer beads on O2 transfer was examined in more detail as described elsewhere (Littlejohns and Daugulis, 2007). The presence of the beads could have 2 effects on O2 uptake: a physical effect arising from their mere presence, and an absorptive effect in which O2 is actually taken up by the polymer. In order to examine this in detail, polymers with negligible O2 affinity (e.g. nylon 6,6 with a diffusivity of 1.6 x 10-9 cm2/s), and polymers with substantial O2 affinity (SB with a diffusivity of


435

Figure 5. Mass transfer coefficients at 400 rpm for nylon 6,6 (square), glass beads (circle), water (triangle and line), silicone rubber (square) and styrene–butadiene copolymer (star).

Figure 6. Oxygen transfer rate between 30% and 80% of liquid saturation by a system of water with silicone beads (circles) and water without particles (squares).

436


AND

NEAL G. BOUDREAU

1.4 x 10-6 cm2/s, and silicone rubber with a diffusivity of 3.4 x 10-5 cm2/s) were tested along with glass beads with essentially zero O2 affinity as a control. The presence of beads on both the kLa and on the oxygen transfer rate (OTR) was examined. The volumetric mass transfer coefficients (kLa) for aqueous systems with and without particles are shown in Figure 5 for different aeration rates at 400 rpm. The kLa values are up to 55% lower for the system with SB beads relative to the system without particles and up to 63% lower for the system containing silicone rubber beads, which at first seems counter-intuitive, given the earlier TPPB results. For systems containing SB and silicone rubber, the measurement of mass transfer coefficient contains the effect that the solid polymer particles may have on kLa, as well as any effects that they may have on absorbing oxygen. The system containing nylon 6,6 shows up to a 268% increase in kLa. Due to the low oxygen diffusion coefficient of nylon 6,6, the effect of the nylon particles on the kLa is isolated, and mass transfer enhancement due to the mere presence of particles alone is clearly observed. In a similar manner to nylon 6,6, glass beads are inert and enhance the kLa up to 159%. Both nylon 6,6 and silicone rubber have very similar dimensions and densities, which have been identified earlier as critical factors for the effect of particles on kLa. Nylon 6,6, can therefore be used to approximate the effect of the presence of silicone rubber beads on kLa, as both the effects of oxygen absorption by the silicone rubber and the effects on the gas–liquid mass transfer are contained within the measured kLa for the silicone rubber system and cannot be separated. In order to demonstrate a larger overall uptake of oxygen into a TPPB system relative to a system without a second phase, the instantaneous oxygen transfer rate (OTR) as a function of time at 400 rpm agitation and 1 L/min aeration is shown in Fig. 5. This plot, which was generated by mathematically «separating» the kLa effect from the overall observed measurements, clearly shows that between 30% and 80% of liquid saturation the system containing silicone rubber beads has a much larger OTR during the progression to liquid saturation than the system without a second phase. As well, the system with a second phase reaches 80% liquid saturation much later than the system without a second phase. This is due to the polymers acting as an oxygen sink within the system, in turn causing the liquid oxygen concentration to be lower relative to the system without polymers, at any given time. This decrease in the liquid concentration causes an increased driving force for oxygen between the gas and liquid phases, which causes a larger oxygen transfer rate for an extended period of time. Therefore, although the kLa is measured as lower for the reasons explained above, the overall oxygen transfer rate into the solid–liquid system is larger for systems containing particles with oxygen affinity (e.g. silicone rubber or SB). This is due to the oxygen transfer rate not only being proportional to the volumetric mass transfer coefficient, but also to the increased instantaneous concentration driving force. The SB and silicone polymers have a large uptake of oxygen and therefore more oxygen can ultimately be contained within the


437

system. These results are comparable to those for liquid–liquid systems that have found that oxygen is transferred at a higher rate due to an increased driving force (Nielsen et al., 2005b). However liquid–liquid systems increase the working volume oxygen saturation concentration, whereas the solid–liquid system increases the driving force by decreasing the liquid concentration at any given time, as well as by enhancing gas–liquid mass transfer. Nevertheless, both liquid–liquid and solid–liquid systems can increase the overall amount of oxygen that can be contained within a working volume.

4 CONCLUSION The presence of a second immiscible phase, whether a liquid, or a carefully selected polymer, has been shown to significantly improve the ability of bioreactors to enhance VOC removal. Although data from the use of a pure strain have confirmed this in the present work, similar results have also been obtained with microbial consortia operating in polymer-based TPPBs. In addition, the presence of these materials can have a significant positive effect on O2 transfer, which can be critical during dynamic periods of operation treating VOC surges. Adding such polymeric materials to more conventional biotreatment devices (e.g. biofilters) may also provide similar positive benefits on VOC buffering and removal, and on O2 transfer.

REFERENCES Aizpuru, A., Khammar, N., Malhautier, L. and Fanlo, J.L. (2003) Biofiltration for the treatment of complex mixtures of voc influence of the packing material. Acta Biotechnology. 23: 211-226. Amsden, B.G., Bochanysz, J. and Daugulis, A.J. (2003) Degradation of xenobiotics in a partitioning bioreactor in which the partitioning phase is a polymer. Biotechnol. Bioeng. 84: 399-405. Davidson, C.T. and Daugulis, A.J. (2003) Addressing biofilter limitations: A two-phase partitioning bioreactor process for the treatment of benzene and toluene contaminated gas streams. Biodegradation. 14: 415-421. Littlejohns, J.V. and Daugulis, A.J. (2007) Oxygen transfer in a gas-liquid system containing solids of varying oxygen affinity. Chem. Eng. J. 129: 67-74. Nielsen, D.R., Daugulis, A.J. and McLellan, P.J. (2003) A novel method of simulating oxygen mass transfer in two-phase partitioning bioreactors. Biotechnol. Bioeng. 83: 735-742.

438


AND

NEAL G. BOUDREAU

Nielsen, D.R., Daugulis, A.J. and McLellan, P.J. (2005a) Quantifying maintenance requirements from the steady-state operation of a two-phase partitioning bioscrubber. Biotechnol. Bioeng. 90: 248-258. Nielsen, D.R., Daugulis, A.J. and McLellan, P.J. (2005b) A Restructured framework for modeling oxygen transfer in two-phase partitioning bioreactors. Biotechnol. Bioeng. 91: 773-777. Prpich, G.P. and Daugulis, A.J. (2004) Polymer development for enhanced delivery of phenol in a solid-liquid two-phase partitioning bioreactor. Biotechnol. Progr. 20: 1725-1732. Prpich, G.P. and Daugulis, A.J. (2005) Enhanced biodegradation of phenol by a microbial consortium in a solid-liquid two phase partitioning bioreactor. Biodegradation. 16: 329-339. Shuler, M.L. and Kargi, F. (2002) Bioprocess Engineering, 2nd ed., Prentice Hall, New Jersey. Tang, H.M., Hwang, S-J and Hwang, S-C. (1995) Dynamics of toluene degradation in biofilters. Hazard Waste Hazard Mater. 12: 207-219. Weber, F.J. and Hartmans, S. (1995) Use of activated carbon as a buffer in biofiltration of waste gases with fluctuating concentrations of toluene. Appl. Microbiol. Biotechnol. 43: 365-369.

439

Microbial solvent regeneration in biotreatment of air contaminated by styrene ERIC DUMONT AND YVES ANDRÈS UMR CNRS 6144 GEPEA, Ecole des Mines de Nantes, 4 rue Alfred Kastler, BP 20722, 44307 Nantes Cedex 03, France

ABSTRACT In this study, the biodegradation of a styrene-polluted waste gas in a reactor containing 5 L of a biphasic mixture (10:90% v/v) of organic solvent (silicone oil) – water – biomass was investigated to establish the ability of a microbial solvent regeneration. Reproducible microbial solvent regenerations have been observed. The regeneration time, which increases with the increase of the styrene load (varying from 543 to 1800 mg), leads to elimination capacity up to 48 gstyrene.m-3mixture.h-1. The solvent regeneration requires roughly 1.2 molecule of oxygen per molecule of styrene and corresponds to the first steps of the biodegradation of the styrene.

1 INTRODUCTION Styrene is of major importance in the petrochemical and polymer-processing industries, which can contribute to the pollution of natural resources via the release of styrenecontaminated effluents and off-gases. Generally the produced polluted air flows are high in volume with low styrene concentrations (around 200 mg.m-3) corresponding to the application area of bioprocesses and some biofilter utilizations are described (Cox et al., 1996; Arnold et al., 1997; Jorio et al., 2000; Zilli et al., 2001; Dehghanzadeh et al., 2005) but with the needing of high bed volume. Another possibility in biological air treatment is the use of a bioscrubber, combination of a column for the pollutant air to liquid transfer and a biological reactor for the solvent recycling. However, for waste gases containing hydrophobic compounds having low solubility in water (like styrene), it is necessary to use a mixture of a non-biodegradable organic solvent and water. The pollutant is preferentially transferred from the gas phase to the organic solvent, and

439

440

ERIC DUMONT and YVES ANDRÈS

diffuses to the aqueous phase in order to be degraded by the microorganisms. Thus complete regeneration of the mixture of organic solvent – water is possible. There are only few studies devoted to the regeneration of the organic solvent for styrene biodegradation (see Dumont et al., 2006a). Consequently, the aim of this study is to present preliminary results about a solvent regeneration due to the degradation of the styrene by a mixed culture able to use this molecule as the unique source of carbon. Silicone oil is specifically used as solvent and the regeneration is followed by the measure of the oxygen variation in the gas phase in a batch reactor.

2. MATERIALS AND METHOD 2.1 MICROORGANISMS AND CHEMICAL The mixed culture was obtained from the Nantes (France) wastewater treatment plant. Styrene was used as the sole source of carbon and energy and a nutrient solution consisting in an aqueous solution of H8N2O4S and H2KO4P was used according to the quantity of carbon to keep the C/N/P ratio around 100/5/1. The biomass was progressively acclimated to styrene. Silicone oil (Rhodorsil fluid 47V5, dimethylpolysiloxane) was obtained from Rhodia Company. The physical properties at 25°C are: viscosity, 5 mPa.s; density, 910 kg.m-3; styrene solubility, 38g/L (Dumont et al., 2006b); oxygen solubility, 7 higher than in water (Dumont et al., 2006a). The styrene solubility in water at 25°C is 320 mg/L (Kirk-Othmer, 1983). 2.2. EXPERIMENTAL SETUP The description of the reactor used in thus study is shown in Figure 1. The reactor has an 11.5 L total volume (height 0.33 m, diameter 0.21 m). In the experiments, air was supplied from a compressor and sparged through an elliptical distributor (75x150mm) with 50 holes (1mm diameter). All experiments were carried out at a constant temperature of 25°C maintained by a thermostatic bath. The total volume of the mixture was 5 L (silicone oil volume fraction: 10%) allowing to absorb roughly 3g of styrene (Dumont et al., 2006b). Each experiment was carried out according to the following sequential procedure: 1) A synthetic waste gas stream was prepared by passing compressed air through a styrene generator filled with liquid styrene (flowmeters 3 and 4 opened; valves 1 and 5 opened; valve 2 closed). The known styrene stream bubbled through the mixture silicone oil – water – biomass during 1 hour in order to be absorbed by the liquids. The styrene concentrations in the gas phase were measured simultaneously at the inlet and outlet of the reactor using a

MICROBIAL SOLVENT REGENERATION IN BIOTREATMENT OF AIR CONTAMINATED

441

Figure 1. Scheme of the laboratory scale multiphase reactor.

Flame Ionisation Detector (Combustion HFR 400 FFID) calibrated before each experiments with standards. From these measures the amount of styrene absorbed by liquids is known. 2) The system was then completely isolated from the outside air by closing the valves 1 and 5. By closing the flowmeter 4 and opening the valve 2, a finite volume of air (6.7 L) was continuously flowed through the mixture in order to supply the biomass in oxygen (gas flow rate 0.9 10-4 m3/s). The decrease in oxygen concentration in air due to the biodegradation of the styrene by microorganisms was monitored and recorded as a function of time during 23h for further analysis. The oxygen fraction in the gas phase was determined using a paramagnetic oxygen analyser (Cosma Cristal 300). It was assumed that: (i) in the reactor the ideal gas law is applicable to calculate the number of moles of oxygen absorbed by the liquids (temperature and pressure of the gas phase was low), (ii) the presence of silicone oil in the emulsion does not change Henry constant for oxygen in water, (iii) the response time for the paramagnetic oxygen analyser is 99%), was purchased from Merck Limited, India. 2.2 MICROBIAL CULTURE AND MEDIA COMPOSITION The microbial mixed consortium was obtained from a municipal sewage treatment plant. The final inoculum was obtained by series of repeated inoculation in a mineral salt medium (MSM) that had the following composition: (NH4)2SO4 – 2.0 g/L, KH2PO4 – 2.0 g/L, NaCl – 0.5 g/L, MgSO4·7H2O – 0.025 g/L (Krausova, et al., 2003). The pH of the mineral salt media was adjusted to (7±1). 2.3 EXPERIMENTAL The schematic of the experimental setup is given in Figure 1. The VOC vapor generated by passing air at controlled rate through a VOC reservoir was mixed with humidified air in a glass mixing chamber to obtain the desired concentration of the feed gas stream. The biofilter consisted of a poly-acrylic tube (5’70 cm) having 6 sampling ports sealed with a rubber septa at 10 cm interval along the biofilter height. It was loosely packed with a mixture of compost and polystyrene spheres. A perforated plate at the bottom provided the support for the packing and also ensured uniform distribution of the vapor stream. The moisture content of the filter bed was maintained by prehumidifying the incoming vapor and by periodically sprinkling fresh media from the top of the biofilter. The modified RBC was made from two semi-cylindrical poly-acrylic parts with flanges to get a leak free operation. It contained 20 poly-acrylic discs each of 21 cm diameter, 5 mm thickness with 10 mm spacing between the discs. The discs were mounted on a stainless steel shaft and rotated slowly with variable speed motor. Baffles fixed to the top cover allowed the gas flow to pass through each chamber and increase the residence time of the gas in the reactor. The experimental runs were performed by passing the VOC vapor at different concentration and flow rates to vary the influent organic load. Samples were collected at the inlet and outlet at different intervals using gas tight syringe and analyzed for residual VOC concentration. 2.4 ANALYTICAL METHODS Residual concentration of DCM in the aqueous phase was measured in cell free samples using a Gas Chromatograph (AIMIL, India) fitted with a flame ionization detector and a (10% FFAP) packed column. Nitrogen was used as the carrier gas at a

472

R. RAVI, LIGY PHILIP AND T. SWAMINATHAN

Figure 1. Schematic of the experimental setup.

flow rate of 25 mL/min. The temperatures of injector, oven and detector were 200, 120 and 220°C respectively and retention time was 1.4 mins.

3 RESULTS AND DISCUSSION 3.1 STARTUP AND ACCLIMATIZATION OF BIOFILTER AND ROTATING BIOLOGICAL CONTACTOR (RBC) Startup studies were initiated with both biofilter and RBC in order to develop the required biofilm acclimatized to DCM as the main carbon source.

DICHLOROMETHANE REMOVAL USING MIXED CULTURES IN A BIOFILTER

473

3.1.1 BIOFILTER STARTUP Through a diverse microbial population exist in the compost, an external microbial consortium acclimatized to DCM in batch culture was added to the biofilter. Acclimatization is critical for successful operation of any biofilter and it depends on the activity of the microorganisms present in the compost.The biofilter was initially operated at an inlet concentration of 0.15 – 0.21 g/m3 and at a gas flow rate of 0.06 m3/h corresponding to an empty bed residence time (EMBRT) of 1.47 mins. The removal efficiency profiles were monitored continuously for 90 days till they achieved steady state. Figure 2 and 3 show the concentration profile of DCM and its removal efficiency in the biofilter. The removal efficiency during the first few days was high, which could be mainly due to the absorptive and adsorptive capacity of the compost. After this initial sorption phase, the biofilter efficiency decreased to 49%. However, there was a slow gain in the efficiency after 4 days. Zhu et al. (1998) have reported similar observation during the treatment of benzene vapors in biofilters containing compost and granular activated carbon as the packing material. Hodge and Devinny (1994) and Arulneyam and Swaminathan (2004) have also reported similar adsorption and biodegradation phases during startup operation in biofilters handling ethanol and methanol vapours. The removal mechanism shifted towards biodegradation and during this phase, the removal efficiency progressively increased with fluctuations, until becoming constant at about 88%.

Figure 2. Startup operation of the biofilter treating DCM vapors.

474


Figure 3. Removal efficiency of the biofilter treating DCM vapors.

3.1.2 STARTUP WITH RBC The RBC was initially inoculated with a seed culture acclimatized to DCM and operated at an inlet concentration of 0.25 – 0.3 g/m3 at a gas flow rate 0.06 m3/h for10 days. It was observed that fairly thick and uniform biofilm was formed on the disc surface. As the removal efficiency was observed to be low, the DCM concentration in the feed was reduced to 0.15 – 0.2 g/m3 and the study was continued. The removal efficiency profiles were monitored continuously till they achieved steady state. Figures 4 and 5 show the concentration profile and the removal efficiency in RBC. The removal efficiency increased from 19% to 54% when the RBC was operated at high concentration. After decreasing the concentration, the removal efficiency increased rapidly and reached about 85% in 32 -35 days. The shorter time to achieve the steady state may be due to the thick growth of biofilm in the RBC. The total biomass in the RBC may be more than in the biofilter. The fact that such a thick biofilm can develop in RBC with DCM as the only carbon source proves the potential application of RBC for VOC removal.

4 CONCLUSIONS The RBC system was found to be a potentially attractive configuration for gas phase biodegradation of VOC. For comparative organic load, both RBC and biofilter gave reasonably high (>80 %) removal of DCM, but RBC was able to reach it in shorter time.

DICHLOROMETHANE REMOVAL USING MIXED CULTURES IN A BIOFILTER

475

Figure 4. Startup operation of the RBC treating DCM vapors.

Figure 5. Removal efficiency RBC treating DCM vapors.

REFERENCES Arulneyam, D. and Swaminathan, T. (2004) Biodegradation of mixture of VOC’s in a biofilter. J. Environ. Sci. 16(1): 30-33. Brunner, W., Staub, D. and Leisinger, T. (1980) Bacterial degradation of dichloromethane. Appl. Environ. Microbiol. 40: 950-958.

476


Diks, R., Ottengraf, S.P.P. and Vanden Oever, A.H.C. (1994) The inflence of NaCl on the degradation rate of dichloromethane by Hyphomicrobium sp. Biodegradation 5: 129-141. Green, T. (1991) Human exposure. Toxicol. 16: 2981-2985. Herbst, B. and Wiseman, U. (1996) Kinetics and reaction engineering aspects of the biodegradaton of dichloromethane and dichloroethane. Wat. Res. 30(5): 1069-1076. Hodge, D.S. and Devinny, J.S. (1994) Biofilter treatment of ethanol vapours. Environ. Prog. 13: 167-173. Wang, J. and Chen, J.M. (2006) A Removal of dichloromethane from waste gases with a biocontact oxidation reactor. Chemical Eng. J. 123(3): 103-107. Krausova, V.I., Robb, F.T. and Gonzalez, J.M. (2003) Bacterial degradation of dichloromethane in cultures and natural enviroments. J. Microbiol. Methods 54: 419-422. Kennes, C. and Veiga, M.C. (2001) Conventional Biofilters. In: Bioreactors for Waste Gas Treatment, Kennes, C. and M. C. Veiga (Eds.) Kluwer Academic Publisher, Dordrecht, pp. 47-98. Machay, D.M. and Cherry, J.A. (1989) Ground water contamination pump-and-treat remediation. Environ. Sci. Technol. 23: 630-636. Rittmann, B.E. and McCarty, P.M. (1980) Utilization of dichloromethane by suspended and fixedfilm bacteria. Appl. Environ. Microbiol. 39(6): 1225-1226. Rudiger, P. (1999) Abluftreinigung in Biofilmreaktoren mit inerten Tragern. Diss. ETH No: 13229, ETH Zurich. Vinage, I. and Rudolf Von Rohr, P. (2003) Biological waste treatment with a modified rotating biological contactor. I. Control of biofilm growth and long-term performance. Bioprocess Biosyst. Engin. 26: 49-57. Zhu, L., Abumaizar, R.J. and Kocher, W.M. (1998) Biofiltration of benzene contaminated air streams using compost activated carbon filter media. Environ. Prog. 17(3): 168-172.

