Romain Cresson

Patrick Dabert

Nicolas Bernet
email: bernet@supagro.inra.fr

Microbiology and performance of a methanogenic biofilm reactor during the 1 start-up period 2

Keywords: Biofilm, Anaerobic digestion, Bioreactor start-up, Microbial population 21 dynamics, Methanogenesis, Acetogenesis

Microbiology and performance of a methanogenic

biofilm reactor during the start-up period R. Cresson, P. Dabert, N.

Introduction 1

Anaerobic digestion is an effective method for treating many organic wastes. Compared with 2 aerobic process, the advantages are : lower energy requirement, higher treatment efficiency, 3 lower sludge amount and valuable gas production (Lettinga et al. 1983). The limitation of 4 anaerobic digestion processes resides in the duration and instability of their transitory phases 5 (starting, restarting, organic load increase) because of the anaerobic microorganisms low The natural ability of microorganisms to aggregate between them and/or to adhere on a 9 solid surface and form a biofilm can be exploited to improve significantly the efficiency of 10 anaerobic wastewater treatment processes by increasing their capacity of microbial biomass 11 retention (Hall 1987;[START_REF] Annachhatre | Anaerobic treatment of industrial wastewaters[END_REF]. If the advantages of fixed biomass processes for 12 anaerobic digestion have been underlined (i.e. biomass retention, robustness, high loads…) 13 (Nicolella et al. 2000), the optimization of the start-up period seems necessary to increase the 14 competitiveness of anaerobic high rate processes. This initial phase can be defined as the time 15 required for the selection and organisation of an effective microbial consortium for a given 16 pollutant (Weiland et al. 1991;Heppner et al. 1992). Regarding process performances, the 17 objective of this phase is to reach as quickly as possible the process nominal load with the 18 desired treatment efficiency without creating instability of the anaerobic ecosystem. [START_REF] Amann | Fluorescent-oligonucleotide probing of 20 whole cells for determinative, phylogenetic, and environmental studies in microbiology[END_REF] For Heijnen et al. (1989) however, the organic load increase during start up should not be 20 related to a criterion of treatment performance, but rather with criteria of biological activities.

21

During start-up, a continuous and progressive load increase should occur while preserving an 22 optimal biological activity in order to respect as well as possible biomass nutritional 23 requirements to avoid transitory shocks (Stronach et al. 1986). Hence, the influent organic 24 composition would significantly select the microorganisms involved in the anaerobic biofilm increased biogas production, the loading rate was increased by increasing the wine 17 concentration in the influent as above. In both strategies, the objective was to increase the 18 COD load of the reactor from 0.5 to 20 g COD .L -1 .d -1 as quickly as possible without inhibiting concentration remained below 0.1 g COD .L -1 . As a consequence COD removal rates increased 1 quickly to reach more than 90% of removal efficiency for both systems (Fig. 1).

2

Considering the next 28 days (phase II), the OLR was increased from 0.5 g COD .L -1 .d -1 to 3 6.5 g COD .L -1 .d was kept stable at 6.5 g COD .L -1 .d -1 . The OLR of R B was still slightly increased from 6.5 to 10.7 12 g COD .L -1 .d -1 , while its COD removal efficiency stood above 70%. Chromatographic analysis 13 showed that propionic acid represented 13.5 and 13.4% of total VFA accumulated for R A and 14 R B respectively. This VFA accumulation (intermediates and also inhibitors of anaerobic identical profiles with only two strongly dominant peaks (E and K in Fig. 2). These peaks co-18 migrated with distinguishable but not very dominant peaks from the inoculum profile. After The Sporomusa are acetogenic bacteria able to grow with many substrates such as

6

  growth rate and sensitivity to perturbations such as organic overloads (Henze et al. 1983; 7 Weiland et al. 1991; Austermann-Haun et al. 1994; Puñal et al. 2001; Arnaiz et al. 2003).

  8

  25

(

  precisely the proportions of hydrogenotrophic and acetoclastic 1 methanogens(Ehlinger et al. 1989). However, until recently, it was difficult to assess the 2 microbiology of such processes, which is now more accessible thanks to the development of 3 molecular tools. Thus, FISH and/or molecular fingerprints (DGGE, SSCP) have been recently 4 used to study the dynamic transition of a methanogenic population in a thermophilic digester 5 Hori et al., 2006), the microbial community and population dynamics in an anaerobic hybrid 6 reactor (Boonapatcharoen et al., 2007), or the microbial community during biofilm 7 development in an Upflow Anaerobic Sludge Bed (UASB) reactor (Fernandez et al, 2008). 8 It is clear that controlling the anaerobic processes start-up requires a thorough knowledge 9 of the biological processes and microorganisms which implement them. The objective of the 10 present study was to establish a link between process performances and implementation, 11 evaluated during the start-up period of two anaerobic three phase reactors and the 12 microbiology of their corresponding ecosystems. The installation of biofilm microbial 13 communities was investigated by 16S rRNA targeted PCR and Single Strand Conformation 14 Polymorphism (SSCP) capillary electrophoresis of the PCR products (Zumstein et al. 2000). 15 Moreover, colonized particles were observed at the end of the start-up period by FISH using a 16 confocal laser scanning microscopy to get information on the spatial distribution of the 17 Archae and bacterial communities in the biofilm. for this study were two inverse turbulent bed biofilm reactors. With this 21 process, anaerobic biomass is grown on low density particles that are fluidised by an up-flow 22 current of biogas recirculated from the top of the reactors. Because of an important 23 solubilisation of CO2 in the liquid phase, the biogas was mainly composed of methane 24 . Biogas injection velocity is used to control the height of fluidization and clogging 1 problems. The reactors had an active volume of 7.24 L and were filled with the solid carrier 2 material up to 24% of its active volume. This coke manufacture residue (U-Spheres , 3 provided by Omega Minerals, Hamburg -Deutschland) is a mineral granular material 4 composed mainly of silica and alumina. The solid particles had an average equivalent 5 diameter of 172 µm and a specific density of 696 g.L -1 . The use of such a small particulate 6 carrier increases considerably the area available for biofilm growth. Further details on 7 experimental devices, carrier characteristics and operational parameters are available in a 8 previous study (Cresson et al. 2006). 9 Operating conditions 10 Both reactors were inoculated with 3 litres of a sludge at 13.6 g.L -1 of volatile suspended 11 solids (VSS) gathered from a pilot scale fixed bed anaerobic digester (1m 3 ) treating winery 12 wastewater. Sludge specific activity was measured at 3 successive times in batch mode, in a 6 13 litres reactor, with ethanol as the only substrate. It was equal to 0.11 g COD .g VSS -1 .d -1 for the 14 third essay (COD, chemical oxygen demand). 15 The influent used in this work was a wine-based wastewaterconsisting of diluted red wine 16 (ethanol as the carbon substrate) complemented with nitrogen as NH 4 Cl and phosphorus as 17 NaH 2 PO 4 , at a ratio of COD/N/P equal to 400/7/1 and, from day 58 until the end of the 18 experiment, with a solution of trace elements at concentration of 5 ppm. The dilution factor 19 was adapted to the potential of the microbial consortium. The COD concentration of the 20 influent was increased from 0.5 to 20 g COD .L-1 from the beginning to the end of the start-up 21 period. Influent was stored at 4°C. Composition of the influent and trace element combination 22 are reported in a previous study (Cresson et al. 2006). 23 During the inoculation period, both reactors have been operated in batch mode for 12 24 hours in order to enhance attachment of planktonic bacteria on the solid carrier (Cresson et al. In the meantime, the sludge was recycled from the bottom to the top of the reactor with 1 a peristaltic pump, to ensure a good contact between the microorganisms and the carrier. After 2 this short inoculation period, a continuous inlet flow of diluted wastewater was applied to 3 maintain a constant hydraulic retention time of 24 hours, in order to enhance biofilm 4 formation and to minimize planktonic bacterial growth in the liquid phase by encouraging 5 their washout (Tijhuis et al. 1994; Michaud et al. 2005; Cresson et al. 2008). Thereafter, the 6 only difference between both reactors was the start-up strategy applied to increase the organic 7 loading rate (OLR) and to bring both reactors to their expected organic load and removal 8 efficiency. 9 For strategy A (reactor R A ), the wine concentration in the influent was stepwise increased 10 (from 0.25% to 10%) when the reactor COD removal efficiency reached 80%. The hydraulic 11 retention time (HRT) was kept constant and equal to one day. For the maximum load strategy 12 (reactor R B ), the reactor was frequently submitted to pulse injections of wastewater. The 13 effect of these disturbances on the amount of biogas produced by the digester was analysed. 14 In case of a decreased biogas production (e.g. the disturbance induces an overload of the 15 reactor), the loading rate was decreased (or kept constant) (Steyer et al. 1999). In case of an 16