477

Behaviour and optimization of a novel monolith bioreactor for waste gas treatment YAOMIN JIN, MARÍA C. VEIGA AND CHRISTIAN KENNES Chemical Engineering Laboratory, Faculty of Sciences, University of La Coruña, Rúa Alejandro de la Sota, 1, 15008 – La Coruña, Spain

ABSTRACT Treatment of waste gases in bioreactors is cost-effective and environmental-friendly compared to the conventional techniques used for treating large flow rates of gas streams with low concentrations of pollutants. Nowadays, significant research is dedicated at the development of new bioreactor configurations, improved biocatalysts or new packing materials, among others. In the present study, a novel bioreactor packed with ceramic monolith was developed for treating VOCs (toluene or methanol) polluted air. Operational parameters that were considered included start-up of the bioreactor, inlet loading, changes in gas flow rate, liquid feed mode, and monolith blockage and biomass growth. Preliminary data on performance and stability have been obtained showing that this system can efficiently be used for waste gas treatment.

1 INTRODUCTION Biological treatment is an established technology for air pollution control and the alternative of choice to physical and chemical treatment techniques because of its environmental friendly and cost-effective for treating waste gases characterized by high gas flow rates and low pollutant concentrations (Kennes and Thalasso, 1998; Kennes and Veiga, 2001). The most widely utilized bioreactors for air pollution control are biofilters and biotrickling filters. Biofilters are reactors in which a humid polluted air stream is passed through a porous packed bed on which a mixed culture of pollutantdegrading organisms is naturally immobilized. In biotrickling filters, a distinct free water phase containing various nutrients is trickled over a packed bed. Both biofilters and biotrickling filters have some limitations concerning performance although they

477

478

YAOMIN JIN, MARÍA C. VEIGA AND CHRISTIAN KENNES

are currently largely used for efficiently treating air polluted with volatile organic compounds (VOCs), odorous compounds, and other air pollutants. Conventional biofilters are usually packed with natural carriers, such as compost, peat or soil. They decay over time, causing compaction, clogging, short circuiting and increased headloss across the bed. In addition, using biofiltration to control hydrophobic compounds is difficult because of mass transfer rate limitations. For biotrickling filters, packed with inert carrier materials (Kennes and Veiga, 2002), the circulating trickling liquid allows controlling the pH value and supplies the fixed biofilm with the essential inorganic nutrients. However, the trickling phase and the presence of the liquid film slow down the transfer of pollutants and oxygen from the gas to the liquid phase. Over the recent past decades, great efforts have been dedicated to the development of new bioreactor configurations in order to improve the mass transfer. Poppe and Schippert (1992) demonstrated the advantages of adding water-immiscible organic solvents to the liquid phase of bioscrubbers for the elimination of hydrophobic VOCs. By adding organic solvents with high boiling points in a range of 10-30% of the total volume, 100 to 1000 times larger amounts of hydrophobic target compounds were absorbed in the scrubber solution. In the bioreactor, the target compounds were transferred from the organic phase to the water phase driven by a concentration gradient between oil and water as biological degradation of the compounds occurred in the water phase. This new technique was demonstrated by treatment of a mixture of 13 volatile compounds in air by a two-stage scrubber. The optimization of such twoliquid-phase systems is nowadays widely being studied (Daugulis, 2001). Reij et al. (1997) used a microporous hydrophobic membrane as a support for biofilms that remove the poorly soluble propene from air. In the membrane bioreactor, the pollutant in the gas phase is transferred through a membrane to the biofilm, attached to the other side of the membrane, where nutrients and oxygen are provided (Kennes and Veiga, 2001). Vinage and von Rohr (2003) developed the rotating biological contactor for waste gas treatment. The polluted air passes through the headspace of the reactor, containing discs mounted on a rotating shaft that serve as support for a biofilm. The shaft in rotated (~ 2 rpm), and the discs are partially wetted in water containing nutrients and other additives (Kennes and Veiga, 2001). The movement of the discs favors mass transfer and the control of the fixed biomass. Kan and Deshusses (2003) developed a new vapor phase bioreactor named the foamed emulsion bioreactor (FEBR) that overcomes some of the limitations of biofilters and biotrickling filters. The FEBR consists of an emulsion of highly active pollutant-degrading microorganisms and a water-immiscible organic phase, which is made into a foam with the air being treated. The monolith, which is widely used as catalyst support for gas treatment, e.g., cleaning of automotive exhaust gases and industrial off gases, can be tailored to meet the needs of a relatively inexpensive, light weight, inert, bioreactor packing that

BEHAVIOUR AND OPTIMIZATION OF A NOVEL MONOLITH BIOREACTOR FOR WASTE

479

provides a high specific surface area (surface-to-volume) to greatly increase the mass transfer rate. Typical monoliths consist of many parallel channels separated by thin, porous ceramic walls, representing a collection of parallel microreactors. They are formed in several configurations, usually from cordierite (2MgO·2Al2O3·5SiO2) or similar silica-alumina compounds. The geometry of monolithic supports yields one major advantage over particulate packing materials since they offer very little resistance to flow. Reactors using monolithic catalyst supports may be an attractive alternative to conventional multi-phase reactors and have been used in bioconversion and fermentation processes. Nevertheless, hardly any study has been done on their application in environmental technology. In monolithic channels bubble-train or Taylor flow usually occurs. Gas bubbles and liquid slugs move with constant velocity through the monolith channels approaching plug flow behavior. The gas is separated from the catalyst by a very thin liquid film and during their travel through the channels the liquid slugs show internal recirculation. These two properties result in optimal mass transfer. The present work is related to the study and development of a novel monolith bioreactor for the treatment of waste gases containing volatile organic compounds, i.e. toluene and methanol. The feasibility of using monolith bioreactors for treating VOC-polluted air has been proved. The effect of operating conditions as the gas flow rate, liquid flow rate, and the inlet loading rate of the system have been studied.

2 MATERIALS AND METHODS 2.1 MICROBIAL CULTURES AND CULTURE MEDIUM Studies undertaken in our group on toluene-treating monolith bioreactors were performed with a microbial consortium from a conventional biofilter treating the same pollutant. Batch experiments and growth of the inocula were undertaken with an aqueous culture medium containing (per liter): 4.5 g KH2PO4, 0.5 g K2HPO4, 2.0 g NH4Cl and 0.1 g MgSO4·7H2O (Jin et al., 2005). The culture medium was autoclaved at 120 oC for 20 min before adding filter-sterilized solutions of vitamins and trace minerals. Experiments on the removal of methanol were conducted with a pure culture of Candida boidinii. The nutrient solution used both for batch assays and for bioreactor studies contained the following macronutrients (per liter): 2 g KH2PO4, 2 g K2HPO4, 0.4 g NH4Cl and 0.2 g MgCl2·6H2O, and 0.01 g FeSO4·7H2O. The culture medium was sterilized at 120 oC for 20 minutes.

480


2.2 MONOLITH SUPPORT The ceramic monolith packing has the following characteristics: geometry: square ducts; length: 150 mm; cross section: 100×100 mm; number of channels: 26×26; channel width: 3.0 mm; weight: 850 kg.m-3; geometric surface: 800 m2.m-3; voids fraction: 64%. Details of the composition and the preparation procedure of the monolith used in this work are proprietary information of Rauschert Verfahrenstechnik GmbH (Germany). 2.3 EXPERIMENTAL SETUP The schematic of the monolith bioreactor used in this study is shown in Figure 1 and has been described previously in detail (Jin et al., 2006). The reactors were usually maintained at room temperature. The polluted gas was fed to the bioreactors by mixing a large air stream flowing through a humidification chamber with a smaller air stream passing through a flask containing the pollutant, i.e. either toluene or methanol. The bioreactors were fed in a downflow mode.

Figure 1. Schematic of the laboratory scale monolith bioreactor.

2.4 ANALYTICAL METHODS Methanol or toluene concentrations were measured by means of a HP-6890 gas chromatograph (Agilent Technologies, Spain) equipped with a 30 m×0.53 mm HP-PLOT Q column and a flame ionization detector, operating in splitless mode. Oven temperature was 130 °C, while both the injector and detector temperature was


481

150 °C. Samples were injected using a 2.5 cm3 gas-tight Hamilton syringe. Under these conditions, the retention time of methanol and toluene were 3.5 min and 4.3 min, respectively (Prado et al., 2005). Similarly, CO2 concentrations were measured on another Hewlett-Packard 5890 series II GC equipped with a thermal conductivity detector (TCD). The CO2 concentrations were determined at an injection temperature of 90 °C, an oven temperature of 25 °C and using a TCD at 100 °C.

3 RESULTS AND DISCUSSION 3.1 STARTUP OF THE BIOREACTORS The microbial consortia used to inoculate the bioreactors were obtained from either a previous bioreactor or from batch enrichments. The nutrient solution containing the desired biomass was continuously recirculated over the packing material. Simultaneously, a visible biofilm developed on the surface of the square channels of the monolith. Afterwards, the monolith was transferred to the bioreactor and VOCspolluted air was fed continuously. The start-up period of the bioreactor treating toluene lasted around 24 days with removal efficiencies of 60-100% while slowly increasing the load from 0.395 to 29.5 g-toluene.m-3.h-1. These data suggest that the start-up phase is quite slow for that pollutant. However, a shorter start-up period of a few days was required for the bioreactor treating methanol. After this period, the inlet concentration was kept at 200 mg.m-3 with an EBRT of 30 s, reaching an elimination capacity of 30 g.m -3.h -1, while maintaining the removal efficiency above 95%. Acclimated biomass allowed to shorten the start-up phase, as also observed by others (Veiga and Kennes, 2001). 3.2 INFLUENCE OF THE GAS FLOW RATE The influence of the gas flow rate on the reactor´s performance was evaluated in the bioreactor treating toluene. In the range of gas flow rates of 18 to 110 l.h-1, the removal efficiency first remained constant, around 90%, while gradually increasing the gas flow rate. When the gas flow rate was further increased, the biofilm thickness decreased due to the shear force. The mass transfer limitation step was determined by the laminar film thickness between the gas and liquid phase. When the gas flow rate was increased, the turbulence of the gas increased, and the laminar film became thinner. Hence, the resistance decreased and mass transfer was enhanced. The results show that the highest elimination capacity was reached at the highest gas flow rate, although this led to a lower removal efficiency (Figure 2).

482


Figure 2. Effect of the gas flow rate on toluene removal.

3.3 INFLUENCE OF THE TRICKLING LIQUID PHASE The gas and liquid mixture travels through the channels of the monolith reactor. Depending on the flow rate of each phase and on the feed method, a number of different flow regimes can occur, such as dispersed bubble flow, bubble flow, aerated Taylor flow, Taylor flow, churn turbulent flow, slug flow, annular flow and mist flow. In the co-current downflow trickling operation, the gas and liquid phases travel in the same direction through the channels. In this operation mode the Taylor flow regime is preferred. In this regime the gas and liquid move through the channels as separate packages, ensuring plug flow behaviour. The gas bubbles are separated from the bio-catalytic wall, containing the attached biofilm, only by a thin liquid film. Gas adsorbed in this film can immediately be consumed by the bio-catalyst attached on the walls of the channels. Adsorbed gas that is not consumed at the film exchanges with the liquid plug. The recirculation pattern in the liquid plug facilitates a rapid exchange with the film. Because of these properties of Taylor flow in capillaries, a high gas-liquid mass transfer rate is obtained. In order to check the effect of the mode of feeding of the liquid phase, a toluenefed bioreactor was first operated in a trickling mode and later without trickling phase (no recirculation of the liquid medium) during the treatment of toluene. The flow rate of the gas and liquid were 80 and 1 l.h-1, respectively. The removal of toluene in the trickling mode was lower than without trickling phase in the experimental range of inlet concentrations used in this work (Figure 3). This seems to contradict the theory that Taylor flow could enhance the mass transfer from the gas phase to the liquid


483

phase, which may be due to the following reasons: (1) when no liquid is recirculated in the reactor, the gas flow is uniform, and the liquid inside the channels forms a very thin film, then the resistance between the gas phase and the liquid phase is low, (2) in the trickling mode, although the Taylor flow generated by the liquid flow could enhance the mass transfer, the liquid was not uniformly distributed in the monolith. This could cause non-homogenous mass transfer in the different channels.

Figure 3. Performance of the monolith bioreactor with or without trickling phase.

3.4 MONOLITH BLOCKAGE AND BIOMASS GROWTH The clogging of the monolith channels was first observed during the treatment of toluene. The pressure drop sharply increased from initially zero to 0.5 cm H2O. The biofilm growth made the gas flow and liquid flow regime become nonhomogenous, and the performance of the biofilter decreased dramatically. In order to remove excess biomass, a high flow rate of trickling liquid was used. The turbulence that was generated allowed to efficiently wash out part of the biofilm from the reactor. It seems that controlling the biofilm growth is a very important parameter for long term stable operation of monolith bioreactors, as also recently observed by others (Ebrahimi et al., 2006). This problem can be solved by optimizing the dimensions of the channels or by means of a high flow rate of the trickling liquid. Previous studies undertaken with smaller channels (channel width 1.27 mm instead of 3.0 mm) resulted in a still faster clogging. Studies on biomass accumulation were also performed with methanol. A low biomass growth rate acidophilic yeast, Candida boidinii, was inoculated for treatment

484


of methanol. The pressure drop across the bioreactor indicates that biomass accumulation was relatively insignificant until day 50 of operation. The pressure drop remained around 6 mm H2O/m after 35 days operation. On day 60 of this operational period, the inlet loading of methanol was increased from 75 to 150 g.m-3.h-1, this high loading of methanol enhanced excess biomass growth causing clogging of the channels. Finally, the accumulated biomass led to a dramatic increase in pressure drop across the bioreactor on day 75. Biomass accumulation has also been observed on the top view of the monolith packing. In order to remove excess biomass, a high liquid flow rate was used (3 l.h-1) to generate shear forces and remove some biomass attached on the channel walls allowing the pressure drop to return to its original value, around 6 mm H2O.m-1. It is also important to note that the biomass accumulation, as reflected in Figure 4, had very little effect on methanol removal even at high values of the pressure drop. Physical operational problems are, however, encountered at such high pressures, necessitating backwashing to remove excess biomass. Overall, the monolith bioreactor showed a higher elimination capacity and much lower pressure drops compared to other conventional bioreactors, which could save on operation costs when the bioreactor is scaled up for application in the field.

Figure 4. Development of pressure difference across the monolith bioreactor.


485

4 CONCLUSIONS According to the data available so far, it appears that monolith bioreactors are able to reach relatively high removal rates and very good performances compared to conventional systems. However, more research is still needed in order to confirm if such good results can be generalized to all monolith bioreactors and if they can be maintained over longer operation periods.

5 ACKNOWLEDGEMENTS The present research was funded by the Spanish Ministry of Education and Science (Project CTM2007-62700/TECNO) and through European FEDER funds. Yaomin Jin was financially supported by the Xunta de Galicia through Project PGIDIT05PCIC10304PN.

REFERENCES Daugulis, A.J. (2001) Two-phase partitioning bioreactors: a new technology platform for destroying xenobiotics. Trends Biotechnol. 19: 457-462. Ebrahimi, S., Kleerebezem, R., Kreutzer, M.T., Kapteijn, F., Moulijn, J.A., Heijnen, J.J. and van Loosdrecht, M.C.M. (2006) Potential application of monolith packed columns as bioreactors, control of biofilm formation. Biotechnol. Bioeng. 93: 238-245. Jin, Y., Veiga, M.C. and Kennes, C. (2005) Autotrophic deodorization of hydrogen sulfide in a biotrickling filter. J. Chem. Technol. Biotechnol. 80(9): 998-1004. Jin, Y., Veiga, M.C. and Kennes, C. (2006) Development of a novel monolith-bioreactor for the treatment of VOC-polluted air. Environ. Technol. 27(11): 1271-1277. Kan, E. and Deshusses, M.A. (2003) Development of a foamed emulsion bioreactor for air pollution control. Biotechnol. Bioeng. 84: 240-244. Kennes, C. and Thalasso, F. (1998) Waste gas biotreatment technology. J. Chem. Technol. Biotechnol. 72(4): 303-319. Kennes, C. and Veiga, M.C. (2001) Bioreactors for Waste Gas Treatment. Kluwer Academic Publishers, Dordrecht, The Netherlands. Kennes, C. and Veiga, M.C. (2002) Inert filter media for the biofiltration of waste gases – characteristics and biomass control. Rev. Environ. Sci. Bio/Technol. 1: 201-214. Poppe, W. and Schippert, E. Das KCH-biosolv-ver-fahren in kombination mit einembiowäscher herkömlicher art-eine verfahrensentwicklung zur abluftreinigung fur wasserlösliche und schwer wasserlösliche schadstoffe. In Biotechniques for Air Pollution Abatements and Odour

486


Control Policies (Maastricht, The Netherlands, 1992), A. J. Dragt and J. van Ham, Eds., Elsevier Science Publishers B.V., pp. 71-76. Prado, O.J., Veiga, M.C. and Kennes, C. (2005) Treatment of gas-phase methanol in conventional biofilters packed with lava rock. Water Res. 39(11): 2385-2393. Reij, M.W., Hamann, E.K. and Hartmans, S. (1997) Biofiltration of air containing low concentrations of propene using a membrane bioreactor. Biotechnol. Prog., 13(4): 380-386. Veiga, M.C. and Kennes, C. (2001) Parameters affecting performance and modeling of biofilters treating alkylbenzene-polluted air. Appl. Microbiol. Biotechnol. 55: 254-258. Vinage, I. and von Rohr, R. (2003) Biological waste gas treatment with a modified rotating biological contactor. I. Control of biofilm growth and long-term performance. Bioprocess Biosyst. Eng. 26(1): 69-74.

487

Biodegradation of BTXS and substrate interactions in a Bioactive Foam Reactor JIHYEON SONG AND SHONG-GYU SHIN Dept. of Civil & Environmental Engineering, Sejong University, Kunja-Dong, Kwangjin-Gu, Seoul, Korea 143-747

ABSTRACT A bioactive foam reactor (BFR), using surfactant-driven bubbles and suspended microorganisms, has emerged as a potential alternative to packed-bed biofiltration systems for the treatment of volatile organic compounds (VOCs). The study presented herein was designed to investigate the effects of VOC mixtures (benzene, toluene, p-xylene, and styrene) on biodegradation efficiencies and substrate interactions in the BFR. Benzene, toluene p-xylene, and styrene were applied individually to the toluene-acclimated BFR at the same inlet concentration (0.78 g/m3), and then paired BTXS mixtures (BT, BX, BS, TX, TS, XS, and BTXS in the same ratio by volume) were applied but the total inlet concentration were maintained constant. The overall removal rates of each of the four VOCs were in the following order: toluene, styrene, benzene, and p-xylene in the inlet concentration range tested. However, styrene biodegradation was the highest in the presence of other VOC compounds. The removal efficiency for toluene as a single substrate was 82%, but toluene removal efficiencies dropped when the paired mixtures were applied. The removal efficiency for benzene also decreased in the presence of other TXS compounds. In contrast, the removal efficiency for p-xylene as a single substrate was only 21% in the BFR, but p-xylene removal efficiencies ranged 35 – 41% in the presence of other BTS compounds. As a result, the biodegradation of benzene and toluene was inhibited by the other carbon sources, whereas the biodegradation of styrene and p-xylene was enhanced by the others. Consequently, a careful attention needs to be given when BFR performance and biodegradation rates of mixed VOCs are utilized for system design and operational purposes.

487

488

JIHYEON SONG

AND

SHONG-GYU SHIN

1 INTRODUCTION Packed-bed bioreactors including biofilters and biotrickling filters have drawn increased interests for the treatment of volatile organic compounds (VOCs). Many studies have shown that these bioreactors can successfully treat a wide range of VOCs including benzene, toluene and xylenes (van Groenestijn and Hesselink, 1993; Devinny et al., 1999). However, operational problems such as excess biomass accumulation and biodegradation activity loss make these treatment methods less attractive (Song and Kinney, 2000), especially when subjected to high concentrations of VOC mixtures. In order to overcome those problems, several new approaches have been made using suspended microorganisms instead of the fixed biofilm (Kennes and Veiga, 2001). One of them is a bioactive foam reactor (BFR) that is operated with a surfactant bubble solution containing pollutant-degrading microorganisms (Phipps, 1998). In the BFR, a VOC-laden air stream is sparged into the surfactant solution and forms biologically active foams. The fine foams can provide a large surface area for the mass transfer of VOCs as well as enhance microbial activity without a significant accumulation of biomass over time. Kan and Deshusses (2003, 2005) demonstrated that a foamed bioreactor using an organic-phase emulsion and active microorganisms could achieve a high elimination capacity for toluene. Recently we have modified the defoamer of the BFR system from its patented prototype to make it simple, and an organic emulsion such as oleyl alcohol was not used in the liquid phase. Since BFR performance mainly relies on the mass transfer of VOCs and the subsequent microbial degradation, substrate interactions between the VOCs present in inlet streams are another important factor that must be considered to achieve successful BFR operation when it is subjected to different chemical mixtures. Emissions from various industrial sources often consist of a mixture of compounds with different chemical characteristics and biodegradability. Therefore, several bioreactor studies have examined the biodegradation of binary mixtures of VOCs, such as BTEX (Collins and Daugulis, 1999; Strauss et al., 2004), and methanol and α-pinene (Mohseni and Allen, 2000). These studies have consistently indicated that the presence of a readily degradable compound can inhibit the removal of the other recalcitrant compound. However, few attempts have been made to investigate the effects of substrate inhibition between complex VOC mixtures in bioreactors using suspended microbial cultures such as the BFR. The study presented herein was therefore designed to investigate the effects of a VOC mixture (benzene, toluene, p-xylene, and styrene) on biodegradation efficiencies and substrate interactions in the BFR.

BIODEGRADATION OF BTXS AND SUBSTRATE INTERACTIONS

489

2 MATERIALS AND METHODS 2.1 MICROBIAL CULTURE AND SURFACTANT In order to obtain the toluene-degrading culture used in this study, a mixed microbial culture was initially collected from a wastewater treatment plant in Seoul, Korea. Pseudomonas putida TDB4 was isolated from the mixed culture and cultivated in our laboratory using gaseous toluene as a sole carbon and energy source. The mineral medium used for cultivating the mixed culture and P. putida TDB4 was slightly modified from the composition described by Song and Kinney (2000), which contained 1.36 g/L KH2PO4, 1.42 g/L Na2HPO4, 3.03 g/L KNO3 and trace metals per liter of distilled water. The surfactant used in the BFR was TritonX-100 (Sigma-Aldrich, USA), which was selected based on a bottle test that showed no adverse effects on toluene degradation of P. putida TDB4. The surfactant concentration in the liquid phase was maintained at 0.013% (v/v) that was slightly lower than its critical micelle concentration (0.014%).

Figure 1. Schematic of the bioactive foam reactor (BFR) tested in this study.

2.2 BFR CONFIGURATION As shown in Figure 1, the lab-scale bioreactor used in this study consisted of a foam column (volume 1.8 L), where the mass transfer of VOCs between the gas and foams took place, and a cell reservoir (volume 2.5 L), where biologically active microorganisms degraded the VOCs in the liquid phase. An air stream, generated by a

490

JIHYEON SONG

AND

SHONG-GYU SHIN

compressor and controlled by a flow meter, was contaminated with VOC vapour by the slow injection of research-grade, pure compounds using a syringe pump (KD Scientific, USA). The VOC-contaminated air stream was introduced to the bottom of the foam column through an aeration stone. As the air stream was sparged into the nutrient solution containing the surfactant and the microbial culture, fine foams were generated and moved along with the air stream in the foam column. The foam itself broke down at the top of the cell reservoir, returned to the liquid phase, and then the air trapped in the moving foam was released and exited the reactor. The liquid phase was continuously recirculated from the cell reservoir to the foam column in a closedloop, and the total liquid volume in the entire system was 1.8 L. 2.3 BFR OPERATION AND SUBSTRATE INTERACTIONS Two sets of BFR experiments were conducted to determine the effect of various combinations and concentrations of the VOCs (BTXS) on removal efficiencies at a constant gas retention time of 40 seconds. Prior to each BFR start-up, a microbial solution consisted of 1 L of pre-grown P. putida TDB4 was mixed with 1 L of the nutrient medium containing the surfactant. And then, the BFR was acclimated to toluene biodegradation at an inlet toluene concentration of 0.38 g/m3 (i.e., 100 ppmv) until the steady state condition was obtained. First, in order to determine short-term responses of the bioreactor system at the inlet concentration of 200 ppmv, benzene, toluene, p-xylene, and styrene were applied individually to the toluene-acclimated BFR for five hours. And the short-term experiments were repeated for various VOC mixtures (BT, BX, BS, TX, TS, XS, and BTXS) at the same total concentration of 200 ppm v. Between each short-term experiment, the acclimation condition was restored to maintain the BFR at steady state. The second set of BFR experiments was conducted to observe the performance of the BFR operated continuously using BTXS that were mixed in a 1:1:1:1 ratio by volume. Throughout the 13-day operational period, the inlet VOC concentrations were changed twice. Initially the VOC mixture at the total concentration of 100 ppmv was introduced to the toluene-acclimated BFR (referred to as «Phase M1»). On day 6, the inlet concentrations of the VOC mixture were increased stepwise to 200 ppmv («Phase M2»). As a measure of substrate interaction between the compounds in the mixtures, a substrate interaction index (SII) was defined as a ratio of the changes in removal efficiency due to the other coexisting carbon source to the removal efficiency determined when a single carbon source was supplied. The SIIs were calculated for each compound in the presence of other compounds at the different concentrations tested in this study.


491

Substrate Interaction Index (SII) = (VOCmix – VOCsingle)/VOCsingle

(1)

Where VOCmix is the removal efficiency of the target compound in mixture, and VOCsingle is theremoval efficiency of the compound as a single substrate at the given concentration. 2.4 ANALYTICAL METHODS To determine VOC removals, gas samples were collected from three ports located at the inlet, at the middle of the connecting line between the foam and the microbial columns, and at the outlet of the BFR. The samples were collected with 0.5mL gas-tight syringes and immediately analyzed using a gas chromatograph (HP 6890, Agilent, USA) equipped with a flame ionization detector.

3 RESULTS AND DISCUSSION 3.1 SUBSTRATE INTERACTIONS IN THE SHORT-TERM RESPONSES The short-term changes in VOC removal efficiencies were determined when various combinations of different VOCs were applied to the toluene-acclimated BFR. The biodegradation of toluene in the absence of other compounds (BXS) by the toluenedegrading pure culture (P. putida TDB4) showed a high and stable removal efficiency. The removal efficiencies dropped when BXS compounds were applied individually to the BFR during the short-term period. The overall removal efficiencies of the four VOCs were in the following order: toluene, styrene, benzene, and p-xylene. Styrene is the most soluble compound among the VOCs tested, but its removal efficiency was lower than that of toluene in the toluene-acclimated microbial system. In addition, the biodegradation efficiency of pxylene was the lowest indicating that p-xylene was the most recalcitrant compound, and this finding was similar to other results reported in the biofiltration literature (Collins and Daugulis, 1999; Deshusses and Johnson, 2000; Strauss et al., 2004). The removal efficiencies of a target compound in the presence of the other compounds were monitored in the toluene-acclimated BFR. Figure 2 illustrated the SIIs calculated using the experimental data obtained in the short-term BFR operation. For instance, the removal efficiency of toluene as a single substrate was 82% at the inlet concentration of 200 ppmv, and it dropped to 55% when both toluene (100 ppmv) and styrene (100 ppmv) was applied. Therefore, the SII for toluene in the presence of styrene («Ts» in Figure 2) was -0.32, indicating that the toluene biodegradation was inhibited by the styrene addition. In comparison, the styrene removal efficiency was found to be 77% when styrene was individually supplied to the BFR, but it increased to 91% when toluene and styrene were applied together in the paired mixture, yielding

492

JIHYEON SONG

AND

SHONG-GYU SHIN

the SII of 0.18 for styrene in the presence of toluene («St» in Figure 2). These findings indicate that the styrene biodegradation was enhanced by the presence of toluene. Overall, as shown in Figure 2, the SIIs for both benzene and toluene were negative values, implying that the biodegradation of these compounds was always inhibited by the presence of the other compounds. Strauss et al. (2004) reported the results obtained from a packed-bed biofilter system that the toluene biodegradation was inhibited by the other compounds (benzene, p-xylene and ethylbenzene), and toluene had an enhancing effect on the removal efficiency of the other compounds when paired. However, in this study, the presence of toluene and benzene in the paired mixtures resulted in the mutual deterioration of the removal efficiencies of both compounds. The biodegradation of both p-xylene and styrene in the mixtures was enhanced by the presence of other compounds. The improved removal efficiencies of p-xylene and styrene occurred at the expense of toluene removal efficiency, presumably due to the similarities of the enzymatic systems used in the metabolic pathways for the aromatic compounds. Therefore, p-xylene, the most recalcitrant compound, had a greatest enhancing effect by the presence of other structurally-related aromatics. In addition, the presence of p-xylene and styrene in the paired mixture resulted in the mutual enhancement of the removal efficiencies of both compounds.

Figure 2. Substrate interaction indices for target compounds in the presence of the other compounds in paired mixtures.


493

3.2 SUBSTRATE INTERACTIONS IN THE CONTINUOUS BFR OPERATION The response of the BFR operated continuously using BTXS in the paired mixture of a 1:1:1:1 ratio by volume was monitored throughout the 13-day period. Figure 3 illustrates that the inlet concentration of each compound and its outlet concentration during the operational period. In the continuous experiment when subjected to the mixture of the four compounds, the removal efficiency was ranked in the following order: styrene, toluene, benzene, and p-xylene. The ranking of removal efficiencies for styrene differed from that of the short-term experiment. This finding implies that the toluene-acclimated microbial culture could become adapted to the biodegradation of styrene as the BFR operation continued, since styrene is a more soluble and readily biodegradable compound than the others. An averaged removal efficiency of each compound at pseudo-steady-state during Phase M2 was used to calculate SII values. The SII values were -0.01, -0.10, 2.02, 0.10 for Btxs, Tbxs, Xbts, Sbtx, respectively. Similar to the short-term BFR experiment, the SII values showed that the biodegradation of benzene and toluene was inhibited, but the biodegradation of p-xylene and styrene was enhanced by the presence of other compounds in the continuous BFR experiment.

Figure 3. Changes of inlet and outlet concentrations of (a) benzene, (b) toluene, (c) p-xylene, and (d) styrene in the VOC mixture during the continuous BFR operation.

494

JIHYEON SONG

AND

SHONG-GYU SHIN

It is interesting to note that all the SII values obtained in the continuous operation shifted to the positive direction from its value obtained from the previous short-term experiment (e.g., -0.20 -0.01 for Btxs). The enhancement of substrate interaction in the paired mixture was presumably due to an increase in microbial density in the cell reservoir as well as gradual adaptation of the microbial strain to the biodegradation of the BTXS mixture over time. Consequently, a careful attention needs to be given when BFR performance and biodegradation rates of mixed VOCs are utilized for system design and operational purposes.

4 ACKNOWLEDGEMENTS This work was supported by grant No. R01-2005-000-10675-0 from the Basic Research Program of the Korea Science and Engineering Foundation.

REFERENCES Collins, L.D. and Daugulis, A.J. (1999) Benzene/toluene/p-xylene degradation: solvent selection and toluene degradation in a two-phase partitioning bioreactor. Appl. Microbiol. Biotechnol. 52: 359-365. Deshusses, M.A. and Johnson, C.T. (2000) Development and validation of a simple protocol to rapidly determine the performance of biofilters for VOC treatment. Environ. Sci. Technol. 34: 461-467. Devinny, J.S., Deshusses, M.A. and Webster, T.S. (1999) Biofiltration for air pollution control. Boca Raton: Lewis Publishers. Kan, E. and Deshusses, M.A. (2003) Development of foamed emulsion bioreactor for air pollution control. Biotechnol. Bioeng. 84: 204-244. Kan, E. and Deshusses, M.A. (2005) Continuous operation of foamed emulsion bioreactors treating toluene vapors. Biotechnol. Bioeng. 92: 364-371. Kennes, C. and Veiga, M.C. (2001) Bioreactors for waste gas treatment. Kluwer Academic Publishers. Mohseni, M. and Allen, D.G. (2000) Biofiltration of mixtures of hydrophilic and hydrophobic volatile organic compounds. Chem. Eng. Sci. 55: 1545-1558. Phipps, D.W. (1998) Biodegradation of volatile organic contaminants from air using biologically activated foam. US Patent No.5,714,379. Song, J. and Kinney, K.A. (2000) Effect of Vapor-Phase Bioreactor operation on biomass accumulation, distribution, and activity. Biotechnol. Bioeng. 68: 508-516. Strauss, J.M., Riedel, K.J. and du Plessis, C.A. (2004) Mesophillic and thermophilic BTEX substrate interactions for a toluene-acclimated biofilter. Appl. Microbiol. Biotechnol. 64: 855-861. van Groenestijn J.W. and Hesselink P.G.M. (1993) Biotechniques for air pollution control. Biodegradation. 4: 283-301.

495

Characterization and performance evaluation of a two-phase partitioning bioreactor for volatiles organic compounds treatment in off-gas JEAN-MARC ALDRIC AND PHILIPPE THONART Centre Wallon de Biologie Industrielle, Unité de Bio-industries, Faculté des Sciences Agronomiques de Gembloux, Passage des déportés, 2, B – 5030 Gembloux, Belgium

ABSTRACT The treatment of the industrials off-gas strongly evolves with change of the environmental legislation on a worldwide scale. Biotechnics existing for their treatment sometimes present limits for some volatile organic compounds (VOC) such as BTEX because of their poor water solubility. The use of two phase-partitioning bioreactors (TPPB) is an interesting alternative in this case. In this work, a laboratory scale TPPB (water / silicone oil) was monitored at high level of Isopropropylbenzene (IPB) air pollution (7g/Nm3) and a flow of 1 VVM. We focused ourselves on the inoculation with the strain Rhodococcus erythropolis T 902.1. We showed that the increase of the inoculums size to 5 g DM/l. induces a better initial abatement of the pollutant, however performances of the TPPB decrease quickly because of cellular mortality. The use of a smaller inoculum (0,2g DM /l) seems to be a good compromise to observe progressive improvement of the IPB abatement with adaptation of biomass. The TPPB was followed during 38 days in order to confirm its potentialities and characterise its evolution. We showed that the performances of the TPPB are maintained with an elimination rate near to 63 % for IPB polluted air (7 g/Nm3) and punctually reach 92 %. The biomass grows gradually and stabilizes itself around 10 g/l. With fluorescent double stain Rhodamine/ propidium iode, we also shown that cellular viability strongly evolve: cellular viability was low (30 %) in the first operating hours but is quickly increased after adaptation to IPB (80%); we also suggested an endorespiration phenomenon in the bioreactor. In this work, we could confirm the previously estimated elimination performances of the two- phase partitioning bioreactor with silicone-oil as second phase. Elimination rate of a monoaromatic compound at high concentration (7 g / Nm3) can be maintained between 240 g /m3.h and 360 g /m3.h. in the TPPB.

495

496

JEAN-MARC ALDRIC

AND

PHILIPPE THONART

1 INTRODUCTION Numerous polluting organic compounds are released by human activities and persist in the environment, because of their low solubility in water and their high concentration. Monoaromatic hydrocarbons such as benzene are produced in large amounts, and used in fuels, as solvents and as starting materials for the production of plastics, synthetic fibres and pesticides (Budavari et al., 1996). Monoaromatics have become prevalent environmental contaminants, and thirty of them are on the «EPA Priority Pollutant List» (1996). Eleven of these compounds are in the top 100 chemicals on the Priority «List of Hazardous Substances» published by the Agency for Toxic Substances and Disease Registry (ASTDR, 1997). In recent years, biological techniques have been applied more frequently to control these emissions, because they eliminate many of the drawbacks of classical physical-chemical techniques. Disadvantages of usual air treatment techniques are high-energy costs (incinerators), the use of expensive chemicals that may require special operational safety procedures (chemical scrubbers) and the generation of waste products such as spent chemical solutions or spent activated-carbon (Van Groenestijn et al., 2005; Davidson et al., 2003). Biological methods involving biofilter has been shown to be promising alternatives compared to the traditional technologies for the control of many gaseous pollutants (Rene et al., 2005). Bio filters present however several limits such as the ripening period of the bio filters during which cells proliferate to the point where the bed can be used and their restriction to the treatment of low VOC (volatile organic compounds) concentration (below 1g / Nm3). This is partly due to the poor water-solubility of gaseous pollutants. Some researches were carried out in order to improve transfer and solubility of hydrophobic pollutants during biological treatment. Budwill and Coleman (1997) showed the positive effect of silicone-oil addition on the biodegradation rate of n-hexane vapours in peat-based bio filters. Cesàrio et al. (1997) showed an enhancement of the toluene mass transfer rate by a factor of 1,1 using a dispersion containing 10% (v/v) FC40 solvent and a twofold oxygen transfer rate. More recently, many researches were devoted to the Two-Phase Partitioning Bioreactors (TPPB) as a new technology for xenobiotic degradation, The TPPB. concept has been demonstrated to be effective for the degradation oh high levels of organic compounds (Daugulis et al., 2001). Dumont and Delmas (2003) reviewed the mass transfer enhancement of gas absorption in oil-in-water systems and conclude that our understanding, of the influence of oil addition on the mass transfer parameters kL and a could be improved. By improving the oxygen and pollutant mass transfer and reduce the inhibitory substances by lowering their concentration in the aqueous phase, the elimination capacity reported for two-phase partitioning bioreactor often exceeded the performance

CHARACTERIZATION AND PERFORMANCE EVALUATION OF A TWO-PHASE

497

of bio filters. For nonexhaustive example, the elimination capacity quoted in some studies are summarized in Table 1. Table 1. Some examples of elimination capacity quoted in literature for some compounds Compounds

Microorganism

Hexane

Pseudomonas aeruginosa Sphingomonas aromaticivorans Alcaligenes xylosoxidans Y234 Alcaligenes xylosoxidans

HAP (naphtalène et phénanthrène) Benzene Toluene

Elimination capacity (EC) g/m³réact.h

Reference

135

Muñoz et al. (2006)

238

Daugulis and Janikowski (2002) Yeom and Daugulis (2001) Daugulis and Boudreau (2002)

291 727 727

To compare, a recent study (Arriaga et al., 2006) reported in a fungal biofilter supplied with silicone-oil quote EC of 100 g hexane /m3.h, however higher than those reported in classical fungal and bacterial biofilters (60g/m3react.h). The open nature of biofilters limits the control of parameters, inoculation with a single chosen species may thus fail (Devinny et al., 1999). In a TPPB, the best control of parameter permits the inoculation with an adaptated strain. In a similar way, the acclimatation in compost biofilters treating gasoline vapors, was much more rapid when they were inoculated with adaptated culture but do not affect ultimater removal efficiencies (Wright et al., 1997; Leson and Smith, 1997). In this study, we review the merit and limitation of a water / silicone oil TPPB used for treating isopropropylbenzene (IPB) gaseous vapours at high concentration. We also consider the influence of various quantities of inoculum on the initial evolution of EC.