  the dominant peaks revealed on SSCP profiles was obtained as described 27 in Dabert et al.(2001). Briefly, the 16S rDNA V3 region of bacterial and archaeal 28 again from total genomic DNA as described above but using the 1 RedTaq DNA polymerase (Sigma) and unlabeled primers. The resulting PCR products were 2 purified with the QIAquick Spin kit (Quiagen) before to be cloned into E. coli using the 3 TOPO TA cloning kit according to supplier instructions (Invitrogen). E. coli with inserts of 4 proper size (about 200 bp) were screened by PCR on colony using plasmid-targeted primers 5 (T7 and M13) and 2% agarose gel electrophoresis. Correct amplified fragments were 6 amplified again using the above PCR-SSCP protocols. All inserts producing a peak that co-7 migrated with distinguishable peaks from the bacterial or archaeal SSCP profiles were 8 sequenced to finalize the process of peak identification. 9 DNA sequences were performed by MillGen (Toulouse, France). Sequences were 10 identified by comparison with their closest relative available in databases using Blast from the 11 National Center for Biotechnology Information and the Ribosomal Database Project. DNA 12 sequences have been deposited in the EMBL database under accession numbers and FISH hybridization were performed using standard conditions 16 (Amann et al. 1990) except for colonized particles handling. Colonized particle samples (2 17 ml) were retrieved from each reactor using a plastic disposable syringe. Sample fixation was 18 performed by sucking three volumes of a 4% paraformaldehyde PBS 1X solution in the 19 syringes, rolling them gently and placing them at 4°C during 3 to 5 hours. The colonized 20 particles were then rinsed 3 times within the syringe by gently sucking in and out PBS 1X 21 buffer. Finally, colonized particles were stored at -20°C in a 1:1 PBS 1X and absolute ethanol 22 solution. For hybridization, fixed particles were rinsed once with hybridization buffer and 23 then hybridization was carried out in 0.6 mL Eppendorf tubes by soaking the colonized 24 particles in 500 µL of 20% formamide hybridization solution (Amann et al., 1990). Labelled -CY3, or ARC915 -FAM, (Batstone et al. 2004)) were added to a final 1 concentration of 4ng.L -1 . Hybridization was done under gentle agitation at 46°C for 3 hours in 2 the dark. After hybridization, particles were washed twice at 48°C in 2 mL washing solution 3 preheated and under gentle agitation (once for 15 minutes and then for 30 minutes). Total 4 cells labelling was carried out by addition of the DNA specific dye chromomycine to washing 5 solutions. Observations were carried out using a laser scanning confocal microscope Leica 6 TCS SP2 equipped with three lasers (wavelengths of excitation of 633, 543, 488 nm).7Microscopic fields were digitized to one µm depth using 40X and 20 X objectives. up period of both reactors was investigated for 81 days. A carbon mass balance was 11 drawn up, considering that ethanol was the only organic carbon source in the diluted wine 12 influent and that carbon flux distributed primarily between anabolic reactions (i.e. biomass 13 production), and catabolic reactions (i.e. biogas and VFA production). These hypotheses were 14 verified by comparing theoretical COD values calculated from ethanol concentration in the 15 influent to COD values measured at different influent organic loads. The differences observed 16 were always below 2% (data not shown). Fig.1 and

  15 digestion) indicated an incomplete degradation of the substrate and signed more particularly a 16 failure of methanogenesis, the last phase of anaerobic digestion. This failure was due to a lack 17 of micronutrients, in particular cobalt (Co) and nickel (Ni), which were found in very low 18 concentrations in the diluted red wine used as influent (Cresson et al. 2006). 19 Therefore, from day 58 until the end of experiment (phase IV), a solution of trace 20 elements was added to the influent (5 ppm). After 5 days of complementation, high COD 21 removal efficiencies were restored (more than 90 %) in both reactors, VFA concentrations 22 dropped below 0.2 g COD .L -1 and the part of input carbon converted into methane exceeded 23 53% (Fig. 1 and

19 22

 19 more days, carrier particles colonization was underway as it is shown by the appearance of 20 other dominant peaks in the profiles and the increased area under the peaks.21The evolutions of bacterial populations during both reactors start-up are shown in Fig.3.[START_REF] Annachhatre | Anaerobic treatment of industrial wastewaters[END_REF] Broadly, the dominant species changed much between the first 16 th days of operation, and 23 remained relatively stable thereafter until the end of the experiment. The dominance of peaks was progressively replaced by the one of peaks F and J for R A and peaks G, H and J 1 for R B . 2 Identification of the observed SSCP peaks was obtained by constructing two bacterial 3 16S rDNA V3 region clone libraries from reactor samples retrieved on day 16 for R A and on 4 day 23 for R B . These representative samples have been selected according to the presence and 5 the height of their peaks. Screening of 50 bacterial clones was enough to assign all clearly 6 distinguishable peaks of both bacterial profiles (Fig. 3, peaks A to M). At least two 7 independent clones were sequenced for each assigned peak. The resulting sequences, of about 8 200 bp length, allowed peaks identification at the genus level (table 2). Six out of the 11 9 peaks identified were related with more than 97% similarity to species or environmental 10 clones belonging to the Delta proteobacteria group, as Geobacter, Pelobacter and the sulfate-11 reducing Desulfovibrio genus. Four belonged to the Firmicutes, three of them to the acetogen 12 group Sporomusa. The last one belonged to the Chloroflexi. Peaks L and M were not -targeted SSCP profile obtained from the inoculum archaeal community 16 showed a low diversity with only 7 distinguishable peaks. Evolution of this community in 17 both reactors is shown in Fig. 4. Biofilms from both reactors on day 1 showed almost 18 identical SSCP profiles containing five to six peaks that all co-migrated with peaks present in 19 the inoculum (Fig. 4). Then, on the contrary of bacterial diversity, archaeal diversity remained 20 low during all the experiment with one largely dominant peak. Library construction for peaks 21 identification was done on an archaeal DNA sample obtained from R A on day 54. PCR-SSCP 22 analyses were conducted on 20 archaeal clones for peaks assignation. All clones co-migrated 23 either with the dominant peak O or with the transient peak N. The former was identified as an sampled on day 72, were hybridized with a combination of probes 4 allowing the selective detection of archaeal (probe ARC915 -FAM) and bacterial populations 5 (probe EUB338 -CY3) and observed by confocal scanning laser microscopy (Fig. 5). The 6 pictures showed a relatively homogeneous distribution of archaeal and bacterial populations 7 with a few archaeal clusters disseminated in the totality of the biofilm without preferential , the start up period of two identical anaerobic fluidized bed reactors operated 11 with different strategies has been monitored. The simplified composition of the influent used 12 resulted in the installation of simplified methanogenic ecosystems. Identification of the 13 dominant microbial populations, linked to the present knowledge of their major metabolic 14 activities, can be used to propose a global scheme of the degradation routes involved within 15 each reactor (Fig. 6). 16 Ethanol degradation 17 After one day of operation, the dominant microbial populations detected on carrier particles 18 belonged to the Sporomusa and the Desulfovibrio groups for the bacterial domain and the 19 Methanosaeta and Methanobacterium groups for the Archaea.

  20

14 enrichment in 2biofilm organization 3 FISH 22 The

 in322 Geobacter in R B and the presence of Fe (0.26mg.L -1 in the diluted wine before 15 addition of the complementation solution (Cresson et al. 2006)) in the reactor feeding ( may 16 have brought an alternative metabolic pathway to the acetoclastic activity of the Sporomusa-17 Methanosaeta couple in this reactor. 18 If the hypothesis presented above can explain the evolution of the dominant microbial 19 populations and process parameters of both reactors, the initial event that conducted to the 20 installation of different microbial communities in the reactors remained unreachable. It 21 certainly appended very early in the process since the Pelobacter and Geobacter populations 22 that enriched during the reactor start up were already detectable on the particles at day 1. suggest a slightly better performances for the R B since it removed 2 30% more COD, produced 32% more methane and accumulated 19% less VFA than R A 3 during the whole course of the experiment. This advantage could be linked to the slight 4 difference in dominant bacterial populations present in both reactors. During the period of 5 trace element limitation (from days 30 to 54), a slight reduction of the Pelobacter dominance 6 compared to the one of Sporomusa was observed in R A , while the same phenomenon was 7 unseen for the Geobacter/Sporomusa couple within R B . While methanogenesis was primarily 8 supported in R A by acetate production from Sporomusa and Pelobacter populations, the 9 presence of two Geobacter populations in R B may have bring alternative degradation 10 pathways and provided an advantage to the microbial consortium in term of catabolic 11 opportunities. The largest biodiversity observed in R B may have also helped the community to 12 support the period of trace element limitation, resulting in a positive impact on the 13 performances and stability of the reactor. 14 The link between ecosystem biodiversity and process stability have been proposed in 15 other studies (Fernandez et al. (1999; Zumstein et al. 2000). These authors have shown 16 independently that the apparent functional stability of two anaerobic digesters was hiding a 17 permanent evolution of their microbial populations (as detected by 16S rDNA-targeted 18 ARDRA or SSCP). In another study, the comparison of two ecosystems subjected to substrate 19 loading shocks, as it was the case in this study for R B , suggested that the microbial 20 community appearing the most stable in its structure was associated with the greatest 21 functional instability (Fernandez et al. 1999). Finally, the advantage of microbial diversity 22 was also pointed out by Schmidt and Ahring (1999): a UASB anaerobic digester inoculated 23 with two species of acetate-consuming methanogenic Archaea was more stable and resistant 24 to fluctuations than a second one inoculated with only one species. Our study, like all the 25 , provides microbial ecologists with data that point out the link between 1 microbial diversity and functional stability. Spatial hybridizations performed on colonized particles at the end of the experiment showed a 4 relatively homogeneous distribution of the archaeal and bacterial populations in the biofilm. 5 As suggested by Batstone et al. (2004), the nature of the substrate probably has a dominating 6 impact on the spatial arrangement of anaerobic digester process biofilms. Complex substrates 7 composed of organic macromolecules (such as brewery or sugarcane refinery effluents) and 8 requiring an important and limiting phase of hydrolysis tends to generate layered distribution 9 of the various trophic groups within the biofilm. Schematically, these biofilms are composed 10 of an external fine layer of acidogenic bacteria and of an internal layer composed of obligate 11 syntrophic bacteria and Archaea methanogens. The core of the biofilm is, most of the time, 12 only slightly or even not hybridized indicating an absence of active cells. 13 Ethanol, the main organic substrate in this study, is small, easily biodegradabile, compact 14 and hydrophilic molecule. These properties favour its diffusion within the biofilm. 15 Furthermore, considering the recent particles colonization, mass transfer phenomena were 16 undoubtedly facilitated by the biofilm low thickness. The bacteria involved in ethanol 17 degradation are not limited by the access to the substrate as it is shown by the strong signal 18 observed in FISH hybridization. Viable and active bacteria cells were hybridized in the whole 19 biofilm, even in zones close to the carrier. The space arrangement of the archaeal and 20 bacterial populations within the biofilm thus seems dictated by the nature of the substrate. 21 Conclusion work presented here aims to contribute to the comprehension of the mechanisms of 23 installation and maturation of anaerobic biofilms, in order to better understand their formation 24 start-up phase of bioreactors. This project, which implies a multi-disciplinary 1 approach, consisted into a global apprehension of two systems started with the same initial 2 conditions (same process, same biomass inoculated) but operated under different start-up 3 strategies. Evolution of the reactor performance as well as the microbial biodiversity appeared 4 to be very similar for both systems. Simplification of the bacterial and archaeal ecosystems 5 compared to the inoculum diversity has been observed and could be ascribed to the nature of 6 the feeding influent. Hence, bacterial biodiversity shows an alternation of the dominant 7 populations supposed to ensure similar or identical metabolic functions for anaerobic ethanol 8 degradation. The slight differences between both systems in term of treatment performance 9 could be explained by the dominance of one Pelobacter in R A versus two Geobacter 10 populations in R B . These two Geobacter populations may have brought alternative 11 degradation pathways or better adaptive response to trace element limitation and provided an 12 advantage to the microbial consortium of the R B in term of catabolic opportunities. Hence, the 13 largest biodiversity observed in the R B seems to have a positive impact on the performances 14 and the stability of the reactor. Furthermore the spatial arrangement of the archaeal and 15 bacterial populations within the biofilm observed by CLSM seems also dictated by the nature 16 of the substrate.