2 MATERIALS AND METHODS 2.1 MICROORGANISMS AND CULTIVATION The Rhodococcus erythropolis strain was obtained from the collection of the Walloon Centre of Industrial Biology (C.W.B.I.; Belgium). All the substrates and other chemicals were purchased at VWR international (Leuven, Belgium) or Aldrich (Bornem, Belgium).

498

JEAN-MARC ALDRIC

AND

PHILIPPE THONART

The culture of Rhodococcus erythropolis in 868 medium (glucose 20g/l.; casein peptone 20 g/l; yeast extract 10 g/l) is harvested after 64 hours (optical density 600 nm =1,4). The inoculum for the biological reactor is obtained by centrifugation of various volume of this culture in function of experiment. The pellet obtained is washed twice and diluted in 200 ml saline water (9g/l NaCl). The inoculum is then introduced into the bioreactor where the medium for biodegradation is composed of silicone oil (10% V/V) and aqueous medium M284 (90 % V/V) whose composition is : Na2HPO4 17,7g/l; NaH2PO4 24,33 g/l (buffer pH 7); NaCl 4,68g/l ; KCl 1,49g/l; NH4Cl 1,07g/ l; Na2SO4 0,43g/l; MgCl2. 6H2O 0,20g/l ; Na2HPO4 2H2O 40mg/l ; CaCl2. 2H2O 30mg/ l Fe(III)NH4citrate 4,8mg/l; ZnSO4.7 H2O 0,144 mg/l ; MnCl2.4 H2O 0.1 mg/l ; H3BO3 0,062 mg/l ; CoCl2.6 H2O 0,19 mg/l; CuCl2.2 H2O 0,017 mg/l ; NiCl2.6 H2O 0,024 mg/l ; Na2MoO4.2 H2O 0,036 mg/l; ethanol 1g/l. 2.2 BIOREACTOR AND ASSEMBLY The stirred bioreactor used for biodegradation (LSL Bio Lafitte BL06.1, Saint German en Laye, France) described by Aldric (2005). Its reactional volume reaches 4,5l and the stirring speed was maintained at 600 rpm. The assembly is schematized in figure 1.

Figure 1. Schematic of the laboratory scale TPPB. A flow-meter (2) allows the control of the gas flow by means of a control valve (3). A bottle (5) containing pollutant (IPB) allows the generation of the polluted effluent with a temperature control (7). A valve is used to dilute the polluted effluent and to control the concentration at level of the sampling bubble preceding the bioreactor (9). A septum (10) is used to sample gas. The stirred bioreactor is the seat of the biodegradation (12). A sparger (13) is used to allow a diffusion of little bubbles of gas effluent within the bioreactor. Stirring and mixing are carried out by means of two agitation modules : a TD4 module (14) and a helicoidal module (15). Baffles (16) avoid the formation of vortex within the bioreactor. Lastly, a sampling bubble permits to measure the residual concentration at the exit of the bioreactor.


499

The IPB concentrated gas is generated by stripping within a thermostatized glas bottle. The gas flow is permanently controlled by a flow meter and fixed at 4,5 l/min. The concentration of IPB in the gas coming in the bioreactor is controlled by an adjustable mixture between polluted gas and air and maintained at 7 g / Nm3. 2.3 SAMPLING AND ANALYTICAL METHODS Gas samples are regularly taken from each bubble of sampling as well as in the liquid reactional medium. IPB concentration was estimated thanks to a Perkin Elmer headspace sampler HS 40 XL (for liquid samples) and a gaz chromatograph Hewlett Packard 5890 equiped with a Alltech INC. Deerfield EC-WAX column and flame ionisation detector (for gas-samples). Temperatures of the injector, column and detector were respectively 153, 150 and 250 °C. 2.4 CELLULAR VIABILITY Liquid samples are regularly taken from TPPB and diluted to approximately 106 UFC/ml. 1ml was then centrifuged (8000 rpm; 10 min), the pellet was twice washed with sterile saline water (9g/l NaCl). The pellet was double stained with 5μl of Rhoda mine 123 (1,25 mM) and 5μl of propidium iode (1,25mM), homogenized and incubated 5 minutes at ambient temperature. The cellular vialbility can be evaluated by a fluorescent microscopy, viable cells appear green and nonviable cells appear red. The proportion of viable cells were estimated by enumeration of red and green cells in 5 microscopic fields of vision / sample.

3 RESULTS AND DISCUSSION 3.1 INFLUENCE OF INOCULUM DENSITY ON THE REMOVAL EFFICIENCY In this experiment, the TPPB was inoculated with several inocula in order to evaluate impact of initial cellular biomass on the performances of bioreactor during the firsts four days. Indeed, the ratio pollutant concentration / adapted micro-organisms concentration could be an important factor when the TPPB is started. According to the data shown in Figure 2, it appeared that the use of a high size of inoculum (2,4 g DM / l) allows reaching initial removal efficiency near to 95 % but a reduction in RE is subsequently observed during the first 4 days. On the other and, for a lower size (respectively between 0,2 and 1,53 g DM/ l) the RE was maintained with a low value (around 80%) during the first 4 days. Singularly, smallest size of inoculum (0,2 g DM / l), allows to reach the highest RE after 4 operating days (84 %).

500

JEAN-MARC ALDRIC

AND

PHILIPPE THONART

Figure 2. Influence of size of inoculums on the performance of a TPPB removing IPB.

Figure 3. Evolution of biomass dry matter during the first 4 days of experimentation for various size of inoculum.

Figure 3 clearly indicates that the cellular multiplication is inversely proportional to the size of the inoculum. When the inoculum size is high (beyond 1,5 g/l), no beneficial effect can be observed, neither on the cell multiplication, nor on the removal efficiency. However, under the same conditions, a TPPB inoculated with an inoculum


501

size of 0,2 g/l reach 3,75 g/l and a RE maintained at 85 % after 10 operating days (results not shown). Small inoculum (0,2 g/l) seems to be the best compromise to quickly reach high RE and an adaptation of the biomass to high IPB concentration. 3.2 EVALUATION AND EVOLUTION OF THE PERFORMANCES OF THE TPPB The performances of the TPPB were evaluated while following the elimination flow of IPB (during 38 days). The TPPB was first continuously sparged during the first four days with an inlet gaz flow (7g/Nm3 IPB; 1 VVM) (phase 1). Then, the TPPB was sparged only during the day (phase 2) because of the too much low RE during the first phase. Lastly, the TPPB was again continuously sparged between day 30 and day 38 to evaluate the performances under extreme conditions (phase 3).

Figure 4. Evolution of the flows within the bioreactor Qin (rhombus) : IPB flow coming in TPPB; Qelim (triangle): IPB elimination flow (mg/min.l). Small dotted lines and large dotted lines represent respectively the average of Qin and Qelim during phase 2.

Figure 4 shows the performances evolution for a TPPB during each phase. It should be specified that le TPPB is used in the limiting predetermined conditions (Aldric et al., 2005), however, under these conditions, the removal efficiency is supported at approximately 63 % and punctualy reach 81% and 92 % with twelfth and thirteenth days, the elimination flow during phase 2 is thus estimated at 4mg/l.min (240g/m3react .h). On the other hand, when starting the third phase, the removal efficiency fall to 25 % (elimination flow below 2mg/l.min) at day 31 but follows a readadaptation at the end of experiment.

502

JEAN-MARC ALDRIC

AND

PHILIPPE THONART

Figure 5. Evolution of IPB concentration in water-silicone oil media.

IPB concentration in the two-phase media (Figure 5) increases gradually during phase 2 and stabilizes itself between 1000 and 1500 mg/l. The absorbtion of pollutant is indeed followed of a consequent biodegrdation since the accumulation of IPB in the TPPB is limited. Nevertheless, continuous sparging with IPB gas flow (night and day; phase 3) increases strongly the IPB accumulation. This leads to a reduction of driving force term (CL0– CL) resulting from well-known equation 1, this can explain the reduction of removal efficiency observed. eq. 1

3.3. EVOLUTION OF BIOMASS Growth of Rhodococcus erythropolis T 902.1 within the TPPB and viability of cells constitute significant parameters to evaluate the influence of the IPB load on the biomass. Figure 6 shows a high growth of the biomass during the first 19 days, from 0,3 g/l to 12,75 g/l. Thereafter, the biomass is stabilized between 9 and 11 g/l as from day 20 until day 38. Figure 6 show a very low viability of the cells when the TPPB is starded, only 15h after inoculation with washed biomass. Cellular viability increases then continuously to reach 84 % after 7 operating days and to be maintained between 68 % and 78 % during phase 2. When phase 3 is started (day 30), the biomass seems to undergo a shock probably induced by the increase of IPB concentration in two-


Figure 6. Evolution of biomass dry matter (DM)

503

Figure 6. Evolution of cellular viablity

phase media, however the biomass seems to subsequently readapt with the news operating conditions. The stabilization of the biomass as well from the quantitative point of view (dry matter) as qualitative (viability) suggest a cellular metabolism more directed towards the biodegradation and the endorespiration phenomenon that towards the cell multiplication.

4 CONCLUSIONS The following conclusions can be drawn from the results presented in this study (1) The use of a small inoculums size (0,2 g DM/l) is preferable with the use of more significant sizes because of a better adaptation of the biomass to IPB load. (2) The use of a TPPB water-silicone oil to treat hight concentration IPB flows is confirmed. For an IPB load near to 390 g.m -3.h-1, an IPB average elimination flow of 240 g.m-3.h-1 can be obtained and maintained during 31 days of noncontinuous operating conditions, in addition the biomass remains functional. (3) The limits of the TPPB seem to be reached when the TPPB is continuously fed with strongs IPB flows and concentration, although the biomass seems to relatively readapt after the shock.

5 ACKNOWLEDGEMENTS The authors wish to acknowledge Dow-Corning Society and Irs. Tiangoua Kone and Didier Mundella for their participation in this work.

504

JEAN-MARC ALDRIC

AND

PHILIPPE THONART

REFERENCES Aldric, J.M., Destain, J. and Thonart, P. (2005) The two-phase bioreactor water /silicon-oil: prospects in the off-gas treatment. Appl. Biochem. Biotechnol. 121-124: 707-720. Arriaga, S., Muñoz, R., Hernandez, S., Guieysse, B. and Revah, S. (2006) Gaseous hexane biodegradation by Fusarium solani in two liquid phase packed-bed and stirred tank bioreactors. Environ. Sci. Technol. 40: 2390-2395. ASTDR, Priority List of Hazardous Substances, Agency of Toxic Substances and Disease Registry, USA, 1997. Budavari, S., O’Neil, M.J., Smith, A. and Heckelman, P.E. (1996). http://www.amazon.com/exec/ obidos/search-handle-url/index=books&field-author-exact=Merck&rank=relevance%2C%2Bavailability%2C-daterank/102-6297388-6039339. The Merck Index: An Encyclopaedia of Chemicals, Drugs and Biological, Merck, Whitehouse Station, NJ. Budwill, K. and Coleman, R.N. (1997) Effect of silicone-oil on bio filtration of n-hexane vapours. Med. Fac. Univ. Gent. 62: 1521-1528. Cesário, M .T., Beverloo, W.A., Tramper, J. and Beeftink, H.H. (1997) Enhancement of gas-liquid mass transfer rate of polar pollutants in the biological waste gas treatment by a dispersed organic solvent. Enz. Microb. Technol. 21: 578-588. Daugulis, A.J. and Boudreau, N.G. (2003) Removal and destruction of high concentration of gaseous toluene in a two-phase partitioning bioreactor by Alcaligenes xylosoxidans. Biotechnol. Lett. 25: 1421-1424. Davidson, C.T. and Daugulis, A.J. (2003) The treatment of gaseous benzene by two-phase partitioning bioreactors: a high performance alternative to the use of bio filters. Appl. Microbiol. Biotechnol. 62: 297-301. Devinny, J.S., Deshusses, M.A. and Webster, T.S. (1999) Biofiltration for air pollution control. Boca Raton: Lewis Publishers. 299 p. Dumont, E. and Delmas, H. (2003) Mass transfer enhancement of gas absorption in oil-in-water systems: a review. Chemical Engineering and Processing. 42: 419-438. EPA, Priority Pollutants, Code of Federal Regulations, (1996). Title 40, Part 423, Appendix A, USA, Chapter 1. Leson G.; Smith B.J., Petroleum environmental research forum field study on bio filters for control of volatile hydrocarbons. J. Environ. Eng. 123: 556. Rene E.R., Murthy D.V.S. and Swaminathan T. (2005) Performance evaluation of a compost biofilter treating toluene vapours. Proc. Biochem. 40: 2771-2779. Van Groenestijn, J.W. and Kraakman, N.J.R. (2005) Recent developments in biological waste gas purification in Europe. Chem. Engin. J. 113: 85-91. Wright, W.F., Schroeder, E.D., Chang, D.P.Y. and Romstad, K. (1997). Performance of a pilotscale compost infielder treating gasoline vapour. J. Environ. Eng.123: 547.

505

Removal of dichloromethane from waste gases using a fixed-bed biotrickling filter and a continuous stirred tank bioreactor LAURA BAILÓN1, YOLANDA DOPICO1, MARCELL NIKOLAUSZ2, MATTHIAS KÄSTNER2, MARÍA C. VEIGA1 AND CHRISTIAN KENNES1 1

Chemical Engineering Laboratory, Faculty of Sciences, University of La Coruña, Rúa Alejandro de la Sota, 1, 15008 – La Coruña, Spain 2 Department of Bioremediation, UFZ-Centre for Environmental Research Leipzig-Halle GmbH, Permoserstr. 15, 04318 – Leipzig, Germany

ABSTRACT A laboratory scale fixed bed biotrickling filter (BTF) and a continuous stirred tank bioreactor (CSTB) have been studied and compared for the elimination of dichloromethane from waste gases. The DCM removal efficiency in the trickling filter was > 85% for inlet loads up to 25 g.m-3.h-1 and a maximal removal capacity of about 170 g.m-3.h-1 was achieved at a load of 350 g.m-3.h-1. The continuous stirred tank bioreactor showed removal efficiencies > 90% for inlet loads up to 120 g.m-3.h-1. At this load the maximal removal capacity of the system was reached, i.e. about 100 g.m-3.h-1. Thus, higher maximum elimination capacities were reached in the BTF while higher removal efficiencies were obtained at high loads with the CSTB. Both systems presented good stability against overloads.

1 INTRODUCTION Dichloromethane (DCM) is produced in large amounts by the chemical industry. It is widely used as solvent in paint removers, acetate film production, pharmaceutical processes, metal degreasing and as an aerosol propellant. Due to its low boiling point (40.1 ºC) and high vapour pressure (47kPa at 20 ºC) significant amounts of DCM reach the environment via gaseous emissions. The global consumption of DCM in 2004 was about 600000 tonnes and it is still growing further (Wang and Chen, 2006).

505

506

L. BAILÓN, Y. DOPICO, M. NIKOLAUSZ, M. KÄSTNER, M. C. VEIGA AND C. KENNES

Although some DCM containing wastes are incinerated, it can be assumed that the greater part of the DCM produced is eventually lost into the environment. Besides being present in gaseous emissions, DCM is also detected in many aqueous industrial effluents (Hartmans and Tramper, 1991). As most halogenated compounds, DCM is considered a health hazard. It is an irritating compound with fragrant odour, which is harmful to the respiratory system and central nervous system. Besides, it is a potential human carcinogen. Exposure to high concentrations may cause unconsciousness and death. Its MAK-value (maximal concentration at work) is 360 mg.m-3. The emission control of DCM and other VOCs has been the subject of recent environmental regulations in several countries. Therefore, a considerable interest exists in the development of techniques for the elimination of DCM and other chemicals from waste gases. The conventional control technologies for VOCs treatment, such as thermal incineration and wet scrubbing are usually costly, mainly when the pollutant concentration is low and the air flow rate is large. Biological techniques, using biofilters, bioscrubbers, biotrickling filters, suspended-growth bioreactors, or membrane bioreactors among others, have been studied and applied successfully over the past decades to solve problems of polluted air emissions containing VOCs and odours (Kennes and Veiga, 2001). These biotechnological methods often exhibit similar or even higher efficiencies than the traditional physical-chemical processes. They avoid the need of expensive catalysts, do generally not generate secondary streams that have to be treated again, and present lower operating costs (Gadre, 1989; Groenestijn van and Kraakman, 2005; Kennes and Thalasso, 1998). DCM can be readily degraded under aerobic and anaerobic conditions by several different microbial genera using it as their sole carbon and energy source. The aerobic degradation of DCM yields 1 mole of carbon dioxide and 2 moles of hydrogen chloride, with medium acidification, according to the following reaction (Herbst and Wiesmann, 1996; Kennes et al., 2006): (1) DCM degrading bacteria are isolated quite readily from activated sludge and many water and soil samples. Strains belonging, among others, to the genera Hyphomicrobium, Pseudomonas and Methylobacterium have been described (Brunner et al., 1980; Stucki et al., 1981; Gälli and Leisinger, 1985). Free liquid phase bioreactors have proven to be superior in the removal of acid producing pollutants than biofilters, as pH control and removal of metabolites are easier. This paper presents the results of an investigation of DCM removal from air in a Biological Trickling Filter (BTF) and in a Continuous Stirred Tank Bioreactor (CSTB). Studies have very recently also been started to compare the results obtained

REMOVAL OF DICHLOROMETHANE FROM WASTE GASES USING A FIXED

507

with the BTF and CSTB of this work to those that would be reached with the same kind of reactors and conditions but working with a two-phase system, above all to check the effect of shock loads. The presence of a free liquid phase in the BTF and CSTB is interesting because of the non-negligible water solubility of DCM (Kennes et al., 2006). Besides, the addition of an organic phase makes sense when dealing with fluctuating pollutant concentrations and shock load-conditions in order to buffer load variations.

2 MATERIALS AND METHODS 2.1 BACTERIA AND MEDIUM The BTF and the CSTB were inoculated with a biomass suspension of Hyphomicrobium KDM2 and KDM4 cultivated in shake-flasks. These dichloromethane-degrading strains have been described previously (Nikolausz et al., 2006) and belong to the Department of Bioremediation of the UFZ-Centre for Environmental Research Leipzig-Halle GmbH in Germany. The mineral medium used in this studuy contained per litre of distilled water: 1.5 g KH2PO4, 4.69 g Na2HPO4.12H2O, 0.5 g (NH4)2SO4, 0.2 g MgSO4.7H2O, 1 ml of a trace mineral solution and 1 ml of a vitamins solution. The trace mineral solution contained per litre of distilled water: 5.3 mg CaCl2, 2 mg FeSO4.7H2O, 0.2 mg MnSO4.5H2O, 0.2 mg CuSO4.5H2O, 0.2 mg ZnSO4.7H2O, 0.03 mg H3BO3, 0.4 mg CoCl2 and 4 mg Na2MoO4.2H2O. The vitamins solution contained per litre of distilled water: 0.2 mg biotin, 0.2 mg folic acid, 0.5 mg riboflavin, 0.5 mg thiamine, 0.5 mg nicotinic acid, 0.001 mg vitamin B12, 0.5 mg p-aminobenzoic acid, 1 mg pyridoxamine and 0.5 mg lipoic acid. The pH of the medium was adjusted at 7. 2.2 FIXED BED BIOTRICKLING FILTER A scheme of the biotrickling filter is shown in Figure 1. The reactor consisted of a glass column of 0.09 m internal diameter, with a cone at the bottom and at the top. A total working volume of 2.1 litre was filled with lava rock and the liquid medium was distributed or collected by means of perforated plates at the top and bottom of the column. The waste gas was introduced at the top of the filter in co-current operation. The polluted gas stream was artificially created by mixing two air streams. A small stream of air was bubbled through a vaporization flask containing pure dichloromethane and was mixed afterwards in a mixing chamber with a large pure air stream. Gas phase concentrations ranging from 0.1 to 15.7 g.m-3 were obtained by changing the flow rate of the dichloromethane stream. The total gas flow rate was kept constant at 0.084 m3.h-1. The gas velocity was 13.2 m.h-1 and the empty bed retention time 90 s.

508


Figure 1. Scheme of the laboratory biotrickling filter system.

The gas flow rates for air and DCM were adjusted utilizing two rotameters with valve (Brook Sho-Räte and Aalborg) and the liquid velocity was kept at 6.9 m.h-1 through a Watson Marlow peristaltic pump. The reactor was maintained at room temperature (21 ± 2 ºC). The liquid in the holding tank was gently mixed with a magnetic stirrer. The pH of the culture was kept between 6.9 and 7.05, with the aid of a pH controller (D09765T, Labprocess) coupled to an electrovalve, by dosing a 2 N NaOH alkaline solution to neutralize the HCl formed during the biological DCM degradation. This resulted in NaCl accumulation. High concentrations of this compound are expected to inhibit the biological activity (Gälli and Leisinger, 1985; Ottengraf et al., 1986; Hartmans and Tramper, 1991; Okkerse et al., 1999a, Diks et al., 1994). Thus, the conductivity was continuously measured in the holding tank (D09765T, Labprocess) and was kept below 28 mS.cm-2 by intermittent draining of liquid from the reactor. To maintain the required level of inorganic nutrients and compensate for drain and evaporation, fresh medium was added when needed. The limit of 28 mS.cm-2 was obtained from a salt tolerance test described in the Results. Air samples were taken with a Hamilton gas tight syringe at the inlet and outlet ports of the bioreactors. The DCM concentration of the samples was determined using a gas chromatograph, Hewlett Packard HP 6890 GC, equipped with a flame ionisation detector. 2.3 CONTINUOUS STIRRED TANK BIOREACTOR (CSTB) The experimental set-up is shown in Figure 2. The bioreactor consisted of an air-tight cylindrical glass vessel with a total volume of 2 litre (BioFlo 110, New Brunswick Scientific). The BioFlo-reactor was equipped with a PCU, a level control


509

module, a dO2/pH control module, a power controller (with rotameter), a 4 peristaltic pumps unit and an injection and extraction ports for culture medium. The temperature was controlled by a water jacket with thermostated water. Mixing was achieved by a turbine stirrer.

Figure 2. Scheme of the continuous stirred tank bioreactor.

The vessel was filled with 1.5 litre of biomass suspension. The temperature was kept at 30 ºC and the stirrer at 400 rpm. As in the BTF, the pH of the culture was kept at 7 utilizing 2 N NaOH as a neutralizing agent, added automatically by means of a peristaltic pump. The conductivity was here also kept below 28 mS.cm-2. Fresh medium and biomass culture were added and removed continuously in order to maintain a constant salt concentration. The inlet and outlet liquid flow rates were controlled to keep an equilibrium between conductivity and biomass concentration; being the total reactor volume constant. The removed biomass solution was decanted and the supernatant or the settled biomass were recycled to the reactor when needed. The waste gas stream was created in the same way as for the BTF. The flow rate was also 0.084 m3.h-1 while the gas empty bed retention time was 64 s. The DCM concentration was determined at the inlet and outlet gas streams of the bioreactor in the same way as in the trickling biofilter.

510


2.4 SALT TOLERANCE TEST In this test the degradation of DCM was evaluated in batch cultures at NaCl concentrations ranging from 0 to 500 mM. The experiment was carried out in 250 ml Erlenmeyer flasks containing 30 ml mineral medium inoculated with 10 ml of microbial suspension from the CSTB, previously centrifuged and diluted with fresh medium to minimize the amount NaCl that it could contain. After the addition of 10 ml pure DCM, the flasks were sealed with a screw cap containing two septa. The one in contact with the content of the flask was made of Viton, an inert material towards DCM, while the second septum, of ordinary rubber, was put on top of it in order to guarantee sealing. Subsequently, the flasks were vigorously shaken by hand until the compound had dissolved. To keep the liquid and gas phases in equilibrium during the rest of the experiment the Erlenmeyers were shaken horizontally at 120 rpm and 30ºC. Under these equilibrium conditions the degradation of DCM was followed by determining the gas phase concentration by gas chromatography as a function of time.

3 RESULTS AND DISCUSSION 3.1 BIOREACTORS PERFORMANCE The performance curves for both bioreactors are plotted in Figure 3.

Figure 3. Elimination capacity of both bioreactors versus the DCM inlet load.

The biotrickling filter had a maximal elimination capacity (EC) of around 170 g.m-3.h-1 (Figure 3). Up to inlet loads of 25 g.m-3.h-1 the BTF presented removal efficiencies (RE) between 85-100%. Between 25 and 170 g.m-3.h-1 the efficiency decreased sharply

511


from 85 to 50%; and then more smoothly down to 45% at 350 g.m-3.h-1, which is the pseudo-critical load of the performance curve, where the maximal elimination capacity was reached. At higher loads of 600 g.m-3.h-1 the RE was only 28%. These results of RE and EC are in the same range as those obtained by other authors (Table 1) (Kennes et al., 2006). The continuous stirrer tank bioreactor shows a maximal elimination capacity of around 100 g.m-3.h-1 (Figure 3). High removal efficiencies between 90-100% are found for inlet loads up to 120 g.m-3.h-1. At this point the maximal elimination capacity was reached and the RE quickly decreased at higher loads. At 350 g.m-3.h-1 the RE was about 25%. Comparing the two different bioreactor configurations studied here, it can be concluded that for low inlet DCM loads the CSTB exhibits a better performance, while that for higher loads the BTF works better. The changing point happens at the cross point of the performance curves of both reactors, once the maximum removal efficiency of the CSTB is reached (Figure 3). To the best of our knowledge, no previous study has been published on the removal of DCM in a CSTB and on its comparison with its removal in a DCM-treating BTF. Table 1. Air biotrickling filters for the removal of DCM working at neutral pH and T ~20-30ºC. Bacteria

Packing material

Inlet range (g.m-3)

Gas / liquid velocity (m.h-1)

EBRT (s)

RE (%)

EC (g.m-3 h-1)

Hypomicrobium Sp. GJ21 Hypomicrobium DM20 Hypomicrobium Sp. GJ21 Hyphomicrobium KDM2 & KDM4

Keramic Novalox Saddles Polypropylene

References

0.5-10

160 / 36

29

32

150 (max)

0.0660.727

233 / 15.15

16

87.2 59.3

12.68 103.51

PVC

2

163 / 7.3

60

84

102

Okkerse et al., 1999a

Lava rock

0.1-15.7

13.2 / 6.9

90

90 45

13 170 (max)

This study

Diks and Ottengraf, 1991a, b Hartmans and Tramper, 1991

512


After somewhat more than one year operation samples were taken from both bioreactors for identification of the dominant microorganisms by molecular techniques. It was observed that the originally inoculated Hyphomicrobium strains remained dominant in both bioreactors together with some other new species, despite working under completely non-sterile conditions as requested for practical environmental applications. The identification process was carried out at the UFZ-Centre for Environmental Research in Leipzig (non published data). To check the stability of the bioreactors both systems were subjected to 6 hours overloads. In a first experiment, the load was increased from approximately 70 g.m-3.h-1 to 250 and, later, 460 g.m-3.h-1 and, in a second experiment, it was increased from 15 g.m-3.h-1 to about the same maximum values. In all cases the recovery of both systems was almost immediate. Total recovery when the original inlet load was 70 g.m-3.h-1 took less than 1 hour; and less than 3 hours were required when the original load was 15 g.m-3.h-1. 3.2 SALT TOLERANCE TEST As DCM degradation is accompanied by HCl formation, which results in NaCl accumulation when neutralized with NaOH, the salt tolerance of our culture was tested. In Figure 4 it can be observed that at a concentration of 300 mM NaCl some limited inhibition of the bacterial activity is noticeable. At a concentration of 500 mM NaCl the inhibition becomes much more significant. Therefore, to avoid biological inhibition, the concentration of NaCl in the liquid medium of the bioreactors must be kept below 300 mM, which corresponds to a conductivity of 28 mS.cm-2 or less.

Figure 4. Degradation of DCM in batch cultures at different NaCl concentrations.


513

Ottengraf et al. (1986) found complete inhibition at concentration levels exceeding 250-300 mM NaCl, while Hartmans and Tramper (1991) reported high inhibitory effects with 200 mM NaCl. In other bioreactor studies, the NaCl concentration was kept below 100-150 mM (Diks and Ottengraf, 1991a; Okkerse et al., 1999a). In a study of Diks et al. (1994), microbial growth of Hyphomicrobium GJ21 was shown to be strongly inhibited at NaCl concentrations exceeding 350 mM. Nevertheless, for a «TF-enrichment culture» from a laboratory trickling filter degrading DCM, initially inoculated with Hyphomicrobium GJ21, inhibition was less severe as good growth was observed up to 600 mM NaCl. So, it seems that some adaptation towards increased salt concentrations could take place in continuously operating trickling biofilters.

4 ACKNOWLEDGEMENTS The present research was partly financed by the Spanish Ministry of Science and Education (project CTM2007-62700/TECNO) and European FEDER funds, as well as the Xunta de Galicia (project PGIDIT05PCIC10304PN). We acknowledge the collaboration of Hana Simova during part of the experimental study.

REFERENCES Brunner, W.B., Staub, D. and Leisinger, T. (1980) Bacterial degradation of dichloromethane. Appl. Environ. Microbiol. 40: 950-958. Diks, R.M.M. and Ottengraf, S.P.P. (1991a) Verification studies of a simplified model for the removal of dichloromethane from waste gases using a biological trickling filter (Part I). Bioproc. Eng. 6: 93-99. Diks, R.M.M. and Ottengraf, S.P.P. (1991b) Verification studies of a simplified model for the removal of dichloromethane from waste gases using a biological trickling filter (Part II). Bioproc. Eng. 6: 131-140. Diks, R.M.M., Ottengraf, S.P.P. and Van den Oever, A.H.C. (1994) The influence of NaCl on the degradation rate of dichloromethane by Hyphomicrobium sp. Biodegradation 5: 129-141. Gadre, R.V. (1989) Removal of hydrogen sulfide from biogas by chemoautotrophic fixed-film bioreactor. Biotechnol. Bioeng. 34: 410-414. Gälli, R. and Leisinger, T. (1985) Specialized strains for the removal of dichloromethane from industrial waste. Cons. Recycling 8: 91-100. Groenestijn, J.W. van and Kraakman, N.J.R. (2005) Recent developments in biological waste gas purification in Europe. Chem. Eng. J. 113: 85-91.

514


Hartmans S. and Tramper J. (1991) Dichloromethane removal from waste gases with a trickle-bed bioreactor. Bioproc. Eng. 6: 83-92. Herbst, B. and Wiesmann, U. (1996) Kinetics and reaction engineering aspects of the biodegradation of dichloromethane and dichloroethane. Water Res. 30: 1069-1076. Kennes, C. and Thalasso, F. (1998) Waste gas biotreatment technology. J. Chem. Technol. Biotechnol 72: 303-319. Kennes, C. and Veiga, M.C. (2001) Bioreactors for Waste Gas Treatment, Kluwer Academic Publishers, Dordrecht, The Netherlands. 312 pp. Kennes, C., Jin, Y. and Veiga, M.C. (2006) Fungal and dechlorinating biocatalysts in waste gas treatment. In: (Lens, P., Kennes, C., LeCloirec, P. and Deshusses, M.A., Eds), Waste Gas Treatment for Resource Recovery, IWA Publishing Co., London, UK, p. 277-301. Nikolausz, M., Nijenhuis, I., Ziller, K., Richnow, H. and Kästner, M. (2006) Stable carbon isotope fractionation during degradation of dichloromethane by methylotrophic bacteria. Environ. Microbiol. 8: 156-164. Okkerse, W.J., Ottengraf, S.P.P., Osinga-Kuipers, B. and Okkerse, M. (1999) Biomass accumulation and clogging in biotrickling filters for waste gas treatment. Evaluation of a dynamic model using dichloromethane as a model pollutant. Biotechnol. Bioeng. 63: 418-30. Ottengraf, S.P.P., Meesters, J.J.P., Van den Oever, A.H.C. and Rozema, H.R. (1986) Biological elimination of volatile xenobiotic compounds in biofilters. Bioproc. Biosys. Eng. 1: 61-69. Stucki, G., Gaelli, R., Ebershold, H.R. and Leisinger, T. (1981) Dehalogenation of dichloromethane by cell extracts of Hyphomicrobium DM». Arch. Microbiol. 130: 366-371. Wang, J. and Chen, J. (2006) Removal of dichloromethane from waste gases with a bio-contact oxidation reactor. Chem. Eng. J.123: 103-107.

515

MICROBIAL ECOLOGY AND BIOFILMS


517

Development of a reliable extraction method for the recovery of total genomic DNA from woodchip colonizing biofilm involved in gas biofiltration LÉA CABROL1, LUC MALHAUTIER2, JANICK ROCHER2, FRANCK POLY3, XAVIER LE ROUX3, MARC JOVIC1, ANNE-SOPHIE LEPEUPLE1 AND JEAN-LOUIS FANLO2 1

Anjou Recherche-Veolia Water, Chemin de la Digue, BP76, 78600, Maisons Laffitte, France Laboratoire Génie de l’Environnement Industriel, Ecole des Mines d’Alès, Avenue de Clavières 6, 30319, Alès Cedex, France 3 Laboratoire Ecologie Microbienne du Sol, UMR-CNRS 5557, Bâtiment Gregor Mendel, Université Claude Bernard Lyon I, 43, boulevard du 11 Novembre 1918, 69622, Villeurbanne Cedex, France 2

ABSTRACT This preliminary study focused on a critical step for the characterization of microbial ecosystem involved in biofiltration. Two aspects of nucleic acid recovery were explored: (i) cell dispersion (three methods tested) and (ii) total DNA extraction (four methods tested). The objective is to select the optimal combination of desorption/extraction methods, allowing subsequent molecular investigations to be reliable. Three relevant criteria are used to assess extraction efficiency: DNA amount and purity, and subsequent amplification feasibility.

1 INTRODUCTION During the past two decades, most studies concerning biofiltration concentrated on two main objectives: (i) assessment of operating parameter impact (e.g. packing material impact, pH effect: Kim et al., 2000; Prado et al., 2006); and (ii) definition of the system limits under different loading conditions (Aizpuru et al., 2001; VergaraFernandez et al., 2007). Both problematic were based on elimination performance evaluation, without regard for intrinsic biological phenomena. It was noticeable that a

517

518

L. CABROL, L. MALHAUTIER, J. ROCHER, F. POLY, X. LE ROUX, M. JOVIC, A. S. LEPEUPLE AND J. L. FANLO

number of studies tended to consider the system according to a «black-box» approach, while in a biofilter the pollutant removal is mainly due to the microbial component, whose structure and activity still remain unclear or even unknown. Microbial ecology aims to characterize microbial communities by their structure (i.e. their diversity, stability, spatial and temporal dynamics, occurrence of specific groups), as well as their interactions with the environment. Along the last decade, staggering progress in molecular biotechnologies offered powerful tools which made possible the fine characterization of microbial communities, granting access to uncultivable microorganisms. Among these molecular tools are quantitative PCR, fingerprintings, clone libraries sequencing, and, more recently, metagenome shotgun sequencing. They have been applied in various ecosystems, such as soil (Patra et al., 2006), sea sediments (Venter et al., 2004), anaerobic sludge (Godon et al., 1997), wastewater treatment biofilters (Ahn et al., 2004). More and more studies are carried out to elucidate community structures in gas biofilters. To date, several molecular tools have been used to gain insight into the dynamic diversity of bacterial communities in biofilters: ARISA (Steele et al., 2005), SSCP (Khammar et al., 2005), RFLP (Khammar et al., 2005), amoA gene PCR, cloning and sequencing (Sakano and Kerkhof, 1998), DGGE on 16S rDNA (Sercu et al., 2005; Cai et al., 2006; Li and Moe, 2004; Shim et al., 2006; Chung, 2007), DGGE on 16S rRNA (Sercu et al., 2006), FISH (Friedrich et al., 1999; Friedrich et al., 2003). Adopting a microbial ecology approach is of prime interest to reach a better understanding of biological mechanisms occurring within a biofilter. This better understanding may help to control, stabilize and optimize the biological process. Biofilms involved in biofilters are constituted of a complex and uncharacterized microflora attached to the packing material. Thus an essential preliminary task for the investigation of microbial communities with molecular tools is to implement and optimize a methodology for the recovery of nucleic acids. To get samples as representative as possible of the initial diversity, this methodology has to be the least selective as possible. In other ecosystem studies, such as soil, a lot of work was carried out to compare and implement DNA recovery methods that exhibit an unbiased sampling of the investigated community (Robe et al., 2003). However, within biofiltration context, very little attention was paid to the methodological aspects of nucleic acids recovery, despite their crucial importance in final results significance (Khammar et al., 2004; Li and Moe, 2004). DNA recovery methods are very heterogeneous and have neither been standardized nor optimized to date. Indeed, they are hugely dependent on the packing material specificities (size, organic/inorganic nature, density, hardness, porosity). In this preliminary study we explored methodological aspects of the nucleic acids recovery from microbial communities involved in a laboratory scale biofilter

DEVELOPMENT OF A RELIABLE EXTRACTION METHOD FOR THE RECOVERY

519

filled with pine bark woodchips. Considering the packing material used in this work, direct DNA extraction could not be applied. Hence two successive steps had to be performed: cell desorption (crushing, shaking, sonication) and DNA extraction (three commercial kits –two of which being specific for soil– and a reference protocol). The objectives of the present work were (i) to optimize cell desorption from the packing material, and (ii) to select the optimal combination of desorption and extraction methods. To assess DNA recovery efficiency, importance was attached to three relevant criteria: extracted DNA amount and purity, as well as subsequent amplification yield (this latter data will only be presented orally).

2 MATERIALS AND METHODS 2.1 FIRST STEP: OPTIMIZATION OF CELL DESORPTION 2.1.1 BIOFILTER SETUP Experiments were conducted on the biomass which colonized a lab-scale biofilter (1 m height, 125 mm diameter) filled with pine bark woodchips (initial porosity of 37%) and treating an H2S stream (10 mgH2S/m3; 500 m/h). The packing material was kept at constant humidity by regularly spraying a salt mineral nutrient solution, whose composition was previously described (Lalanne et al., 2007), at a rate of 150 mL every six hours. The biofilter was run at ambient temperature. 2.1.2 DESORPTION METHODS AND OPERATING CONDITIONS FOR OPTIMIZATION Three commonly used detachment methods were investigated for microorganism removal from woodchip support: blending (performed by an Ultra-Turrax -T25 basic, Ika); shaking (performed by a Vibro-Shaker –Retsch MM200); sonication (performed by an ultrasonic bath -Branson sonifier bath, Energy). One or two parameters were retained as potentially influent to define optimal conditions for each method, as shown in Fig.1. Other parameters (revolution speed, rotating frequency) were maximal.