  17

  inverse fluidization, The inverse fluidized bed and the inverse 2 turbulent bed reactors. Environ Technol 24, 1431-1443. 3 Austermann-Haun, U., Seyfried, C.F., Zellner, G. and Diekmann, H. (1994) Start-up of 4 anaerobic fixed film reactors, Technical aspects. Water Sci Technol 29(10-11), 297-308. 5 Batstone, D.J., Keller J. and Blackall L.L. (2004) The influence of substrate kinetics on the 6 microbial community structure in granular anaerobic biomass.

3Figure 1 5 Figure 2 Figure 3 Figure 4 Figure 5 22 Figure 6

 152345226 Figure 1 Variation of OLR, COD removal efficiency and Ethanol to VFA conversion (as

  FIG 3

Samples preservation and total genomic DNA extraction

	1
	2	Two millilitres of colonized particles were sampled from the reactors using a disposable
	3	syringe and gently washed 3 times with 10 millilitres of PBS 1X buffer. These washed
	4	colonized particles were ground in 500 µl of Guanidine Thiocyanate 4M-Tris-Cl 0.1M pH 7.5
	5	and 150 µl of N-Lauroyl sarcosine 10% using a mortar. 500 µl samples were immediately
	6	frozen and stored at -20 °C. Total genomic DNA was extracted from frozen pellets using the
	7	QIAAMP DNA STOOL kit (QIAGEN, France) according to supplier instruction. DNA
	8	purity, integrity and concentration were assessed by 0.7% agarose gel electrophoresis and F r o ethidium bromide staining.
		P
		e
		e r
		R e v
		i e
	19	w
	20	the system.
	21	Reactor monitoring
	22	pH, biogas and liquid flow rates, biogas composition, total suspended solids (TSS), volatile
	23	suspended solids (VSS), attached volatile solids (AVS), volatile fatty acids (VFA),
		bicarbonate (HCO 3
	25

-), and chemical oxygen demand (COD) were routinely analyzed according 24 to the methods described in (Cresson et al. 2006). 9 Bacterial

and archaeal 16S rDNA-targeted PCR-SSCP amplifications

	1	GNTACCTTGTTACGAAACTT (reverse, E. coli position 1492) and the Red Taq DNA
	2	polymerase (Sigma). Reaction mix was the same as above except for the Red Taq DNA
	3	polymerase enzyme and buffer. PCR reaction was started by a denaturation step of 2 min. at
	4	94°C followed by 30 cycles of 1 min at 94°C, 1 min at 50°C and 1 min at 72°C. Reaction was
	5	ended by a 10 min elongation at 72°C and cooling at 4°C. PCR products were then diluted
	6	100 times and used for a V3 region-targeted PCR -SSCP using primers W104 and W116
	7	TCCAGGCCCTACGGGG (forward Archaea, E. coli position 333). Reaction mix and PCR
	10 8 9 10	conditions using Pfu turbo DNA polymerase (Stratagene) were the same as described above F for the bacterial PCR -SSCP reaction, except for primer annealing temperature that was o r 51°C. All PCR reactions were performed using a GeneAmp PCR system 9700 thermocycler
	11 11 12 13	For bacterial community, the PCR -SSCP was performed by direct amplification of the 16S (Applied Biosystems). PCR products size and concentration were estimated by 2% agarose P were: W49 AGGTCCAGACTCCTACGGG (forward, E. coli position 330 (Brosius et al. r e rRNA gene (16S rDNA) V3 region using the W49-W104 primer couple. Primers sequences e gel electrophoresis and ethidium bromide staining.
	14 15 16 17	1981)) and W104 TTACCGCGGCTGCTGGCAC (reverse, E. coli position 533). Primer W104 was labelled with the 5'-FAM fluorochrome (Applied Biosystems, Perkin-Elmer). PCR reaction mix was as follow: about 100 ng of total DNA, Pfu turbo supplied buffer 1X, dNTP R e v 0.2 mM each, primers 2.6 ng/µl each, H 2 O up to 50 µl and 1.25 U of Pfu turbo DNA i e
	18	polymerase (Stratagene). After an initial denaturation step of 2 min at 94°C, 25 cycles were w
	19	applied with 30 s at 94°C, 30 s at 61°C and 30 s at 72°C. Reactions were ended by a 10 min
	20	elongation at 72°C and cooling at 4°C.
	21	For archaeal community, the PCR -SSCP attempted by direct amplification of the 16S rRNA
	22	gene did not give a positive signal on faintly colonised particles, probably because of the low
	23	amount of biomass attached to the particles. A nested PCR protocol was thus applied as
	24	follow. A first amplification of the total archaeal 16S rDNA was performed using primers
	25	W17 ATTCYGGTTGATTCCYGSCRG (forward, E. coli position 6); W02
		8

12

SSCP

electrophoresis and microbial community profiles analysis

	13
	14	SSCP capillary electrophoresis was performed as described in Delbès et al. (2000). One µl of
	15	PCR -SSCP product (diluted 10 to 20 times in water) was mixed with 18.8 µl of formamide
	16	and 0.2 µl of the internal standard GS-400 Rox (Applied Biosystems). The samples were then
	17	denaturated 5 min at 95°C and placed directly on ice for 5 min. The resulting single-stranded
	18	DNA fragments were separated by capillary electrophoresis using a CAP 5.58 %-Glycerol
	19	10% polymer and the ABI 310 Genetic Analyser equipped with a capillary tube (47 cm x 50
	20	mm) (Applied Biosystems). Laser detection of the sample PCR products and internal
	21	standards generated molecular typing profiles (so called SSCP profiles) that were aligned and
	22	compared using the GeneMapper software (Applied Biosystems). To facilitate comparison
	23	when a large number of profiles was analysed, each profile was scanned and represented as a
	24	row of discs where each disc corresponds to a distinguishable peak with its area being
		proportional to the peak proportion in the total profile.

  -1 according to each strategy. COD removal rate remained similar for both

	4
	5	reactors, above 87% with an average around 92%. The fraction of input carbon metabolized
	6	into VFA increased from 6.5% to values as high as 32% on day 30 for the R A and 17% on day
	7	38 for the R B . Carbon fractions dedicated to methane production, which were around 40% at
		the beginning of phase II, dropped accordingly.

8

From day 38 to 58 (phase III), COD removal efficiency and methane production 9 collapsed for both systems, while VFA accumulated to about 2.5 g COD .L -1 . The R A was more 10 affected, with COD removal efficiency falling below 60%. According to strategy A, the OLR 11

  CO 2 to produce acetate(Kuhner et al. 1997). It can thus be supposed 1 that they took an active part in acetate production in both reactors, thus supporting growth of Sporomusa was exchanged with the same Sporomusa species as the one observed in R A

	Thus, the simplest scenario during the first days after inoculation would be that ethanol was degraded into methane via the two conventional acetotrophic and hydrogenotrophic ways limitation, probably took part in the strong accumulation of acetate in R A during all the initial phase of the study (until the 59th day). Furthermore, acetate accumulation could also be a homoacetogens using H 2 -3 1 2 1 2 promoted respectively by the Sporomusa-Methanosaeta and Desulfovibrio-Methanobacterium 3 consequence of the much lower maximum rate of acetate consumption compared to the one of 3
	4 4 4	The Desulfovibrio are type I sulphate-reducing bacteria that use sulphate as terminal couples. The absence of visible methane production during the first 10 days of reactors start ethanol consumption during syntrophic ethanol degradation by anaerobic biofilm (Wu et al.
	5 5 5	electrons acceptor. Many species have the capacity to function either like hydrogen up and the requirement for a nested PCR for detection of the archaeal groups suggest a very 1991).
	6 6 6	consumers or producers. In a sulphate rich environment, they use simple substrates such as low amount and activity of the methanogenic groups. Concerning R B , the Desulfovibrio population dominant at the beginning of the
	7 7 7	ethanol, fumarate, malate or pyruvate as carbon and energy sources and oxidize it After two weeks of run, the two dominant bacterial groups were gradually replaced by experiment was replaced by two Geobacter populations while the initial species of
		F o r incompletely to form acetate and carbon dioxide by producing hydrogen sulphide (H 2 S). They F o r other species that differed between reactors A and B. Within R A , the Desulfovibrio population F o r
		P P P
		e e e
		e r e r e r
		R e v R e v R e v
	18 19 19	i e w i e i e OLR increase. It varies from 0.75 at the beginning of the experiment for an OLR equal to 0.5 w w
	20 20	gCOD.L -1 .d -1 to 30 mg.L -1 for an OLR of 20 gCOD.L -1 .d -1 at the end of the experiment. Thus, whatever the sulphate and/or hydrogen concentrations were in the reactor, this bacterial
	21 21	However, concentrations ratio remained constant (equal to 660) during all the study. This high group was probably involved in AGV production and acetotrophic methanogenesis.
	22 22	value of COD/sulphate ratio suggests that the Desulfovibrio identified in both reactors had a In R A , from day 16 to 72, the bacterial community was thus dominated by Pelobacter and
	21
		metabolism turned towards hydrogen production and did not compete with the methanogenic Sporomusa populations whose fermentative metabolisms were turned towards acetate,
		pyruvate, lactate, glycerol, methanol, ethanol, formate and fructose. In most cases, and
	25
	24
		15 16

22

particularly when the electrons donor is ethanol, the final reduced compound is acetate 23

(Kuhner et al. 1997; Biebl et al. 2000)

. Alternatively, they can grow as autotrophic 2 acetotrophic methanogens like Methanosaeta. 8 consume hydrogen and compete with the hydrogenotrophic methanogenic Archaea (Widdel 9 1988). On the opposite, in anaerobic environments deprived of sulphate, or in the presence of 10 very weak sulphate concentrations, they can produce hydrogen by using substrates like lactate 11 or ethanol and be involved in an interspecific H 2 transfer with hydrogenotrophic 12 methanogens, represented here by Methanobacterium. A study of Raskin et al. (1996) 13 mentions the persistence of a significant number of Desulfovibrio in the biofilm of an 14 anaerobic digester in the absence of sulphate. 15 The sulphate concentration in both processes, or rather the COD/sulphate ratio, will thus 16 strongly influence the metabolism of the Desulfovibrio ( Percheron et al. 1997). In this study, 17 sulphate concentration in the influent increased during all the start-up proportionally to the 23 Archaea (Hao et al. 1996). 8 was replaced by a Pelobacter population while the initial species of Sporomusa was 9 exchanged with another species of the same genus. Members of the Pelobacter genus are 10 described either as syntrophic bacteria oxidizing ethanol, propanol and butanol in their 11 corresponding acids thanks to a fermentative metabolism with hydrogen production, or as 12 bacteria consuming hydrogen and ethanol to produce acetate or propionate with simultaneous 13 Fe (III) or sulphate reduction (Laanbroek et al. 1982; Schink et al. 1987; Tholozan et al. 14 1990

). In the presence of sulphate, Pelobacter propionicus oxidizes ethanol in acetate by 15 using sulphate as final electrons acceptor. In the absence of sulphate and with high enough 16 hydrogen partial pressure, ethanol is condensed with bicarbonate or acetate to form propionate 17 or butyrate respectively. At low hydrogen partial pressure, the reaction also produces acetate 18 and propionate (Bornstein et al. 1948; Braun et al. 1981; Eichler et al. 1984; Wu et al. 1996). 23 propionate or butyrate production. This strong acidogenic activity, associated with the 24 decrease of Archaea methanogens acetoclastic activity consecutive to micronutrient 8 evolution. Within the Deltaproteobacteria, members of the genera Pelobacter, Geobacter, 9 Desulfuromonas and Desulfuromusa form a monophyletic group (Lovley et al. 1995; 10 Lonergan et al. 1996). However, in contrast to Pelobacter, Geobacter species have a 11 respiratory metabolism with Fe (III) or sulphur S 0 serving as common terminal electrons 12 acceptors. They can completely degrade ethanol, acetate and other short-chain fatty acids like 13 butyrate or propionate to carbon dioxide (Cord-Ruwisch et al. 1998; Coates et al. 2001). The

Table 1

 1 Carbon mass balance in percentage of carbon input as ethanol. Data are the mean of 1 measurements realized during the considered period.

	2
	3
	4	Reactor A:
		Phase	Length (d)	VFA Bicarbonate	CH 4	CO 2	AVS	VSS	Total
		I	10	8.0%	35.6%	19.1%	3.8%	19.5%	4.5%	90.5%
		II	28	23.2%	23.8%	37.5%	1.5%	5.4%	2.6%	94.1%
		III	21	41.3%	13.7%	32.2%	1.1%	2.4%	3.4%	94.1%
		IV	22	6.1%	17.4%	54.3%	9.3%	4.0%	2.9%	94.0%
		Overall	81 18.6%	17.6% 45.1%	5.7%	3.9%	3.1% 94.1%
	5
	6	Reactor B:
		Phase	Length (d)	VFA Bicarbonate	CH 4	CO 2	AVS	VSS	Total
		I	10	6.3%	44.2%	17.6%	4.6%	19.7%	3.9%	96.3%
		II	28	11.0%	29.3%	41.1%	2.4%	6.1%	3.2%	93.2%
		III	21	27.1%	22.2%	36.9%	2.7%	1.9%	3.5%	94.2%
		IV	22	3.1%	22.1%	53.0%	9.4%	5.4%	2.8%	95.7%
		Overall	81	11%	23.3%	47%	6.4%	4.6%	3.1% 95.5%

Table 2

 2 Phylogenetic affiliation of 16S rDNA sequences

	1
	2
	Peak	Phylum	% ID	Closest	Accession No	Isolated from
	name			microorganism or
				environmental
				clone
	A Clostridia	100	Uncultured	AB266923	Anaerobic UASB reactor
				bacterium clone
				SmB60f1
	B Bacteria	98	Uncultured	AJ852366	Gut Compartments of
		(environmental		bacterium clone		Melolontha melolontha Larvae
		samples)		MKEW-104		(Coleoptera: Scarabaeidae)
	C Bacteria	99 Clone SJA -117	AJ009488	Anaerobic trichlorobenzene-
		Chloroflexi				transforming microbial
						consortium
	D Deltaproteobacteria	98 Desulfovibrio sp.	AY770382	zinc-smelter wastewater effluents
		Desulfovibrio		A2
	E Deltaproteobacteria	98 Desulfovibrio sp.	AY770382	zinc-smelter wastewater effluents
		Desulfovibrio		A2
	F Deltaproteobacteria	99	Pelobacter sp.	AY692043	anaerobic biofilm
		Pelobacter		clone M113		UASB reactor
	G Deltaproteobacteria	100	Anaerobic	AB231802	anaerobic ethanol-oxidizing
		Geobacteraceae		syntrophic		bacterium from an anaerobic
				bacterium NE23-3		digested sludge
	H Deltaproteobacteria	99	Uncultured	AJ009496	anaerobic, trichlorobenzene-
		Geobacteraceae		bacterium clone		transforming microbial
				SJA-152		consortium
	I Bacteria	97	Uncultured	AF275913	anaerobic digester fed with
		(environmental		eubacterium		glucose
		samples)		AA01
	J Firmicutes	94	Sporomusa sp.	Y17760	anoxic bulk soil of flooded rice
		Sporomusa		DR6		microcosms
	K Firmicutes	95 Sporomusa ovata	AJ279800	spore-forming anaerobe that
		Sporomusa				ferments sugar alcohols
	N Archaea	99	Uncultured	AB233298	methanogenic UASB granular
				archaeon clone		sludge
				CG-8
	O Euryarchaeota	99 Methanobacterium	AY552778	anaerobic digesters
		Methanobacterium		beijingense

Microbiol 133, 178-184.