Figure 1. Desorption methods and optimization conditions.

520


Sonication duration range was chosen from previous results (Khammar et al., 2004), which demonstrated that ultrasonic treatment needed higher duration to suspend microorganisms. On the contrary, blending treatment did not need more than two minutes for a complete homogenization of the suspension. 2.1.3 SAMPLING SCHEME At the time of sampling, the biofilter had reached a steady state, with complete H2S removal. Each treatment was carried out in triplicate, on woodchips extracted from the biofilter at the same time and at the same location (about 0.5 m high). Sampling procedure is shown in Fig. 1. Each sample was constituted of 5 g of woodchips, suspended in 15 mL of sterile physiological serum (NaCl, 9 g/L). Aliquots of 1 mL of the liquid suspension were collected and enumerated. 2.1.4 MICROSCOPIC DIRECT COUNTS Total bacteria were enumerated by fluorescence microscopy using DAPI staining (Sigma, USA). Whole experiment is done in sterile conditions. After serial dilutions, samples were stained with DAPI at a final concentration of 20 μg/mL during 30 minutes in the dark in a shaker (200 rpm). Stained bacteria were collected on 0.2 μm polycarbonate membrane filters (Millipore GTBP, Ireland) by vacuum microfiltration. Filters were mounted on microscope slides in Mounting Medium (Sigma, USA) and observed with an epifluorescence microscope (Leica DMLB) equipped with a blue excitation filter (BP 340-380 nm) and a barrier filter LP 425. Thirty microscopic fields per slide were enumerated. 2.2 SECOND STEP: COMBINATION OF CELL DESORPTION AND DNA EXTRACTION 2.2.1 BIOFILTER SETUP Experiments were conducted on the same pilot-scale unit, but the biofilter was treating a VOC mixture made of acetaldehyde, acetone, butanal, MEK, DMDS, butanoic acid, isovaleric acid. At the time of biomass sampling, operating conditions were as follows: 10 mg/m3 for each compound; gas velocity at 100 m/h. 2.2.2 EXPERIMENTAL PROCEDURE The experiment involved 12 samples, each constituted of 5 g of woodchips suspended in 15 mL of sterile physiological serum (NaCl, 9 g/L). Each sample was repeated twice. As shown in Figure 2, each sample was submitted to one of the three desorption methods, under previously optimized conditions (described in section 3.1.). After centrifugation of liquid phase at 10 000 rpm for 10 minutes, the pellet was subjected to DNA extraction, using one of the four following methods: I. PowerSoil DNA Kit, MoBio (Ozyme, France); II. FastDNA® SPIN Kit for Soil, Qbiogene (MP Biomedicals, France); III. NucleoSpin® Tissue Kit (Macherey –Nagel, France); IV. An extraction protocol adapted from Godon et al. (1997).


521

Figure 2. Schematic procedure to assess the optimal combination of desorption and extraction methods (in dotted lines: experiments in prospect).

2.2.3 EXTRACTION AND PURIFICATION OF TOTAL GENOMIC DNA Extraction by commercial kits was performed according to the manufacturers’ instructions, using a Vibro Shaker Retsch for cell disruption. In all cases, elution volumes were 50 μL. The fourth method was slightly modified from the one described by Godon et al. (1997), as follows, to ensure the largest sample size. The microbial cell fractioncontaining pellet obtained after desorption and centrifugation was resuspended in 385 μL of 4M guanidine thiocyanate-0.1 M Tris (pH 7.5) and 115 μL of 10%-Nlauroyl sarcosine. Samples were stocked at -20°C. After the addition of 500 μL of 5% -N-lauroyl sarcosine-0.1 M phosphate buffer (pH 8.0), the sample was incubated at 70°C for 1 h. One volume (500 μL) of 0.1 mm-diameter sterile zirconium beads (Sigma) was added and the sample was shaken at maximum speed (30 Hz) for 10 min in a Vibro Shaker (Retsch MM200). Polyvinylpolypyrrolindone (PVPP, 15 mg) was added. The sample was vortexed and centrifuged for 3 min at 12 000 rpm. The supernatant was recovered. The pellet was washed with 500 μL of TENP (50 mM Tris [pH 8.0], 20 mM EDTA [pH 8.0], 100 mM NaCl, 1 % PVPP) and centrifuged for 3 min at 12 000 rpm. The new supernatant was pooled with the first one. The washing step was repeated three times. The pooled supernatants were centrifuged for 3 min at 12 000 rpm to remove particles, and then split into two 2-mL tubes. Samples were incubated for 1h30 at 56 °C with 30 μL of proteinase K (20 mg/ mL), and then incubated 1 h more at 37°C with 20 μL of RNase A (10 mg/mL). Samples were split into 500 μL subsamples and crude DNA was purified by addition of 1 mL of phenol-chloroforme-isoamyl alcohol (25:24:1). After centrifugation for 5

522


min at 10 000 rpm, the upper phase was recovered. Nucleic acids were precipitated by the addition of 50 μL of sodium acetate 3M and 1 mL of cold absolute ethanol. Samples were incubated for 15 min at -80°C and 30 min at -20°C. After centrifugation for 30 min at 14 000 rpm, supernatant was discarded and DNA pellet was washed with about 1 mL of cold 70% ethanol, dried for 10 min at room temperature and resuspended in 50 μL of Tris-EDTA 0.1X. 2.2.3 EVALUATION OF DNA RECOVERY EFFICIENCY Sizing and quantification of extracted DNA were assessed by electrophoresis. 5 μL of extraction product were loaded in 1% agarose gel. Migration was performed at 85 V, for 45 min, in 1X TAE buffer and gel was stained with ethydium bromide. DNA amount was further determined by absorbance at 260 nm using an UV spectrophotometer (Biophotometer, Eppendorf). DNA concentration was calculated considering that 1 UDO corresponds to a double-strand DNA concentration of 50 μg/ mL, in 1 cm cuvettes. DNA purity was determined by the ratio of absorbance at 260 nm and absorbance at 280 nm (Biophotometer, Eppendorf), considering that the absorbance at 280 nm is mainly due to protein contamination.

3 RESULTS AND DISCUSSION 3.1 OPTIMIZATION OF CELL DESORPTION Bacterial counts after desorption are presented in Figure 3. When a single parameter was variable (treatment duration), results were statistically analyzed by ANOVA (analysis of variance). After verifying variance homoscedasticity with a Hartley test, the significance of differences between means was established by the Fisher-Snedecor test with a risk level of 0.05. It appeared that for blending desorption (Figure 3.A), treatment duration between 0.5 and 2 minutes did not influence the amount of recovered cells. Nevertheless, increasing blending duration led to more deviation. Indeed the longer was the blending, the more organic particles were suspended, which seriously hampered microscopic counting, thus leading to higher experimental errors. On the contrary, cell counts obtained after different sonication durations were not statistically equal (Fig. 3.B): cell removal was significantly improved by increasing sonication duration up to 60 minutes. Concerning shaking desorption, a Doehlert matrix was built. In the model provided by NemrodW analysis, the most significant coefficient is the constant one, i.e. the coefficient linked to none experimental factor (data not shown). Hence it can be concluded that neither shaking duration nor shearing force increase (by adding glass beads) improved shaking efficiency. These results are in accordance with those of Khammar et al. (2004), which detected no significant effect of glass


523

Figure 3. Influence of treatment duration and glass bead amount on total microbial cells recovered after desorption and enumerated by DAPI (A: blending by Ultra-Turrax; B: Sonication; C: Vibro-Shaking). Graphs are in logarithmic scale and error bars represent standard deviation calculated on triplicates.

524


beads and treatment duration on cell detachment from peat and activated carbon by blending and shaking and highlighted treatment duration effect for sonication. As a result, optimal conditions for desorption treatments are chosen as follows: 1 min for blending; 60 min for sonication; 10 min without glass beads for shaking. Blending could have been thought to be the most efficient detachment method, as it allowed recovery and homogenization of the initial material in its entirety (no biomass was lost), while sonication and shaking only suspended microorganisms (a fraction of biomass left on the support may be lost). This was observed by Khammar et al. (2004). Nevertheless, in this previous study, biomass detachment was only evaluated by the number of viable and cultivable microorganisms. As enumeration results were in the same range whatever was the desorption method, whole three methods had to be further compared, on the basis of more accurate criteria. 3.2. OPTIMAL COMBINATION OF CELL DESORPTION AND DNA EXTRACTION METHODS Electrophoresis results are shown in Fig. 4. Only one sample is presented for each duplicate (except for MoBio extraction method, where results were not reproducible).

Figure 4. Electrophoresis of DNA extracted by four different methods, after three different desorption treatments (UT: UltraTurrax; US: UltraSonication; VS: VibroShaking).

Conclusions drawn from band intensity observations were confirmed and completed by absorbance measures, as shown in Fig. 5. It is important to note the bad reproducibility of the results between duplicates. This would be explained by the random aspect of bacterial colonization on the initial 5g-sampling. It is obvious that Godon-adapted extraction protocol led to significantly higher DNA amounts, when compared to commercial kit extraction (about 15 times higher on average). This is observed independently of the previous desorption methods used. But this protocol being highly time-consuming, it is not realistic to envisage its use for routine DNA extraction of numerous samples. It is thus considered as a reference protocol.


525

A

B Figure 5. DNA concentration and purity after different methods of desorption and extraction, evaluated by A260 and ratio A260/A280, respectively (concentrations shown for both duplicates).

It should be noted that, whatever the extraction method was, desorption treatment by Ultra Turrax was clearly not suitable for high DNA extraction yield. DNA concentration after kit extraction was about 20 μg/mL and did not exceed 400 μg/mL with reference extraction protocol. Indeed, Ultra Turrax treatment led to a single-phase suspension where the whole initial material was homogenized. Therefore, in the pellet obtained after centrifugation, the relative proportion of cells was very low compared to the proportion of crushed wooden material. Moreover, blending detachment gave the worst results in terms of DNA purity (Fig. 5.B), probably because of the high amount of organic material in blended samples. In all cases, DNA recovery was higher when previous desorption was performed by Vibro Shaker. After shaking detachment, DNA concentration was about 130 μg/mL when kit-extracted and even reached 2000 μg/mL

526


with reference protocol. Cell removal by sonication gave intermediate results in terms of DNA recovery. DNA extraction by MoBio kit was the least efficient for DNA recovery ( 0.05). As a result, a comparison of mechanical agitation and sonication was realised using an application time of 5 minutes and physiological serum as a chemical dispersion agent. The coupling of mechanical agitation and sonication was also evaluated. Statistical analyses of DAPI enumerations indicated that mechanical agitation desorbed significantly more biofilm than sonication (Student test, p < 0.05). Nevertheless, as shown in Figure 1, coupling sonication and mechanical agitation did not have a significant effect on biofilm’s desorption efficiency (Student test, p > 0.05). Finally, the desorption method selected for biofilter monitoring was a 5 minutes mechanical agitation in physiologic serum (8% (p/v) NaCl solution) using a Vortex (Scientific Industries) on position 10.

544

M. JOVIC, L. CABROL, F. DUCRAY, R. GAGNEUX AND A. S. LEPEUPLE

Figure 1. Comparison of three methods for biofilm extraction. The three methods compared were: (i) mechanical agitation (Vortex), (ii) sonication and (iii) coupling sonication/Vortex. Tests were realised on biofilm’s carrier media of biofilter B.

3.2 MONITORING OF POLLUTANT REMOVAL The monitoring of chemical and physical parameters confirmed that biofilters A and B were fed by VOCs and ammonia. For ammonia inlet concentrations varied from 20 to 40 mg/m3. During the time of the study, determination of the abatement of ammonia did not show significant differences for biofilters A and B. As shown in Figure 2, ammonia was fully eliminated by biofilters A and B after 90 days of functioning. Moreover EC for biofilters A and B was close to 2 g/m3/h.

Figure 2. Graphical representation of NH3 abatement in biofilters A and B during the time of the study. NB: for technical reasons the monitoring of VOCs was not performed. However, inlet concentrations were measured at the beginning of the study. They were close to 5 ppmv CH4 equivalent.

MONITORING AND CHARACTERIZATION OF BACTERIAL POPULATIONS

545

3.3 MONITORING OF BIOFILTERS MICROBIOLOGICAL PARAMETERS 3.3.1 TOTAL AND VIABLE BIOMASS MONITORING As described in Figure 3, densities of microorganisms observed on biofilters A and B were in agreement with those reported on other biofilters (Khammar et al., 2004). Statistical analyses of total biomass for biofilters A and B indicated a significant evolution of colonization of biofilm carrier media with time (ANOVA, p < 0.05). Nevertheless, as presented in Figure 3, modifications of the biomass concentrations were not relied with the start up of biofilters A and B. It was conclude that these changes were due to sampling fluctuations or biofilm washing.

Figure 3. Representation of viability rates® and colonization densities (DAPI, TVC) of carrier media of biofilters A and B.

546


As reported in Figure 3, concentration of viable biomass in both biofilter A and B presented significant modifications during the experiment (Kruskall-Wallis test, p < 0.05 (biofilter A) and ANOVA, p< 0.05 (biofiltrer B)). For biofilter B, modifications of the viable biomass concentration were attributed to sampling fluctuations or biofilm washing. However, for biofilter A, a 1-log increase in TVC enumeration appeared during the first week. This increase was relied to nutriments provision by gaseous effluent. The mean of TVC enumerations was calculated during the entire period of the study. It appeared that concentrations of viable biomass in both biofilters A and B were not significantly different (Student test, p > 0.05). Nevertheless, cellular viability rate was higher in biofilter B than in biofilter A. As indicated in Figure 3, cellular viability rate was ranged between 5 to 40 % for biofilter A and between 22 to 100 % for biofilter B. Also, at t = 0, it appeared that viable biomass concentration in carrier media B was significantly greater than in carrier medium A. It is important to point out that, as the carrier medium could be the only source of nutriments in abnormal working conditions, the biomass would better resist to stops of feeding in a biofilter containing medium B. 3.3.2 AOB AND NOB POPULATION MONITORING During the time of the study, the monitoring of nitrifying bacteria using FISH method did not show any presence of AOB and NOB in biofilters A and B. Positive results obtained with Eubacteria probe (EUB 338) underlined that there was not any problem of FISH probes accessibility in cells. Taking into account these results and ammonia abatement efficiency, the following hypothesis were assumed: (i) AOB and NOB were present in other strates of the biofilters (Khammar et al., 2004) (ii) ammonia was metabolized by others nytrifying bacteria genus, (iii) ammonia was not biodegradated but was absorbed by biofilm carrier media and (iv) ammonia was absorbed by biofilters scrubbing solution. 3.3.3 ANALYSIS OF TOTAL DNA BY 16S RDNA PCR-SSCP Analysis of 16S rDNA SSCP profiles based on similarity matrices allowed the comparison of microbial population of biofilters A and B. As shown in Figure 4, it appeared that microbial populations of biofilters A and B were modified in comparison to their initial state. The results also indicated that the most important variations occurred for biofilter A. As indicated in Figure 5, comparison of SSCP profiles with time for biofilter A and B established that microbial populations of both biofilters A and B converged and became nearly identical after 120 days of functioning. At this time, those microbial populations were closer than at their initial state.


547

Figure 4. graphical representation of evolutions of biofilters A and B microbial populations compared with their initial state.

Figure 5. graphical representation of evolution in time of similarities of biofilters A and B microbial populations.

Finally, as presented in Figure 6 comparison of 16S rDNA SSCP profiles between t and t-1, revealed that microbial populations of both biofilters A and B evolved and stabilized. Figure 6 demonstrated that biofilter B stabilized faster than biofilter A. After 60 days of functionning, biofilter B was stabilized whereas this state was reached after 120 days for biofilter A. Those results were attributed to the composition of

548


biofilm carrier media. Indeed, biofilter B was packed with a carrier media containing branches coming from the composting facility. Therefore, these results demonstrated that inoculation of biofilm carrier media with an acclimated microbial population significantly reduce biofilter start up phase.

Figure 6. Graphical representation of evolution of biofilters A and B microbial populations compared with their previous state (t-1 to t).

4 CONCLUSION AND PERSPECTIVES As a conclusion, it appeared that physical and chemical monitoring of biofilters A and B did not provide helpful information to select a biofilm carrier media. Indeed, during the time of the study, abatement efficiencies observed on the two biofilters were not significantly different. Microbiological parameters have provided valuable information. Indeed, it was established, by the quantification of viable biomass, that carrier media B allowed a greater supply in nutriments for the biomass than carrier media A. Dynamic analysis of microbial populations confirmed that important modifications of microbial populations occurred during the start up phases of the biofilters. It was also demonstrated that those modifications were dependant of initial states. Therefore, it appeared that biofilter’s stabilization phase could be reduced using an appropriate inoculum. Finally, the microbial approach proved its interest to better understand phenomena occurring during the start up phase of biofilters. It also demonstrated the interest to use molecular tools to caracterize a biofilm carrier media. Moreover, carrier


549

media B showed benefits compared to carrier media A, as it could significantly reduce the start up phase and provide a greater supply in nutriment. This last point could have a significative impact on biofilter’s stability in the case of feeding stops.

REFERENCES Aizpuru, A., Malhautier, L., Roux, J.C. and Fanlo J.L. (2001) Biofiltration of a mixture of volatile organic emissions. Air & Waste Manage. Assoc. 51: 1662-1670. Amann, R., Krumholz, L.and Stahl, D.A. (1990a) Fluorescent oligonucleotide probing of whole cells for determinative, phylogenetic and environmental studies in microbiology. J. Bacteriol. 172: 762-770. Amann, R., Binder, B.J., Olson, R.J., Chisholm, M.S.W., Devereux, R. and Stahl, D.A. (1990 b) Combination of 16S rRNA-targeted oligonucleotide probes with flow cytometry for analysing mixed microbial populations. Appl. Environ. Microbiol. 56: 1919-1925. Delbes, Leclerc, Zumstein, Godon and Moletta. (2001) A molecular method to study population and activity dynamics in an anaerobic digestor. Water Sci. Technol. 43(1): 51-57. Dijkshoorn, L., Towner, K.J. and Struelens, M. (2001) New approaches for the generation and analysis of microbial typing data. Elsevier. Friedrich, U., Naismith, M., Altendorf, K. and Lipski A. (1999) Community analysis of biofilters using fluorescence in situ hybridization including a new probe for the Xanthomonas branch of the class Proteobacteria. Appl. Environ. Microbiol. 65(8): 3547-3554. Gieseke, A., Purkhold, U., Wagner, M., Amman, R. and Schramm, A. (2001) Community structure and activity dynamics of nitrifying bacteria in a phosphate-removing biofilm. Appl. Environ. Microbiol. 67(3): 1351-1362. Godon, J., Zumstein, E., Dabert, P., Habouzit, F. and Moletta, R. (1997) Microbial 16S rDNA diversity in an anerobic digester. Appl. Environ. Microbiol. 63(7): 2802-2813. Khammar, N., Malhautier, L., Degrange, V., Lensi, R. and Fanlo, J.L. (2004) Evaluation of dispersion methods for enumeration of microorganisms from peat and activated carbon biofilters treating volatile organic compounds. Chemosphere 54: 243-254. Khammar, N., Malhautier, L., Degrange, V., Lensi, R., Godon, J.J. and Fanlo, J.L (2005) Link between spatial structure of microbial communities and degradation of a complex mixture of volatile organic compounds in peat biofilters. J. Appl. Microbiol. 98: 476-490. Khammar, N., Malhautier, L., Bayle, S. and Fanlo, J.L. (2005) Biofiltration of volatile organic compounds. Appl. Environ. Microbiol. 68: 16-22. Moter, A. and Gobel, U. (2000) Fluorescence in situ hybridization (FISH) for direct visualization of microorganisms. J. Microbiol. Met. 41: 85-112.