Bijaya K Adhikari
email: bijayaadhikari@yahoo.com

Suzelle Barrington
email: suzelle.barrington@mcgil.ca

José Martinez
email: jose.martinez@cemagref.fr

Urban food waste generation: challenges and opportunities

Keywords: urban, municipal solid waste, food waste, landfill, methane, compost, fertilizer, community

Greater economic activity and a wider economic gap between rural and urban areas have accelerated the growth of cities worldwide, along with their waste management issues. As a result, urban food waste (UFW) generation is expected to increase by 35% from 2007 to 2025. This paper examines the possible solutions to implement the environmentally safe recycling of UFW. If landfilling seems to be easy and economical for developed countries, it is not affordable for many large cities of Asia, Africa and South America. The on-site recycling of UFW is a more sustainable solution, but can only be justified economically if properly organized, community supported and recognized for its environmental benefits. On-site UFW composting is a solution which has already been implemented with success and which, if introduced worldwide, could provide agricultural soils with a good source of organic matter, capable of improving water management and fertilizing 3 million ha/yr. As compared to composting, anaerobic digestion could provide energy as biogas for a few high energy demanding industries within cities while also producing some organic soil amendment, after dewatering. Accordingly and for Asia and Africa, the on-site composting and anaerobic digestion of UFW could reduce the mass of MSW by 43 and

Author-produced version of the article published in Int. J.

Dr. José Martinez currently leads as research director, the Environmental Management and Biological Treatment of Wastes Research Unit at Cemagref, a public research body in France. He is also the former coordinator for the FAO research network Ramiran (Recycling of Agricultural Municipal and Industrial Residues in Agriculture Network), and adjunct Professor with the Department of Bioresource Engineering of McGill University. In 2004, he was invited by McGill University to deliver a semester course on "livestock farming and climate change". He has published over 50 scientific and technical papers on livestock slurries and manures. He has strong links with other member states through involvement in various competitively achieved EU projects including MATRESA, ALFAM, NUMALEC. Currently, his research interests are centered on N transformation processes in soils and manures, the reduction of greenhouse 1 Introduction Besides population growth, two main factors will impact the earth's environment in the upcoming decades: economic growth especially for countries with a large population, like India and China; and greater economic disparity between rural and urban areas, driving the rural population into cities. While the first factor leads to a greater demand for resources, such as fossil fuels, metals, water and food, the second factor leads to a more costly waste management burden on large cities. Cities world wide are already experiencing waste handling problems and smog issues [START_REF] Kumar | Cost-benefit analysis of landfill system witrh gas recovery for municipal solid waste management: A case study[END_REF].

From 20 to 80% of the mass of municipal solid waste (MSW) is made up of UFW. The UFW percentage was found to be inversely correlated to the economic status of the community, while the mass of UFW produced per capita was directly correlated [START_REF] Adhikari | Predicted growth of world food waste and methane production[END_REF]. In many countries around the globe, the landfill practice is not even feasible, resulting in land and water dumping [START_REF] Louis | A historical context of municipal solid waste management in United States[END_REF][START_REF] Korfmacher | Solid waste collection systems in developing urban areas of South Africa: an overview and case study[END_REF].

Because of its biodegradability, UFW attracts disease vectors such as parasites, pathogens, insects and vermin [START_REF] Louis | A historical context of municipal solid waste management in United States[END_REF][START_REF] Yedla | Economic evaluation of a landfill system with gas recovery for municipal solid waste management: a case study[END_REF] and its proper disposal can improve the environment and reduce health risks. In Asia, for example, a large number of cities can generally afford the management of 10 to 30% of their MSW [START_REF] Sharholy | Municipal solid waste management in Indian cities -A review[END_REF]. Because the UFW fraction of such cities generally constitutes 50 to 70% of MSW, the on-site treatment of UFW could resolve most of the garbage issue, and in parallel, reduce the mass of MSW to be transported outside city limits, as well as the smog and greenhouse gas emissions.

Where landfilling is practiced, the UFW fraction has a major impact on water, soil and air resources. Besides reducing the amount of land available for food production, landfilled UFW brings moisture producing contaminated leachate which can pollute groundwater and soils with heavy metals and toxic organic compounds [START_REF] Louis | A historical context of municipal solid waste management in United States[END_REF][START_REF] El-Fadel | Landfill evolution and treatability assessment of high-strength leachability from MSW with high organic and moisture content[END_REF]. Landfilled UFW accelerates global warming by producing greenhouse gases (GHG) composed of 60% methane (CH 4 ) and 40% carbon dioxide (CO 2 ) [START_REF] Adhikari | Predicted growth of world food waste and methane production[END_REF][START_REF] Kumar | Cost-benefit analysis of landfill system witrh gas recovery for municipal solid waste management: A case study[END_REF][START_REF] Bhide | Methane emission from landfills[END_REF][START_REF] Wang | Methane potential of food waste and anaerobic toxicity of leachate produced during food waste decomposition[END_REF]. Worldwide, the anthropogenic emission of greenhouse gases from landfill sites is estimated at 8% and results mainly from landfilled UFW. When installing systems to recover these greenhouse gases in young landfill sites, methane capture is costly and represent half of that produced from the anaerobic digestion of UFW [START_REF] Ortega-Charleston | Thermophilic adaptation of a mesophilic anaerobic sludge for food waste treatment[END_REF].

Recycling UFW is a challenge even for developed countries. In the late 1990's and for Europe, only about 0.7% of the total organic fraction of MSW was treated anaerobically [START_REF] Mata-Alvarez | Anaerobic digestion of organic solid Wastes' An overview of research achievements and perspectives[END_REF] and between 6 and 27% was composted [START_REF] De Baere | Anaerobic digestion of solid waste: state -of -the-art[END_REF] while the rest was sent to landfills along with the MSW main stream. In the early 2000's and for Canada, only the Maritime Provinces, with less than 10% of the country's population, were actively recycling over 50% their UFW [START_REF] Wagner | A new model for solid waste management: an analysis of the Nova Scotia MSW strategy[END_REF]. The success of this recycling program relied on higher tipping fees for garbage, reaching $100 US/ton, strict legislation and community training and involvement. Far from reaching this level in 2001, other Canadian provinces such as Quebec, with 20% of the country's population, were recycling only 7% of their UFW [START_REF] Da Costa | Creating a composting infrastructure for the city of Montreal[END_REF].

Although source separation and the on-site treatment of UFW looks attractive from a health and environmental point of view, it is far from being implemented because the general public is not convinced of its economic feasibility [START_REF] Isa | Solid waste collection and recycling in Nibong Tebal, Penang, Malaysia: a case study[END_REF]Kumar et al., 2004). The purpose of this paper is therefore to assess the challenges and economic benefits of the on-site recycling of UFW within an urban context, and to evaluate the opportunities resulting from the conversion of UFW into useful resources. The following sections will investigate the cost of UFW handling and landfilling versus the benefits of recycling the organic matter either as a fertilizer, through composting, or an energy source through anaerobic digestion.

Continental MSW and UFW generation

The estimated continental growth in MSW and UFW is presented in Figures 1a and 1b [START_REF] Adhikari | Predicted growth of world food waste and methane production[END_REF], assuming no changes in the present economic trend. With the largest share of world population, Asia produces the largest amount of MSW which is expected to increase from 617 Gkg/yr in 2007 to 967 Gkg/yr in 2025; during this time, UFW production will grow from 278 Gkg/yr to 416 Gkg/yr. Asia is followed by the Americas with 130 Gkg/yr of UFW in 2007, which is expected to increase to 174 Gkg/yr in 2025. Europe and Africa are producing 98 and 53 Gkg/yr of UFW in 2007, and this production is expected to reach 113 Gkg and 87 Gkg/yr in 2025.

INSERT FIGURES 1A AND B

The estimations presented in Figures 1a and1b consider the impact of the wider economic gap between rural and urban populations. In Asia, this gap is widening and attracting rural communities to move towards cities with the expectation of a better life. [START_REF] Adhikari | Predicted growth of world food waste and methane production[END_REF] demonstrated a clear relationship between the economic development of a country and the displacement of its population from rural areas to urban centres. By maintaining their present economic growth, most Asian countries such as China and India, could see their urban population growing from the present 50% to a future 70%, such as found in Europe and North America (Table 1). The recycling of UFW is not an issue for rural populations because of the space and land at their disposal.

INSERT TABLE 1

In comparison to [START_REF] Adhikari | Predicted growth of world food waste and methane production[END_REF], [START_REF] Sharholy | Municipal solid waste management in Indian cities -A review[END_REF] estimated that 90% of the MSW produced in India's urban centres is improperly disposed of, leading to serious environmental and health risks for the population. [START_REF] Sharholy | Municipal solid waste management in Indian cities -A review[END_REF] also reported that the general urban population was producing some 0.4 kg/capita/day of MSW in 1999 and that its UFW content ranged from 45 to 60%, for a net UFW production of 0.2 kg/capita/day. [START_REF] Adhikari | Predicted growth of world food waste and methane production[END_REF] estimated that in 2007, due to economic growth, the urban population of India could be producing 0.34 kg/capita/day of UFW (Table 1).

In the following sections, estimates of resource requirements will be presented for countries around the world to properly dispose of all their UFW through landfilling.

Then, the savings in resources will be estimated if on-site composting and anaerobic digestion were used instead of landfilling.

Disposal of UFW through landfilling

Landfilling is presently the most widely accepted practice for the disposal of MSW [START_REF] De Baere | Anaerobic digestion of solid waste: state -of -the-art[END_REF]. Although this solution is attractive to large cities, it lacks in scope and overall resource sustainability besides creating a poor environment for those living close to the landfill site. Environmental issues introduced by landfill sites are: risk of groundwater contamination from the leachate, greenhouse gas emissions, truck circulation bringing noise, dust and smog, and garbage decomposition creating odours. Furthermore, landfills require more often than not, the sacrifice of good agricultural soils which otherwise could be used to feed the world.