551

Biofilter response to biomass reactivation for VOC treatment ANA ELÍAS, ASTRID BARONA, GORKA GALLASTEGI, MIKEL LARRAÑAGA AND MARÍA FERNÁNDEZ Department of Chemical and Environmental Engineering, Faculty of Engineering, University of the Basque Country, Alda Urquijo s/n. 48013 Bilbao, Spain

ABSTRACT This research has undertaken a comparative study on using a fresh activated sludge or a refrigerated/ reactivated sludge as active biomass source for biofiltration purposes. A sludge sample was initially selected based on the ratio between volatile solid content and total solid content before and after refrigeration at 6 oC for 90 days. The degradation rate of the activated sample for three addition doses of toluene was established before and after refrigeration. The same procedure was also carried out for ethylbenzene and p-xylene after refrigeration/reactivation. Surprisingly, the degradation rate for toluene was higher after refrigeration and the results were very similar for an addition of 2 and 8 μL. Subsequently, one biofilter was inoculated with the activated sample and another with the reactivated sample, and both were fed with toluene ranging from 2.6 to 26.2 g toluene m-3 h-1. Concerning the elimination capacity of both biofilters, no relevant differences were found. It was concluded that the active biomass degrading toluene was not affected by refrigeration, in spite of the fact that the SV/ST ratio decreased after the storage period. The elimination capacity of the other two biofilters (ethylbenzene and p-xylene) was highly influenced by the gas flow rate.

1 INTRODUCTION Environmental regulations for pollution control are frequently enacted before «suitable» (affordable, effective and environmentally friendly) technologies have been fully developed. Amongst biotechnologies, biofiltration is a seemingly simple system whose effectiveness relies on the optimization of several operating parameters and the selection of a suitable packing material and degrading biomass.

551

552

A. ELÍAS, A. BARONA, G. GALLASTEGI, M. LARRAÑAGA AND M. FERNÁNDEZ

The biomass responsible for the degradation of the gaseous contaminant fed into the biofilter can be supplied by the support material itself (Barona et a.l, 2004), purchased from trade catalogues (Christen et al., 2002), isolated from other active bioreactors (Estévez et al., 2004) or even collected from locations contaminated with the target pollutant (González-Sánchez and Revah, 2007). Prior to the inoculation itself, the proper selection, storage, acclimation and activation of the microorganisms is crucial to ensure a long-lasting operation of the bioreactors. In fact, Prado et al. (2005) proved that the previous biomass concentration and biomass adaptation of the inoculum dramatically affected the start-up and performance of conventional biofilters treating methanol during the first stages of operation. Likewise, simple analytical techniques for biomass growth detection are needed. Several methods, such as extra-cellular enzymatic activity (Laurent and Servais, 1995) or electron transportation activity (Fontvieille and Moul, 1985), were developed by microbiologists to estimate biomass activity. Among other methods, respirometry, optical density at 600 nm, plate counting methods and 4’6-diamidino-2-phenylindole (DAPI) staining have also been used in literature (Hwang et al., 2003; Álvarez-Hornos et al., 2005; Kim and Jaffé, 2007). Other simpler analyses, such as carbon balance and volatile suspended solid content, are also practical tools for assessing biomass concentration and adaptation, although they have obvious disadvantages, such as no discrimination between living or dead biomass or no accounting for changes in microorganism physiology. Bearing in mind that biofilter controlling operators require simple and quick techniques for start-up and everyday operation, the objective of this study is to ascertain the influence of using a previously refrigerated and reactivated sample (a stored sample) as inoculum to set-up several biofilters. Likewise, the relevance of certain simple parameters for achieving biomass activity will be studied. The pollutants to be treated in the biofilters were toluene, ethylbenzene and p-xylene.

2. MATERIALS AND METHODS 2.1 SOURCE OF THE MICROORGANISMS AND MEDIA COMPOSITION Three sludge samples were collected in a wastewater treatment plant (W), in a small river close to a petrochemical company (P) and near a synthetic resin-producing industry (F) in Bizkaia (Spain). After sedimentation of the solid phase of each sample for 2 hours, the liquid phase was transferred into glass bottles for further experimentation. A nutrient medium (Barona et al., 2007) was used for enriching and maintaining the cultures.

BIOFILTER RESPONSE TO BIOMASS REACTIVATION FOR VOC TREATMENT

553

2.2 BIOMASS STORAGE AND ACTIVATION About 1 L of the supernatant phase of each sample was mixed with 1 L of the nutrient solution, and the total volume was transferred into 3 litre vessels that were continuously fed with toluene (inlet concentration ranging from 50 to 100 ppmv for 1000 hours at a gas flow rate of 2.5 L min-1). When this procedure concluded, the activated samples were stored at 6 oC for 90 days. The samples were subsequently warmed to room temperature (20 oC) and the reactivation procedure was then repeated for a further 1000 hours. 2.3 BIOMASS RECOVERY IN BATCH ASSAYS Two different parameters were used to compare biomass recovery before and after refrigeration and storage. The first parameter was the ratio between volatile solid content and total solid content (VS/TS), and it was basically used to select the best sample for subsequent experiments. After the preliminary selection of one of the samples based on VS/TS data, the second step was to determine degradation rates in order to ascertain the influence of refrigeration and storage lag on the selected sample activity. The originally activated sample was transferred into a hermetic vessel and 2 μL of toluene were added. The removal of the contaminant was monitored over time until the complete disappearance of the pollutant. Likewise, two consecutive doses of 4 and 8 μL of toluene were also added to determine the degradation rate of the biomass prior to inoculation. The same procedure was repeated with the reactivated sample (after refrigeration), and in this case, toluene, ethylbenzene and xylene were individually added in the respective hermetic vessels. After the addition of 2, 4 and 8 μL of each pollutant, the degradation rate was monitored over time until the complete disappearance of the compound. 2.4 BIOFILTER SETUP The packing material used for filling four biofilters was made up of composted pig manure and sawdust and has already been used in previous work (Elías et al., 2002). The outline of the pilot plant for each of the four biofilters has already been described in detail (Moura et al., 2006). The first biofilter (biofilter 0T) was inoculated with the original activated sample (without refrigeration and for toluene degradation) and the other three were inoculated with the refrigerated/reactivated sample, although the contaminant to be degraded was toluene, ethylbenzene and p-xylene (biofilter T, biofilter E and biofilter X, respectively). The inlet loading rate for the three contaminants ranged from 2.6 to 29.1 g m-3 -1 h . Two different gas flow rates were used during experimentation; the first one was 1 L min-1 and the second one was 2 L min-1, corresponding to a residence time of 180 and 90 s, respectively. Temperature was constant at 23 oC.

554


2.5 ANALYTICAL METHODS Toluene, ethylbenzene and p-xylene were measured in a micro gas chromatograph (microGC CP 4900) equipped with auto-sampling injection mode and a TCD detector. Operating conditions were: injector temperature, 110 oC; column/ transfer temperature 80 oC.

3. RESULTS AND DISCUSSION The original three samples were activated for 1000 hours, and after a 90-day refrigeration and storage period, they were reactivated for a further 1000 hours. The evolution of the biomass during these activation assays was monitored by measuring the changes in the ratio between the volatile solid content and total solid content (VS/ TS). The results, shown in Table 1, revealed that none of the three samples was able to fully recover initial values of the controlling parameter after tempering. In all cases, the VS/ST ratio decreased by about 20% after storage. Nevertheless, the sample from the wastewater facility (W) showed the highest VS/TS ratio before (85) and after refrigeration (65). Consequently, W sample was preliminary selected as the best inoculum for subsequent operation, although further analyses were carried out to ascertain the degradation rate before and after storage. Table 1. The VS/TS ratio for the three activated samples before and after refrigeration.

Original After 1000 hours of activation After refrigeration and 1000 hours of reactivation

W sample 28 85 65

100 x VS/TS P sample F sample 21 20 60 72 48 58

The degradation rate before and after refrigeration for the toluene activated sample is shown in Table 2. The originally activated W sample achieved a relatively constant degradation rate for the addition of 4 μL of toluene (211 ppmv h-1) and for 8 μL (222 ppmv h-1). Surprisingly, the degradation rate was higher after refrigeration and very similar for an addition of 2 and 8 μL (291 and 280 ppmv h-1, respectively). Consequently, the decrease in the VS/TS ratio after refrigeration for W sample was not considered to be a decisive result (Table 1). Thus, the ratio reached after storage was high enough for the biomass to recover activity.


555

Table 2. Degradation rate for the originally activated and refrigerated/reactivated W sample.

Sample type Originally activated Refrigerated/reactivated Refrigerated/reactivated Refrigerated/reactivated

Compound Toluene Toluene Ethylbenzene p-Xylene

Degradation rate (ppmv h-1) Dose 2 μL 4μL 8μL 66 211 222 291 213 280 263 258 526 194 233 204

The degradation rate for ethylbenzene and p-xylene is also shown in Table 2. The results for p-xylene revealed that the biomass in the refrigerated sample was able to degrade this pollutant at a rate ranging from 194 to 233 ppmv h-1, in spite of the contaminant dose. The refrigerated sample also showed a high capacity to degrade ethylbenzene, above all for the highest dose. The degradation rate was fairly constant for the addition of 2 and 4 μL, but it doubled when 8 μL of ethylbenzene were added. As an example, the remaining amount of ethylbenzene after the addition of 4 μL is plotted in Figure 1. Those preliminary results suggested that the refrigeration of W sample did not stop the biomass degrading the three simple alkylbenzenes. Furthermore, the highest degradation rate was reached for ethylbenzene with an addition dose of 8 μL.

Figure 1. Evolution of the remaining concentration of ethylbenzene along time after an addition of 4 μL (refrigerated/reactivated sample).

556


In order to ascertain the behaviour of W sample (originally activated and refrigerated/reactivated) as inoculum for biofiltration purposes, four biofilters were started-up. Two of them were fed with toluene (biofilters 0T and T) and the other two (biofilters E and X) were fed with ethylbenzene and p-xylene, respectively. The inoculation of biofilter 0T was carried out with the originally activated W sample and the inoculation of the others involved the refrigerated and reactivated W sample. In all cases, two flow rates of 1 and 2 L min-1 were tested. The comparison of the response of biofilters 0T and T rendered similar results and, consequently, only the data for biofilter T have been plotted in Figure 2. It was concluded that the active biomass degrading toluene was not affected by the refrigeration, in spite of the fact that the VS/TS ratio reduced after the storage and reactivation period.

Figure 2. Response of Biofilter T treating toluene (inoculation with the refrigerated/reactivated sample).

As shown in Figure 2, when the inlet loading rate (IL) ranged from 2.6 to 26.2 g toluene m-3 h-1 for a residence time of 180 s (gas flow rate of 1 L min-1), the removal efficiency was close to 100%. When the residence time was reduced by half (80 s) as a consequence of increasing the gas flow rate to 2 L min-1, inlet load higher than 18 g m-3 h-1 rendered lower results for elimination capacity (data plotted below the dashed line). When the gas flow rate was 1 L min-1 and the inlet load of ethylbenzene fed into biofilter E ranged from 4.8 to 28.5 g m-3 h-1, the biomass was able to completely


557

degrade the contaminant. In contrast, the change to the higher gas flow rate dramatically reduced elimination capacity and biofilter response was very variable. It is remarkable that the latter result is not consistent with the data shown in Table 2, where ethylbenzene recorded the highest degradation rate. Nevertheless, this behaviour may be explained by the random attachment of biomass to the support material, which is a phenomenon for further research.

Figure 3. Response of Biofilter E treating ethylbenzene (inoculation with the refrigerated/reactivated sample).

The results obtained for the p-xylene biofilter were similar to those obtained for biofilter E, and, for brevity, are not shown.

4 CONCLUSIONS The storage of active biomass under refrigeration (not in freezing conditions) is necessary for reproducing biofilters. In this study, the influence of a 90-day refrigeration lag was studied. Initially, the volatile solid content/total solid content (VS/TS) ratio of three activated sludge samples was measured before and after refrigeration. None of the three samples was able to fully recover initial values of the controlling parameter after tempering. Nevertheless, the sample with the largest VS/ST ratio was selected for determining the degradation rate of toluene before and after refrigeration.

558


Surprisingly, the degradation rate for toluene was higher after refrigeration. Subsequently, one biofilter was inoculated with the activated sample and another one with the refrigerated/reactivated sample, with both of them being fed with a toluene loading rate ranging from 2.6 to 26.2 g m-3 h-1. It was concluded that the active biomass degrading toluene was not affected by the refrigeration, in spite of the fact the SV/ST ratio decreased after the storage and reactivation period. The elimination capacity of the biofilters for ethylbenzene and p-xylene was highly influenced by the gas flow rate, which is possibly related to the uneven attachment of biomass to the support material in these cases.

5 ACKNOWLEDGEMENTS The authors gratefully acknowledge the financial support of the University of the Basque Country (Research group GIU05/12), and the Spanish Ministry of Science and Technology (Project CTM2006-02460 with ERDF funding).

REFERENCES Alvarez-Hornos, F.J., Gabaldón, C., Martínez-Soria, V., Marzal, P. and Penya-Roja, J.M. (2005) Biodegradation of ethyl acetate and toluene mixtures by a peat biofilter: Respirometry monitoring and dynamics of living and dead bacterial cells. Proceedings of the International Congress on Biotechniques for Air Pollution Control, La Coruña, Spain, pp. 413-420. Barona, A., Elías, A., Arias, R., Cano, I. and González, R. (2004) Biofilter response to gradual and sudden variations in operating conditions. Biochem. Eng. J. 22: 25-31. Barona, A., Elías, A., Cano, I., Uriarte, A. and Artetxe, J. (2007) Additional determinations in a biofiltration system for toluene: Adsorption and partition in the nutrient solution. Chem. and Biochem. Eng. Q. CABEQ 763 (in press). Christen, P., Domenech, F., Michelena, G., Auria, R. and Revah, S. (2002) Biofiltration of volatile ethanol using sugar cane bagasse inoculated with Candida utilis. J. Hazard. Mat. B89: 253-265. Elías, A., Barona, A., Arreguy, A., Ríos, J., Aranguiz, I. and Peñas, J. (2002) Evaluation of a packing material for the biodegradation of H2S and product analysis. Proc. Biochem. 37(8): 812-820. Estévez, E., Veiga, M.C. and Kennes, C. (2004) Fungal biodegradation of toluene in gas-phase biofilters. In: Verstraete, W. (Ed.), Proceedings of the 5 th European Symposium on Environmental Biotechnology. Oostende, Belgium, 25-28 April, pp. 337-340.


559

Fontvieille, D. and Moul, A. (1985) Dénombrement et mesure d‘activité, Bactériologie des milieux aquatiques , Point sur l‘épuration et le traitement des effluents (eau, air): Lavoisier Ed.: 79-131. González-Sánchez, A. and Revah, S. (2007) The effect of chemical oxidation on the biological sulphide oxidation by an alkaliphilic sulfoxidizing bacterial consortium. Enz. Microb. Technol. 40: 292-298. Hwang, S.C.J., Lee, C.M., Lee, H.C. and Fang Pua, H. (2003) Biofiltration of waste gases containing both ethyl acetate and toluene using different combinations of bacterial cultures. J. Biotechnol. 105: 83-94. Kim, H.S. and Jaffé, P.R. (2007) Spatial distribution and physiological state of bacteria in a sand column experiment during the biodegradation of toluene. Wat. Res. 41: 2089-2100. Laurent, P. and Servais, P. (1995) Fixed bacterial biomass estimated by potential exoproteolytic activity. Can. J. Microbiol. 41: 749-752. Moura, M. (2006) Sistemática de crecimiento y seguimiento de biomasa para sistemas de biofiltración. Project 2006. School of Engineering, University of the Basque Country. Prado, O.J., Veiga, M.C. and Kennes, C. (2005) Treatment of gas-phase methanol in conventional biofilters packed with lava rock. Wat. Res. 39: 2385-2393.


561

A comparative study of the characteristics and physical behaviour of different packing materials commonly used in biofiltration ANTONI D. DORADO, XAVIER GAMISANS, DAVID GABRIEL1 AND JAVIER LAFUENTE1 Department of Mining Engineering and Natural Resources, Universitat Politècnica de Catalunya, Manresa, Spain 1 Department of Chemical Engineering, ETSE, Universitat Autònoma de Barcelona, Barcelona, Spain

ABSTRACT In this study, the characteristics and physical behaviour of 8 different packing materials were compared. The materials were selected according to previous works in the field of biofiltration including organic and inorganic or synthetic materials. Results pre-selected those more acceptable support materials for the main function they have to perform in the biological system: high surface contact, rugosity to immobilize the biomass, low pressure drop, nutrients supply, water retentivity or a commitment among them. Otherwise, pressure drop have been described by means of the respective mathematic expressions in order to include phenomena in the classical biofiltration models.

1 INTRODUCTION Biological treatment have become and effective and economical alternative to the traditional systems of gas treatment. However, several packing materials have been used in biofiltration without a global agreement about which one is the most adequate to immobilize biomass. Carrier materials may be organic, natural inorganic, or entirely synthetic. The nature of the packing material is a fundamental factor for successful application of biofilters because it affects the frequency at which the medium is replaced and other key factors such as bacterial activity and pressure drop across the bioreactor (Devinny et al., 1999).

561

562

ANTONI D. DORADO, XAVIER GAMISANS, DAVID GABRIEL AND JAVIER LAFUENTE

Moreover, particles vary in size, which affects important medium characteristics such as resistance to air flow and effective biofilm surface area. If the size of the bed pellets is too small provides for large specific surface areas, available for essential mass exchange, but it also creates resistance to gas flow, while if it is too large, it favours gaseous flows but reduces the number of potential sites for the microbial activity (Delhoménie et al., 2002). Adu and Otten (1996) have reported that particle size is a parameter more influential to the performance than the gas flow rate. Among the naturals carriers reported, compost, peat, soil and the wood derivatives are the most extensively used while GAC, perlite, glass beads, ceramic rings, polyurethane foam, polystyrene and vermiculite are some of the several synthetic or inert carriers which have been studied (Kennes and Veiga, 2001). Specific surface area, porosity, density, water retention capacity and the nutrients availability are some of the most important characteristics of the filter media (Janni et al., 2001). In this work, a comprehensive study of physical parameters for different packing materials commonly used in biofiltration has been performed. Pressure drop was also determined for each packing material to determinate the inherent economical cost to flow the air through the bed. To this aim, pressure drop was evaluated in each case depending on the flow rate, the bed porosity and the water content circulating through the material media in countercurrent flow. Pressure drops have been described by means of mathematical expressions relating the effects of the studied factors in order to include this parameter in classical biofiltration models.

2 MATERIALS AND METHODS 2.1 EXPERIMENTAL SETUP Pressure drop assessment experiments were carried out using a lab-scale plant consisting of a PVC column with an inner diameter of 4.6 cm and a height of 70 cm (Figure 1). The compressed air was conducted by 2 different circuits. In the former, the air stream was passed through a water column in order to increase the relative humidity and in the latter, the air stream arrived completely dry to the fixed bed. The inlet air pressure and the flow rate were controlled and measured by means of a pressure regulator (Norgren Excelon) and a flowmeter (Tecfluid 2100) respectively. Throughout this study, the gaseous stream was supplied in up-flow mode. Tap water was sprinkled continuously at the top of the fixed bed be means of a peristaltic pump (Magdos LT10) and the water content was measured by an optical level sensor. Pressure drop was determined by means of two digital differential pressure meter used according to the limit detection and precision (Testo 512-20hPa and Testo 506-200 hPa).