Land used for UFW landfilling

In terms of agricultural land, landfills occupy large areas which can no longer be used for food production, because of contamination risks. Typically and with a density of 260 to 500 kg/m 3 , each Gkg of urban UFW requires 33 ha of land when piled to a height of 15 m using a waste to soil cover ratio of 5:1 [START_REF] Peavy | Environmental Engineering[END_REF][START_REF] Bhide | Methane emission from landfills[END_REF]. When extrapolating this number and assuming that all future UFW generation will be landfilled, MSW management is observed to have an impact on world agricultural land and food production capacity.

In Asia and for 2007, the annual land use is estimated at 9174 ha for UFW landfilling and this area is expected to increase to 13728 ha in 2025. Similarly, the Americas will require 5742 ha/yr in 2025, which is a 40% increase compared to that required in 2007. Europe and Africa (Figure 2) follow in terms of land use for landfilling, with an estimated increase of 18 and 70% in 2025 as compared to 2007, respectively.

INSERT FIGURE 2

Over the upcoming 18 years (2007)(2008)(2009)(2010)(2011)(2012)(2013)(2014)(2015)(2016)(2017)(2018)(2019)(2020)(2021)(2022)(2023)(2024)(2025), the landfilling of all UFW could require some 400 × 10 3 ha of land with a deep soil profile. In Asia alone, the land area required for such purpose is estimated at 210 × 10 3 ha, representing 53 % of total global land requirement for landfilling. If UFW was recycled, an equivalent amount of land could be maintained in agricultural production to produce food for the Asian population.

A surface area of 210 × 10 3 ha can grow enough wheat, at 2.0 metric ton/ha, to provide a population of 1.0 million with 100% of its carbohydrate requirements.

Many large cities around the globe are suffering from lack of landfilling space for the disposal of their MSW. The city of Singapore plans to build an off-shore island of 350 ha to dispose of its MSW for the upcoming 30 years; the creation of this land base alone will cost almost $10 US/ton of MSW, while the land base for a landfill site in Ohio State cost $0.10 US/ton (Ohio State University, 2001). Where the land and handling costs cannot be afforded, open spaces, street corners and river banks serve as uncontrolled dumping sites for as much as 90% of all MSW [START_REF] Sharholy | Municipal solid waste management in Indian cities -A review[END_REF].

In developed countries, landfill sites use good agricultural land which is becoming a national priority, and bring other forms of regional pollution besides. Landfill sites are often forced onto rural communities to allow urban centres to dispose of their MSW, and as such, introduce additional contamination in the form of truck traffic, dust and odours.

Recently, North American authorities introduced laws for the cleaning and restoration of old landfill sites (Minnesota Pollution Control Agency, 2004). Urban sprawl is becoming a threat to agricultural land, as many government authorities are initiating laws to preserve farm land; agricultural land protection legislation has been implemented in Canada, the United States, Australia and Europe (USDA, 2005). In the United States from 1992 to 2001, urban regions have tripled in area, at a growth rate of 100 × 10 3 ha/yr. During this same period, landfilling operations have reduced US land surfaces designated for agriculture at an additional rate of 1.5 × 10 3 ha/yr. Finally, agricultural land will be even more pressing in the years to come, as the transformation of corn into ethanol is expected to compete with fossil fuel production. From 2005 to 2025, in the Americas alone, landfill sites will remove 110 × 10 3 ha of land, which could produce enough ethanol to replace a net (energy above that required to produce the ethanol) amount of 147 × 10 3 m 3 of diesel/yr (USDA, 1995). Although this diesel represents only a small percentage of the world demand of 4 million m 3 of diesel/day for 2007 (US EIA, 2007), it represents 1.3% of Canada's 2007 consumption. In terms of meeting the Kyoto agreement, this is a major share of Canada's 6% reduction commitment.

Cost of handling and landfilling urban UFW

The land base required for landfill sites often represents the tip of the iceberg [START_REF] Isa | Solid waste collection and recycling in Nibong Tebal, Penang, Malaysia: a case study[END_REF]. In North America, the purchasing of land represents only 0.5% to 1.0% of the total cost of handling and disposing of MSW. Besides the land base, environmentally safe landfills require the expertise of professionals, the construction of access roads, the installation of impermeable membranes and the collection and treatment of leachate. In Although the handling of all UFW, world wide, would only use 0.4% of the world's diesel consumption, it is still equivalent to 50% of Canada's requirements. Nevertheless, the energy cost generally represents 10% to the total cost of handling and disposing of MSW, while the labour cost represents 70% and the capitalization and maintenance costs add up to 20%.

Mumbai spends some $29 US/ton for the collection, transportation and landfilling of what ever MSW can be handled (Department of Chemical Engg, 1997). Similarly and in 2000, the Malaysian MSW handling cost of $16 US/ton, along with its waste generation of 0.8 kg/capita/d [START_REF] Tean | Solid waste collection and recycling in Nibong Tebal, Penang, Malaysia: a case study[END_REF], implies that $105 million US/year would be required to the waste generated by a population of 23 million [START_REF] Isa | Solid waste collection and recycling in Nibong Tebal, Penang, Malaysia: a case study[END_REF]. The low labour cost, despite higher land cost, makes landfilling a cheaper operation in Asia than in North America.

If the present landfill practice is maintained and the entire wet mass of UFW is landfilled, the cities around the world are facing huge economic challenges (Figure 3). At present, Asia needs $10 billion US/yr for the handling of its urban UFW and is expected to spend $15 billion US/yr more in 2025. The Americas are next, presently spending $4.5 billion US/yr and expected to increase this spending by $6.2 billion US/yr, in 2025.

Whereas in Europe, $3.4 billion US/yr is estimated for 2007, the cost is expected to increased by 18% in 2025; Africa needs $1.8 billion US$/yr in 2007 and its cost is estimated to increase by 71%, in 2025.

INSERT FIGURE 3

Landfilling brings no economic benefit, besides providing employment. The recovery of methane is costly and only 50% efficient [START_REF] Ortega-Charleston | Thermophilic adaptation of a mesophilic anaerobic sludge for food waste treatment[END_REF]) as compared to directly digesting UFW using an anaerobic digester. Some 200 L of methane/ton of landfilled UFW (dry basis) [START_REF] Wang | Methane potential of food waste and anaerobic toxicity of leachate produced during food waste decomposition[END_REF] need collecting and flaring at a cost of $4.00 US/ton of wet UFW [START_REF] Ngnikam | Evaluationof potentialities to reduce GHG emissions resulting from various treatments of municipal solid wastes in moist tropical climates: application To Yaounde[END_REF]. The transformation of this methane into electricity is even more expensive, because of its low conversion efficiency of 35%.

Total cost of landfilling urban UFW

In total, the collection, transportation, landfilling of UFW and the greenhouse gas ramification is a non sustainable operation which costs in the range of $ 20 to $40 US/ton UFW, depending on labour and land costs. For developing countries, this cost is often not affordable despite the health issues at stake. For developed countries, this cost is too affordable, and the resources required are relatively limited, explaining the lack of UFW recycling. The on-site recycling of UFW will be implemented for both developing and developed countries, when additional benefits are considered, from an aspect of environment and resource conservation.

On-site treatment and utilization of urban UFW

Considering the environmental issues associated with the landfilling of the world's UFW, alternative sustainable solutions must be developed. The implementation of these solutions requires the involvement of communities, through awareness programs and incentives, and the logical valorization of the organic waste.

Source separation

Source separation is a pre-requisite for the on-site treatment of UFW to reduce MSW handling, transportation and landfilling costs, especially for large cities. In 2025, such practices could reduced by 43% the total MSW disposal cost in Asia, 29% in Americas, 29% in Europe and 53% in Africa (Figure 4).

INSERT FIGURE 4

In the past, many countries around the world have developed large scale composting centers for organic waste which was mechanically separated from MSW.

Many of these composting centres have not been kept in operation [START_REF] Sharholy | Municipal solid waste management in Indian cities -A review[END_REF] for three reasons: the poor quality of the compost resulting from its contamination with glass, heavy metals and other pollutants into contact with the acidic leachate during handling; the workers often get hurt while handling and sorting the various components of the MSW; and the value of the compost produced being lower than the cost of composting [START_REF] Furedy | Recycling organic wastes for food production in asian cities: sanitary and economic perspectives[END_REF].

Composting source separated UFW -a success story

The authors of the present paper were involved with the implementation of a successful on-site composting centre operating in the heart of the city of Montreal, Canada [START_REF] Barrington | An urban composter to induce the culture of residential organic waste sorting[END_REF]. In operation since 2003, the composting centre is located in a municipal park, along a popular pedestrian path. The 125 participating families are source selecting and dropping of their UFW on their way to work in the morning, three times weekly, when an attendant operated the centre. The composting centre is visible to all walking by, and those curious enough can stop to ask questions.

The successful operation of this 6.0 m 3 centre was based on three main elements: the implementation of community awareness, training and educational programs before opening the centre; the easily accessible and visible site selected for the urban composting centre; and the fact that most of the local residents were highly educated, with at least a university degree.

In 2007 and with a capacity of 1.0 tons of high quality compost/week, this Montreal centre operates at a cost of $220 US/ton of finished material or $140 US/ton of wet UFW processes. Labour, power, bulking agent and capitalization represent 40, 1, 9 and 50% of the costs. As bulking agent, the centre uses pellets of cereal flour residues, because of their high moisture absorbing capacity. The operating cost of this centre could be reduced to $70 US/ton of UFW ($110 US/ton of compost produced) if the capacity was increased to 5 tons/week of compost. Although similar in cost to that of larger operations, composting obviously does not seem economical when compared to landfilling and its methane recovery at $40 US/ton UFW (US EPA, 2003; US EPA, 1999).