A COMPARATIVE STUDY OF THE CHARACTERISTICS AND PHYSICAL BEHAVIOUR

563

Figure 1. Schematic of the lab-scale setup. 1: humidification column; 2: fixed bed for pressure drop study; 3: fixed bed for water retentivity study; 4: flow meters; 5: pressure regulator; 6: peristaltic pump; 7: digital differential pressure meter.

2.2 PACKING MATERIALS A total of 8 common packing materials used in biofiltration were studied and compared by determining their main physico-schemical properties. Organic packing materials analysed were coconut fibre, pine leaves, peat and compost from sludge of a waste-water treatment plant. The inorganic or synthetic packing materials studied were polyurethane foam, lignite from Mequinenza mines (Spain), lava rock and an advanced material based on a thin layer of compost over a clay pellet. 2.3 ANALYTICAL METHODS Characterization of packing materials were carried out according to standard methods (APHA, 1980; ASTM, 1990; TMECC, 2002). The following properties were compared in each case: specific surface area, elementary analysis, extractable phosphor content, organic matter, humidity, water holding capacity, retentivity, ph, conductivity and buffer capacity of the leachate. Specific surface area and material density were determined by the BET technique in a Micromeritics, model Tristar 3000, apparatus. Elementary analysis was realised by combustion in standard conditions using sulfanilamida as standard (EA-1108 ThermoFisher Scientific). Extractable phosphor was determined by the technique of ICP in a multichannel analyser in standard

564


conditions (Thermo Jarell-Ash model 61E Polyscan) using Baker Instra as digester of the sample. Surface rugosity was observed by means of a Scanning Electron Microscope (Jeol JSM-840). Humidity and organic matter were determined by drying and combustion standard procedures. Water holding capacity was measured keeping the material wet sparkling constantly tap water for 100 minutes and determining the weight changes. Water retentivity was measured by keeping wet material in constant contact with dry air flow circulating through the bed and measuring the loss of weight of the bed. Conductivity, pH and buffer capacity was determined for the materials leachate submerging them in water for 1 hour in controlled conditions of temperature and agitation.

3 RESULTS AND DISCUSSION 3.1 CHARACTERIZATION OF PACKING MATERIALS High nutrient, phosphorous, potassium and sulphate contents, as well as trace elements, are required for the establishment of a dense process culture. Regarding to the elementary composition of organic packing materials (Table 1), it is shown that the compost is the material with the highest content in nitrogen and phosphorus (2.7 % and 14.500 ppm, respectively. It must be pointed out that immature coal (lignite) studied showed a significant concentration of sulphur (8,8 %) which is related to the quality of the material. Also, presence of sulphur has been detected in compost as well. On the other hand, phosphorous concentration in lava rock (1800 ppm) is higher than expected probably due to the pre-treatment of this material to garden applications. Among the organic material analysed, coconut fibre and pine leaves present the highest organic matter content (higher than 85% by weight). The organic matter detected in coal (next to 80%) is a reflex of the immature nature of this material. In general, it is desirable to have media with a high water-holding capacity. Organic media are 40 to 80% water (by weight) when they are saturated (Devinny et al., 2002). Packing materials studied keep a water holding capacity inside the typical interval, being in the higher values for coconut fibre, pine leaves and peat (Table 2). The humidity of the materials is similar in all the studied cases but there are appreciable differences in water retentivity. Regarding to the specific surface, coal is the material with the highest value (6 m2 m-3), while compost is the highest among the organic materials (2,8 m2 m-3).

565


Table 1. Elementary composition of packing materials.

Coconut fiber Pine leaves Peat Compost Advanced magterial Lava rock Coal

Nitrogen (%) 1,17 0,56 1,26 2,68 0,34 0,00 0,85

Carbon (%) 45,05 45,18 21,99 33,86 2,45 0,40 44,37

Hydrogen (%) 6,18 6,10 2,56 4,63 0,18 0,00 4,06

Sulphur (%) 0,12 0,05 0,15 0,63 0,19 0,00 8,81

Phosphorus (ppm) 256 191 455 14487 1259 1821 98

Organic matter (%) 91,62 86,71 66,23 53,56 2,57 0,63 79,69

Table 2. Physical characteristics of packing materials.

Coconut fiber Pine leaves Peat Compost Advanced material Lava rock Coal Polyeurethane foam

Surface area (m2 m-3)

Humidity (%)

Water holding (g·g-1)

1,68 0,50 1,43 2,82 0,76 0,62 5,99 0,02

6,62 7,79 6,97 7,83 37,62 0,06 4,85 –

3,90 1,51 1,80 0,68 0,58 0,18 0,28 –

Water Conductivity pH retentivity (μS) (% dia-1) 192,24 422,78 66,38 57,89 41,90 23,33 41,62 416,45

315 216 338 470 226 33 205 –

5,93 6,90 5,13 7,24 5,72 7,21 6,51 –

Buffer capacity (ml So42-·l-1) 33 120 20 128 13 33 45 –

Packing materials studied showed a pH close to the neutrality or slightly acid (pH ~ 5 for peat) and a buffer capacity inferior to 150 ml SO42- l-1 in all the cases. Leachate conductivity of the materials was similar among them (excepting lava rock), being 470 μS the highest value determined in compost. The surface rugosity of the materials has been observed and compared by means of Scanning Electron Microscopy. As a sample of organic materials, coconut fibre shows an important surface rugosity which could aim to fix the microorganisms to the surface (Figure 2). Conversely, polyurethane foam shows the opposite situation where the surface observation at 1000 magnifications shows a completely flat surface.

566

a)


b)

Figure 2. Microscopic observation of the rugosity of a) coconut fibre at different magnifications by SEM (x30, x1000) and b) polyurethane foam (x30, x1000).

3.2 PARAMETERS INFLUENCE IN PRESSURE DROP Pressure drop tests were carried out at 7 different flow rates, 5 different water contents and 3 different bed porosities. Flow rates were selected in the range to obtain empty bed residence times commonly used in biofiltration (from 5 to 40 seconds). Water content circulating through the bed was regulated by means of the peristaltic pump avoiding flooding episodes. Porosity was selected through different particle size or different degrees of compactation depending on the materials as for instance, coconut fibre or pine leaves. Results were represented in surface plots to observe simultaneously the parameters influence pressure drop. Figure 3 shows the effect for coconut fibre and compost as examples of organic packing materials behaviour. Regarding water content, the influence is very similar for both materials in opposition to empty bed porosity influence where results do not show significant differences. Figure 4 shows an example of the behaviour of non-organic materials, concretely polyurethane foam and the advanced material. Polyurethane foam showed more important pressure drop in the range of study, presenting significantly differences for the different bed porosities tested. Water content is a parameter less influencing in comparison to organic samples. Drop pressure in advanced material, in the only possible porosity allowed by its shape and structure, shows a strong dependence on water content, being more important at high flow rates.


Figure 3. Influence of operational parameters in drop pressure for coconut fibre (a) and compost (b).

567

568


Figure 4. Influence of operational parameters in drop pressure for polyurethane foam (a) and advanced material (b).


569

3.3 MATHEMATICAL EXPRESSIONS IN DROP PRESSURE Drop pressure in a fixed bed has been described through several semiempirical mathematical expressions. In most of works, the pressure drop is described by the well-known Ergun equation (Ergun, 1952), which may be written as (Eq. 1):

(1)

Where: ΔP is the pressure drop in Pa; H is the height of the fixed bed in m; μ is the viscosity of the air in Pa s; ν0 is the superficial velocity in m s-1; ε is the porosity of the bed; dp is the equivalent spherical diameter of the particle in m; a is the first parameter of the Ergun equation, b is the second parameter of the Ergun equation. Parameters a and b are related to the friction factor. In addition the expression term related to parameter a is significant for flow under very viscous conditions while the parameter b term is only significant when viscous effects are not as important as inertia. Some authors have fitted satisfactorily experimental data to a modified Ergun equation adapting the coefficients of the expression be means of a correction factor (Delhoménie et al., 2003). Other authors have used a specific relation due to the heterogeneity of the material and the difficulty to model pressure drop with the classical Ergun equation (Comiti and Renaud, 1989). In this study, parameters a and b from Ergun equation have been fitted as a function of the material, the porosity and the water content. Pressure drop (ΔP/H) as a function of the empty bed velocity has been fitted by a lineal regression. In all cases the correlation coefficient R2 was superior to 0,990 indicating the correct linearity between operational parameters and pressure drop. This experimental study incorporated the effect of water content in pressure drop although this parameter is not present in Ergun equation. For this reason, parameter a and b were fitted as a function of water content in the bed in order to find a relationship which describes this effect in the pressure drop estimation. Table 3 shows the final results of this systematic study in order to compare the water content effect in each material. Water content for compost, lava rock and the advanced material biofilter showed the strongest effect in parameter a. In the case of parameter b the dependence on water content was markedly lower. Thus, it is possible to express a modified Ergun equation incorporating the water content effect in the pressure drop predictions for some packing materials. These results may be useful to incorporate pressure drop phenomena in classical biofilter models.

570


Table 3. Ergun equation parameters as a function of water content in biofilters.

Compost

Coconut fibre

Lava rock

Immature coal

Pine leaves

Advanced material

0,70 0,76 0,79 0,94 0,96 0,99 0,73 0,76 0,77 0,58 0,63 0,64 0,91 0,92 0,96 0,65

a ordinate 12,634 132,370 432,090 0,626 2,145 9,130 75,398 115,150 234,900 14,618 35,182 100,830 1,885 2,524 9,378 12,342

a slope -0,069 0,720 4,077 0,004 0,019 0,750 0,340 0,384 2,276 0,212 0,387 0,270 0,014 0,016 0,059 1,458

b ordinate 0,250 0,595 1,199 0,062 0,090 0,333 0,512 0,531 0,766 0,176 0,248 0,466 1,146 2,517 11,196 0,160

b slope 0,003 0,009 0,004 0,001 0,000 -0,011 0,005 0,001 0,002 0,001 0,002 0,007 0,058 0,144 0,000 0,009

4 CONCLUSIONS Commonly used packing materials in biofiltration have been characterized and compared for a better knowledge of their advantages and drawbacks. Coconut fibre, pine leaves, peat, compost, polyurethane foam, coal, lava rock and an advanced material have been studied. Organic materials, especially compost and coconut fibre, are suitable to release extra inorganic nutrients. Moreover, these materials are able to keep water content at optimal levels for microorganisms and show the highest specific surface. Surface observation by Scanning Electron Microscope shows a better condition to fix the biomass in organic materials. On the contrary, inorganic or synthetic materials offers higher contact surface and produce cleaner drainage water. Otherwise, pressure drop have been determined for each packing materials as a function of flow rate, water content and bed porosity in order to represent the several effects simultaneously and obtain a mathematical expression to include phenomena in classical biofilter models. A water content dependence has been found through a modified Ergun equation for several packing materials.


571

5 ACKNOWLEDGEMENTS This work was supported by the Spanish CICYT project CTQ2006 – 14997-C02-02. A.D. Dorado received a predoctoral fellowship from the M.E.C. (Ministerio de Educación y Ciencia).

REFERENCES Adu, B. and Otten, L. (1996) Modelling the biofiltration characteristics of volatile compounds. Proceedings of the 89th annual meeting and exhibition of the air & waste management association, June 23-26, 1996, Nashville, TN. Pittsburgh, PA: Air & Waste Management Association. APHA (American Public Health Association). (1980) Standard methods for the examination of water and wastewater. ASTM (Standard Test Method). (1990) Standard methods for volume weights, water holding, and air capacity of water saturated peat materials. In (annual book of ASTM Standards) Vol. 04.08. Bohn, H.L. (1996) Biofilter media. Proceedings of the 89th annual meeting and exhibition of the air & waste management association, June 23-26, 1996, Nashville, TN. Pittsburgh, PA: Air & Waste Management Association. Bohn, H.L. and Bohn, K.H. (1999) Moisture in biofilters. Environ. Prog. 18: 156-161. Comiti, J. and Renaud, M. (1989) A new model for determining mean structure parameters of fixed beds from pressure drop measurements: application to beds packed with parallelepipedal particles. Chem. Eng. Sci. 44: 1539-1545. Delhomenie, M.C., Bibeau, L., Gendron, J., Brzezinski, R. and Heitz, M. (2003) A study of clogging in a biofilter treating toluene vapors. Chem. Eng. J. 94: 211-222. Delhomenie, M.C., Bibeau, L. and Heitz, M. (2002) A study of the impact of particle size and adsorption phenomena in a compost-based biological filter. Chem. Eng. Sci. 57: 49995010. Deront, M., Samb, F.M., Adler, N. and Péringer, P. (1998) Biomass growth monitoring using pressure drop in a cocurrent biofilter. Biotechnol. Bioeng. 60: 97-104. Devinny, J. S., Deshusses, M.A. and Webster, T. S. (1999) Biofiltration for air pollution control. Lewis publishers, Boca Raton, Florida. Durham, D.R., Marshall, L.C., Miller, J.G. and Chmurny, A.B. (1994) Characteritzation of inorganic biocarriers that moderate system upsets during fixed-film biotreatment processes,.Appl. Environ. Microbiol. 60(9): 3329. Ergun, S. (1952) Fluid flor thorugh packed columns. Chem. Eng. Prog. 48: 9-94.

572


Janni, K.A., Maier, W.J., Kuehn, T. H., Yang, C.-H., Bridges, B.B., Velsey, D. and Nellis, M. A. (2001) Evaluation of biofiltration of air-An innovative air pollution control technology. ASHRAE Transactions 107: 198-5214. Kennes, C. and Veiga, M.C. (2001) Bioreactors for waste gas treatment. Kluwer academic publishers, Dordrecht, The Netherlands. Morgan-Sagastume, F., Sleep, B.E. and Allen, D.G. (2001) Effects of biomass growth on gas pressure drop in biofilters. J. Environ. Eng. 127: 388-396. Morgan-Sagastume, J.M. and Noyola, A. (2006) Hydrogen sulfide removal by compost biofiltration: Effect of mixing the filter media on operational factors. Biores. Technol. 97: 1546-1553. TMECC (Test Methods for the Examination of Composting and Compost). (2002).

573

Suitability of dust and bioaerosols from a pig stable as inoculum for biological air filters ANJA KRISTIANSEN, PER HALKJÆR NIELSEN, AND JEPPE LUND NIELSEN Department of Biotechnology, Chemistry and Environmental Engineering, Aalborg University, Sohngaardsholmsvej 57, DK-9000 Aalborg

ABSTRACT Biofiltration for removing ammonia and odour compounds from ventilation air of pig stables is a promising approach. In order to reduce the time for starting up a well-functioning biofilter a good inoculum suited for the environment is needed. In this study the microbial identity and quantity of dust and bioaerosols coming from a pig stable were analyzed for its suitability as inoculum. Inoculation of biofilters with dust had similar good ammonia removal capabilities as biofilters inoculated with activated sludge, although analysis of the microbial dust community revealed clear differences. The organic fraction of the dust particles seems to be important for mediating biofilm development on the filter material.

1 INTRODUCTION Increasing productivity combined with a more centralized production of pigs has the last decades created odour nuisance problems and a high load on the environment through air emissions. Biofiltration is a promising and low cost technology for treatment of large volumes of air containing low concentrations of different compounds, such as ventilation air from a pig stable. The quantitatively most important components in ventilation air from pig stables are organic acids and ammonia, which can readily be oxidized by heterotrophic and nitrifying bacteria, respectively. A well-functioning biological air filter requires microorganisms capable of utilizing air components and adapted to live under extreme conditions. Inoculations by addition of activated sludge or material from existing filters often speed up the process of establishing a new biofilm. However, activated sludge might not be optimal

573

574

ANJA KRISTIANSEN, PER HALKJÆR NIELSEN

AND

JEPPE LUND NIELSEN

as the microorganisms herein are not adapted to the conditions on the new filter material nor the substrate fed into the biofilter. Microorganisms indigenous to the ventilation air, aerosols, and dust particles may already be pre-adapted to life in the environment and thus better suited as inoculum. However, only few investigations, most of which applying culture-dependent techniques, have dealt with the composition of airborne dust and aerosols from pig stables. By analyzing the microbial community herein its potential use as inoculum of nitrifying bacteria and heterotrophic microorganisms into biofilters can be evaluated. In the present study we have applied cultivation-independent molecular techniques to characterize the microbial community composition in dust and bioaerosols from a pig stable, and compared this to established biofilters treating air from the same stable.

2 MATERIALS AND METHODS 2.1 SAMPLING Dust and bioaerosols was obtained from a closed gestation unit with 450 sows near Aalborg, Denmark. Dust and bioaerosol samples were collected from the air by introducing a 2 cm thick (30 cm in diameter) water resistant cellulose filter with a pore size of

Biotechniques for Air Pollution Control : proceedings of the II International Congress on Biotechniques for Air Pollution Control, A Coruña, Spain, October 3-5, 2007

Costs of Air Pollution Control

Air Pollution Control Technology Handbook

Air pollution control equipment calculations

Monitoring, Control and Effects of Air Pollution

Hemispheric Transport of Air Pollution 2007 (Air Pollution Studies)

Handbook of Air Pollution Prevention and Control

Alternative policies for the control of air pollution in Poland

Air Pollution and Health (Air Pollution Reviews)

Air Pollution

Air Pollution Control Equipment Selection Guide

Advanced Air and Noise Pollution Control:

Air pollution control equipment selection guide

Fundamentals of Air Pollution

Fundamentals of Air Pollution

Fundamentals of air pollution

Fundamentals of Air Pollution

Modeling Indoor Air Pollution

Air Pollution - New Developments

AIR POLLUTION & HEALTH

Air Pollution Science for the 21st Century

Indoor Air Pollution

Air Pollution and Health

AIR POLLUTION & HEALTH

Air Pollution XV

Fundamentals of Air Pollution Engineering

HANDBOOK OF ENVIRONMENTAL ENGINEERING - Air Pollution Control Engineering

Handbook of Air Pollution Control Systems and Devices

Handbook of Air Pollution Prevention and Control, First Edition

industrial water pollution control

Environmental Pollution Control Microbiology

Biotechniques for Air Pollution Control : proceedings of the II International Congress on Biotechniques for Air Pollution Control, A Coruña, Spain, October 3-5, 2007

Costs of Air Pollution Control

Air Pollution Control Technology Handbook

Air pollution control equipment calculations

Monitoring, Control and Effects of Air Pollution

Hemispheric Transport of Air Pollution 2007 (Air Pollution Studies)

Handbook of Air Pollution Prevention and Control

Alternative policies for the control of air pollution in Poland

Air Pollution and Health (Air Pollution Reviews)

Air Pollution

Air Pollution Control Equipment Selection Guide

Advanced Air and Noise Pollution Control:

Air pollution control equipment selection guide

Fundamentals of Air Pollution

Fundamentals of Air Pollution

Fundamentals of air pollution

Fundamentals of Air Pollution

Modeling Indoor Air Pollution

Air Pollution - New Developments

AIR POLLUTION & HEALTH

Air Pollution Science for the 21st Century

Indoor Air Pollution

Air Pollution and Health

AIR POLLUTION & HEALTH

Air Pollution XV

Fundamentals of Air Pollution Engineering

HANDBOOK OF ENVIRONMENTAL ENGINEERING - Air Pollution Control Engineering

Handbook of Air Pollution Control Systems and Devices

Handbook of Air Pollution Prevention and Control, First Edition

industrial water pollution control

Environmental Pollution Control Microbiology

Recommend Documents