Proposed on-site UFW composting for Mumbai, India

Flooding is a threat to public health in Indian cities because of the uncontrolled street dumping of MSW. Composting could therefore be a highly suitable solution, along with the fact that once recycled as soil amendment, water conservation and soil fertilization practices can be enhanced [START_REF] Kumar | Cost-benefit analysis of landfill system witrh gas recovery for municipal solid waste management: A case study[END_REF].

Mumbai is one of the biggest cities in India with a population 27000 per km 2 (Demographia, 2007). Its MSW management is one of the major challenges because of a waste production rate exceeding 0.50 kg/person/day (Department of Chemical Engg, 1997) with over 50% UFW [START_REF] Adhikari | Predicted growth of world food waste and methane production[END_REF].

Table 2 presents a scenario where the Montreal park composting centre could be implemented in Mumbai. To deal with the intense population, it is proposed to use 6 invessel composters per km 2 , each with a capacity of 27m 3 . The operating cost for such centre could be lower than that of Montreal, because of cheaper labour costs and the larger capacity of the system. The compost production cost is estimated at $85 US/ton, implying a cost of eliminating UFW of $53.00 US/ton on a wet basis. Nevertheless, these composting centres would first of all require the participation of the population to source separate UFW and bring it to the composting centres. Because not all the population is environmentally aware and committed to environmental issues, monetary incentives may be required.

INSERT TABLE 2

As for Montreal, this composting cost of $85.00 US/ton for Mumbai is still above that of $39.00 US for landfilling. Hence, promoting the on-site recycling of UFW requires the demonstration of additional benefits, for both developed and developing countries.

Advantages of on-site UFW composting

Once composted, UFW are stabilized, of lower moisture content and have lost 50% of their original mass (US EPA, 2003). At present, Asia has a compost production potential of 83 Gkg and its production is expected to increase by 51% in 2025 (Table 3). In the Americas, compost production from UFW can potentially amount to 52 Gkg in 2025, which is 37% higher than that of 2007. In Europe and Africa, the compost production potential is estimated at 34 and 26 Gkg, which represents an increase of 18 and 73% compared to 2007.

INSERT TABLE 3

To compare the cost of composting UFW with that of landfilling, all benefits must be considered (Table 4). First of all, this compost has a mineral fertilizer value because each dry ton of compost produced from UFW can potentially offer 24 kg N, 2.4 kg P and 32kg K [START_REF] Adhikari | Predicted growth of world food waste and methane production[END_REF]. In Asia, the value of N, P and K is $0.25, $0.65 and $0.65 US/kg, while in North America the value is $1.15, $1.15 and $0.60 US/kg [START_REF] Barrington | Personnel contact in Canada and at Tamil Nadu University of Coimbatore[END_REF]. The cost recovery from the fertilizer value of compost produced from the entire Asian urban UFW is estimated at $1418 million US/yr in 2007 and is expected to increase by 50% in 2025, while in the Americas it will reach $1550 million US/yr in 2025. In Europe and Africa, the 2025 value is estimated at $995 and $435 million US/yr. If all UFW was properly separated and composted in Asia, some 3 million ha could be fertilized with 40 kg of P/ha (the equivalence of 28kg of P/ha in mineral fertilizer), which in turn, could feed 12 million persons. Illegally disposed of along river banks or leached to groundwater as landfill leachate, this phosphorous has a tremendous eutrophication impact.

INSERT TABLE 4

The application of UFW compost, at a rate of 40 kg of P/ha/yr, adds 33 tons/ha/yr of organic matter or 1.4%/yr, over a depth of 200mm. Such rate of organic matter application to land can help reduce erosion and improve fertilizer and water absorption, thus advantage the quality of water resources. The value of organic matter in reducing risks of erosion and mineral fertilizer leaching needs further investigation, but in this paper will be evaluated as equal to its mineral content. Similarly, 1.0kg of dry compost can absorb up to 3.0kg of water and 2.0kg of this water is readily available to plants [START_REF] Soussi | Biofiltration and greenhouses gas reduction in manure pit and livestock facilities' NABEC technical conference[END_REF]. The land application of compost, at a rate of 33 tons/ha/yr, can therefore improve water retention by 7 mm following rainfall events, and thus reduce irrigation requirements. This is particularly interesting because water is becoming an increasingly precious commodity. Table 4 estimates the water holding value of UFW compost using a cost for wholesale water of $0.50 US/m 3 , which corresponds to water obtained from reservoirs. When the water is obtained by desalination, the cost ranges between $0.75 to $1.00 US/m 3 [START_REF] Hamer | The cost of water and water markets in South California, USA[END_REF].

Finally, the on-site composting of UFW eliminates a major portion of the MSW handling, landfilling and greenhouse gas recovery costs. On-site composting in urban areas allows local residents to bring their UFW to the composting centre as opposed to hiring expensive equipment for the collection. Furthermore, composting reduces the mass of FW by 50%, even after adding a bulking agent, and this compost can be used on farm land on the outskirts of the city whereas landfills are further away to avoid nuisance.

Accordingly, collection and transportation of UFW can be reduced by over 50% and landfill greenhouse gas emissions are dropped to negligible levels, eliminating the need for their recovery.

When considering all benefits introduced with the on-site composting of UFW, composting becomes a profitable operation. Considering all the added benefits as soil amendment, farm land users will most likely be willing to use the compost in Asia as well as in North America. The high quality of the compost will be an added value.

On-site anaerobic treatment of UFW and energy generation

Besides composting, anaerobic digestion is the other available technology to treat UFW on-site. Anaerobic digestion produces biogas which can replace natural gas, does not require the purchase of bulking agents and can deodorize UFW rather than release odours. Nevertheless, composting produces a dry, stabilized and volume reduced product, as opposed to anaerobic digestion which generates a large volume of wet and nondisinfected sludge [START_REF] Bagge | Hygienic aspects of recycling of organic waste in biogas plants[END_REF] requiring dewatering to produce a high quality soil amendment.

At present, the anaerobic treatment of the Asian UFW can produce 17 Gkg of CH 4 and this production is estimated to increase to 25 Gkg, in 2025 (Table 5a). If burned to get the highest conversion efficiency of 56 MJ/kg CH 4 [START_REF] Edm | Energy density of methane[END_REF], the 2025 energy recovery potential from UFW in Asia is equivalent to 1400 x 10 12 KJ/yr which also represent 44 million kW of energy generation on a continuous basis (Table 5b).

Similarly, the urban UFW from the Americas, Europe and Africa can produce 616 x 10 12 KJ, 392 x 10 12 KJ and 280 x 10 12 KJ of energy respectively, in 2025. If all urban UFW could be anaerobically treated on-site, the total global energy recovery for 2025 is estimated at 2688 x 10 12 KJ; Asia alone could produce 52% of total energy recovery.

Nevertheless, this only represents 1.4% of the energy offered by the 15 million m 3 /d of crude oil presently utilized by the world.

INSERT TABLES 5a and b

When compared to the total world energy demand including that obtained from crude oil, the conversion of UFW into biogas represents less than 0.5% (Table 6). Even for India, UFW could only generate 0.66% of all of its energy needs. For the anaerobic digestion of UFW to produce at least 1% of the world energy, in terms of biogas, wastewater sludge and green wastes (ex. grass clippings) will have to be added.

INSERT TABLE 6

Justifying biogas production from UFW therefore requires its use by a plant with a high demand for clean fuels, such as a food processing plant, a foundry or a community heating plant. The anaerobic digestion of the UFW and sludge produced by a population of 1.0 million represents 10 GW of power (100% efficiency), which becomes interesting when used by a single large plant. Sludge added to the UFW can provide sufficient liquid to liquefy the UFW for its anaerobic digestion, but introduces additional health risks for the workers.

Community UFW anaerobic treatment centres for Mumbai, India

Energy is one of the major issues for the city of Mumbai, India. Each anaerobic digester implemented per km 2 of urban area could generate 130 kW of power on a continuous basis, assuming 100% efficiency (Table 7). This energy could easily be used by a factory or industry, but could not be distributed to all the contributing families. As opposed to 6 composting centres/km 2 , one anaerobic digester/km 2 is preferred to optimize the size of the operation. Thus, local communities would need more incentives to bring their UFW for treatment. The sludge produced would need pasteurization and dewatering before being suitable for soil fertilization [START_REF] Hartmann | Efficiency of the anaerobic treatment of the organic fraction of municipal solid waste: collection and pretreatment[END_REF]. Sludge produced by the anaerobic digesters could be used as soil amendment by nearby gardens rather than farm land outside the city limits, to eliminate the need for dewatering.

INSERT TABLE 7

The benefits of anaerobically digesting UFW can be analyzed, as done for composting (Table 8). The anaerobic digestion of UFW generates less organic matter but just as much mineral fertilizer as composting. In exchange, biogas produces energy and eliminates most cost associated with landfilling, except for the collection and 75% of the transportation. As for composting, anaerobic digestion becomes interesting when considering the added benefits on the environment and world resources.

INSERT TABLE 8 5 Conclusions

The rapid urbanization and growing economic activity around the world accelerates the generation of municipal solid waste (MSW) and urban food waste (UFW) within city limits, especially in developing economy. While presently most large city in Asia, Africa and South America does not have the funds to even collect all of its MSW, those in North America and Europe do not see the incentives to recycle UFW.

The on-site recycling of UFW through composting or anaerobic digestion can be demonstrated as attractive when considering all benefits, including those having an effect on the environment and the world water, soil and air resources. Source separation is a pre-requisite to successfully compost and to anaerobically digest UFW, with the least risk of contamination for agricultural soils and the food supply chain. The organic matter recycled by such treatment technologies could produce a soil amendment capable of fertilizing 3 million ha/yr while increasing the soil organic matter by 1.4%/yr and the water retention by 7mm. If composting requires a plentiful source of bulking agents and a biofilter for odour control, anaerobic digestion requires additional hygienic protection for the operators and sludge users. [START_REF] Barrington | Personnel contact in Canada and at Tamil Nadu University of Coimbatore[END_REF] for compost at 50% moisture containing 12, 1.2 and 16 kg of N, P and K per wet ton; 2 less soil erosion and improve fertilizer adsorption efficiency for higher quality water resources; estimated at 100% of fertilizer value; 3 land saved from landfilling after 20 years of recycling UFW for 2 and 1 crops/yr in Asia and North America, valued at $1000/ha; 4 improved water holding capacity estimated at 1.2 m 3 /ton/yr for 4 years, after 6 rainfall events/year, at $0.5 US/m 3 [START_REF] Hamer | The cost of water and water markets in South California, USA[END_REF]; 5 saving 50% of collection and transportation costs ($16.00 and $1.40 US/ton UFW) plus 100% of landfill cost ($18.60 US/ton UFW) in North America, where 1.0 ton of compost represents 1.6 tons of UFW; in Asia, this cost is 20% lower; 6 composting produces 50% less methane than landfilling; 1.6 tons of UFW generates an additional 56 m 3 of methane when landfilled [START_REF] West | Table 3 Estimated annual continental compost production from urban food waste (UFW)[END_REF] at a cost of $4.00 US/ton UFW. [START_REF] Adhikari | Predicted growth of world food waste and methane production[END_REF]. Note : OE -oil equivalent; UFW -food waste; energy from food waste after anaerobic digestion and at 100% efficiency. All UFW production values are on a wet basis.

  2001, the implementation, operation (transportation of MSW to the site and burial) and closure of a safe landfill site for 5.4 Gkg of solid waste were estimated at $20 US/ton, in Ohio State, USA (Ohio State University, 2001). Since the collection cost is in the range of $16.00 US/ton (Smart Storage, 1998), the total MSW management cost amounts to $36 US/ton. The operations of collecting, transporting and landfilling consume fossil fuels at a rate of 10L/ton (Ohio State University, 2001; University of Tennessee, 1993).

Figure 1 Figure 2 Figure 3 Figure 4

 1234 Figure1Continental urban solid waste production on an annual basis: (a) municipal solid waste (MSW); and urban food waste (UFW) (Source:[START_REF] Adhikari | Predicted growth of world food waste and methane production[END_REF]

Figure 2 Figure 3 Figure 4

 234 Figure 2 Estimated land requirement for landfilling of urban food waste (UFW)

Table 1

 1 Total and urban population for various countries and their food waste generation

	Continent Country	Population	Food waste
					generation
			Total	Urban	UFW	UFW
			10 6	%	kg/c/d	Gkg/d
	Africa	Algeria	30.24	59.3	0.34	6.1
		Ethiopia	65.59	17.7	0.33	3.8
		Kenya	30.55	33.1	0.37	3.7
		Morocco	29.11	55.3	0.35	5.6
		Niger	10.74	20.6	0.50	1.1
	Asia	China	1275.2	52.0	0.34	243.0
		India	1071.0	42.0	0.34	153.0
		Kuwait	2.25	97.6	0.60	1.32
		S. Korea	46.84	86.2	0.60	24.2
	Europe	France	59.30	75.6	0.60	26.9
		Germany	82.28	87.5	0.60	43.2
		Poland	38.67	65.6	0.35	8.9
		U.K.	58.69	89.5	0.60	31.5
		Spain	40.75	77.6	0.60	19.0
		Sweden	8.86	83.3	0.60	4.4
	North	Canada	30.77	77.1	0.60	14.2
	America	USA	282.19	77.2	0.70	152.5
		Mexico	98.93	74.4	0.40	29.4
	South	Argentina	39.20	76.0	0.40	11.9
	America	Brazil	179.00	67.0	0.37	44.4
		Peru	27.60	61.0	0.36	6.1
		Chili	16.08	71.0	0.40	4.5
	Note: UFW generation is on a wet mass basis for 2007.

Table 2

 2

	Community composting centers for Mumbai, India
	Description	Numbers and specification
	*Population/km 2	27000
	Food waste production/person/day	0.25kg
	Number of composting center/km 2	6
	Number of in-vessel composters/center	2
	Specification of each composter	6m long and 2.4m diameter
	Composting cost (fixed and operational)	85 US$/ton wet compost
	*Demographia (2007)

Table 4

 4 Value of composted UFW for Asia and North America

	Valorization	Value of composted UFW
		$ US / ton
		Asia	North
			America
	Fertilizer value 1	$14.20	$24.75
	Organic matter advantages 2	$14.20	$24.75
	Production from land saved 3	$ 5.70	$ 2.85
	Agricultural water 4	$14.40	$14.40
	Saved landfilling cost 5	$35.70	$44.60
	Greenhouse gas recovery 6	$ 6.40	$ 6.40
	Total savings	$90.20	$117.75
	Composting cost	$85.00	$110.00
	Note:
	1 $0.25, $0.65 and $0.65/kg for N, P and K in Asia and $1.15, $1.15 and $0.60/kg in
	North America

Table 5a

 5a Continental methane (CH 4 ) production from the anaerobic digestion of UFW in Gkg/yr

	Year	Asia	Americas	Europe	Africa
	2005	17	8	6	3
	2010	19	8	6	4
	2015	21	9	7	4
	2020	23	10	7	5
	2025	25	11	7	5
	Source: Adhikari et al. (2006)
	Note : UFW -urban food waste.

Table 6

 6 Total energy need versus that recovered from UFW anaerobic digestion

	Continent Country	Energy consumption	Energy from UFW
			OE	Energy	UFW	Energy
			kg/c	kW/c/d	kg/c/d	of total
	Africa	Continent	1628	2.74
		Algeria	1038	1.74	0.34	0.35%
		Ethiopia	278	0.47	0.33	1.28%
		Kenya	481	0.81	0.37	0.84%
		Morocco	357	0.60	0.35	1.07%
		Niger	777	1.31	0.50	0.70%
	Asia	Continent	991	1.67
		China	1138	1.91	0.34	0.32%
		India	512	0.86	0.34	0.75%
		Kuwait	9076	15.25	0.60	0.07%
		S. Korea	4347	7.3	0.60	0.15%
	Europe	Continent	3700	6.2
		France	4518	7.6	0.60	0.14%
		Germany	4203	7.1	0.60	0.13%
		Poland	2370	4.00	0.35	0.16%
		U.K.	3918	6.59	0.60	0.16%
		Spain	3228	5.43	0.60	0.20%
		Sweden	5765	9.67	0.60	0.11%
	North	Continent	7844	13.20
	America	Canada	8300	14.00	0.60	0.08%
		USA	7795	13.10	0.70	0.10%
		Mexico	1533	2.57	0.40	0.28%
	South	Continent	1083	1.82
	America	Argentina	1575	2.65	0.40	0.27%
		Brazil	1066	1.79	0.37	0.37%
		Peru	432	0.73	0.36	0.88%
		Chili	1652	2.78	0.40	0.26%
	World		1674	2.81
	Earth Trends (2007);

Table 7

 7 Example of urban anaerobic UFW digester for Mumbai, India

	Description

Acknowledgement

The authors gratefully acknowledge the continued financial supported by the Natural Science and Engineering Research Council of Canada and cemagref, France.

, American Society of Agricultural and Biological Engineers (ASABE), Nepalese Society of Agricultural Engineers (NSAE) and Nepal Engineering Council (NEA). researches organic waste management and odour measurement and control. She is an active member of the Ordre des ingénieurs du Québec, is involved with Engineers Canada and is president of the Canadian Memorial Engineering Foundation supporting women in engineering with Canadian scholarships. She is a Fellow member of the Canadian Society of Bioengineering (CSBE) and a member of the American Society of Agricultural and Biological Engineers (ASABE). Along with Eco-quartier Jeanne Mance-Mile Ends, she received from the Canadian Geographic Society, the 2007 Canadian Environment Award for developing a fully functional and extremely active community composting centre in downtown Montreal.

List of tables:

Table 1 Total and urban population for various countries and their food waste generation 1 same fertilizer value as compost because both process loose nutrients through leachate (compost) and from dewatering (anaerobic sludge); 2 70% loss of organic matter versus 15% loss with composting; estimated at 35% of fertilizer value; 3 land saved from landfilling after 20 years of recycling UFW for 2 and 1 crops/yr In Asia and North America, valued at $1000/ha; 4 at 35% of the capacity of compost, because of organic matter losses during digestion; 5 1.0 ton of wet UFW produces 360 kW-h of energy at 80% efficiency when burned, worth $0.10 US/kW-h; 6 saving of 0% collection, 75% transport and 100% landfilling; 7 same recovery as for composting. 8 the cost of producing this energy without gas scrubbing is evaluated at $0.15 US/kW-h as experienced in Europe.