D16.1 Report on the biofilters performance · D16.1 Report on the biofilters performance . Due date...

018525

REMOVALS

Reduction, modification and valorisation of sludge SPECIFIC TARGETED RESEARCH OR INNOVATION PROJECT PRIORITY 1.1.6.3.: Global change and ecosystems

D16.1 Report on the biofilters performance

Due date of report: 31 MAY 2008 Actual submission date: 30 MARCH 2009

Start date of project: 1 JULY 2006 Duration: 3 YEARS Organisation name of lead contractor for this deliverable: UNIVERSITAT AUTÒNOMA DE BARCELONA Revision 2

Project co-funded by the European Commission within the Sixth Framework Programme (2002-2006) Dissemination Level

PU Public PP Restricted to other programme participants (including the Commission Services) RE Restricted to a group specified by the consortium (including the Commission Services) CO Confidential, only for members of the consortium (including the Commission Services) X

FP6-018525, REMOVALS, D16.1 REPORT, rev 2

TABLE OF CONTENTS 3

1. STATE OF THE ART 4

2. OBJECTIVES 4

3. ANALYTICAL METHODS 5

4. RESULTS

4.1. REDESIGN OF AN EXISTING BIOFILTRATION SETUP

4.2. EVALUATION OF ORGANIC MATERIALS AS PACKING BED IN BIOREACTORS

4.2.1. PHYSICO-CHEMICAL CHARACTERIZATION OF ORGANIC PACKING

MATERIALS

4.2.2. EVALUATION OF BIOFILTER´S PERFORMANCE PARAMETERS

4.2.2.1. EFFECT OF THE EBRT

4.2.2.2. INFLUENCE OF THE NUTRIENTS SUPPLY

4.2.2.3. EFFECT OF THE WATERING RATE

4.3. ASSESSMENT OF SLUDGE-BASED CARBON AS PACKING BED IN A BIOFILTER

4.3.1. PHYSICO-CHEMICAL CHARACTERIZATION OF COMMERCIAL ACTIVATED

CARBON

4.3.2. EVALUATION OF SLUDGE-BASED CARBON AS FILTER BED IN A BIOFILTER

4.3.3. ECONOMICAL ASSESSMENT OF THE USE OF SLUDGE-BASED CARBON AS

FILTER BED IN A BIOFILTER

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5. PUBLISHED, PUBLISHABLE AND IMPLEMENTABLE RESULTS 23

236. CONCLUSIONS

247. REFERENCES

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1. State of the art

Nowadays, the high rate of generation of biological sludge in waste water treatment facilities can be considered one of the main global environmental concerns. As of today, no optimal solution has been developed for the disposal of this sludge. However, it has been recently proposed that its conversion into activated carbon and its subsequent usage as packing material in bioreactors could mean a highly cost-effective and environmental-friendly way of taking advantage of this material. With the purpose of assessing this possibility in lab-scale bioreactors, a series of carbon- and classic filter bed materials-based biofilters were designed and operated. The present report shows the main results obtained in this sense. Bioreactors are a novel technology used for the treatment of gaseous pollutants. They can be described as containers, in most cases filled with a solid support, over which a microbial population develops. Further information on bioreactor technology can be found in Devinny et al. (1999) and Kennes and Veiga (2001). A number of works published by different research groups have shown that the packing material selected may influence the performance of bioreactors to a great extent (Kennes and Thalasso, 1998; Devinny et al., 1999; Maestre et al., 2007). It is, then, obvious, that proper packing material selection is a vital factor to be considered if high performance and stability of the bioreactor are desired. The main characteristics to be considered for a packing material are its specific surface area, chemical and mechanical durability, porosity, pH, water holding capacity, buffering capacity and elemental composition (Oosting et al., 1992; Bohn, 1996). A wide variety of organic and inert materials have been studied in the past, such as compost, perlite, lava rock, clay and polyurethane foam (Kennes and Thalasso, 1998), while others, such as coconut fibre or pine leaves, have been rarely tested in bioreactors so far, although it is expected that they could perform optimally, at a relatively low cost. In this sense, the use of activated carbon has proved to be a reliable alternative to organic packing materials, as its high adsorption capacity contributes to create a more stable environment in terms of pollutant concentration, resulting in higher microbial activity and increasing removal efficiencies. Prado et al. (2004), working with a series of lab-scale conventional biofilters intended for the depuration of gas-phase methanol, proved that the elimination capacity of an activated carbon-based reactor is comparable to those of reactors packed with other common inert packing materials, as perlite and lava rock. Maximum methanol elimination capacities above 300 g/m3h were found in all systems. Also, Yani et al. (1998) employed activated carbon fibre as carrier for a lab-scale ammonia-treating biofilter. In this study, a maximum removal capacity of 3.5 g N/kg dry carbon·day was observed. N-NH3 loads up to 1.5 g/kg dry carbon·day could be completely removed. In addition, Kwon and Cho (2009) studied the biological degradation of a mixture of benzene, toluene, ethylbenzene and xylene vapors in a biofilter packed with activated carbon. In this study a maximum elimination capacity of 67 g/m3·h was reached at a load of around 94 g/m3·h. Also, Rattanapan et al. (2009) studied the effect of inlet H2S concentration, H2S gas flow rate, air gas flow rate and long-term operation on the H2S removal efficiency of a conventional biofilter. It was found that the efficiency of the H2S removal was above 98% even at inlet H2S concentrations as high as 4000 ppm. The maximum elimination capacity found in this study was about 125 g H2S/m3of GAC·h. However, even though all of these studies prove the technical viability of using carbon as filter bed material, its high cost has prevented it to be more extensively applied in bioreactors. As of today, the existence of real-scale bioreactors employing activated carbon as packing material as not been reported, to the best of our knowledge. In addition to that, it has been proved that bacterial-based biofilters can show low elimination capacities when treating waste gases containing highly hydrophobic compounds, due to their low solubility.

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Different researchers have reported elimination capacities in bacterial-based bioreactors below 60 g toluene/m3h, under very different operating conditions (Leson and Winer, 1991; Liu et al., 2005). Additionally, low pH and lack of moisture can negatively affect the operational stability of bacterial reactors to a great extent. Oppositely to this, it has been proved that fungal biofilters may have significantly higher elimination capacities when a waste gas containing pollutants with a low solubility has to be treated (García-Peña et al., 2001; Aizpuru et al., 2005). This is due to the fact that, commonly, fungi are much more resistant than bacteria to low pH and low moisture content conditions. Additionally, it has been proposed that fungal-based biofilms can improve the solubility of hydrophobic compounds due to the higher surface area of the fungal mycelia (compared with bacterial biofilms), which increases the contact area between the biofilm and the gaseous pollutant (van Groenestijn and Liu, 2002). Also, the presence of nutrients will significantly influence the performance of the bioreactor. Commonly, organic packing materials serve as a natural source of nutrients (Leson and Winer, 1991), though, in many cases, an extra addition may be necessary in order to assure high elimination capacities in the long-term.

2. Objectives

The principal target of WP16 is to evaluate the feasibility of employing activated carbon produced from sludge obtained from a wastewater treatment plant as packing material in biological reactors for the abatement of odours generated in industrial facilities, as, for instance, in solid waste or wastewater treatment plants. Even though the research experience with such activated carbons is very limited, it is expected that their technical characteristics will not differ greatly from those of conventional activated carbons. Then, the main interest arises from the environmental benefits associated to the use of these carbons. In this work package, a series of bench-scale biofilters were designed and tested, packed with commercial organic packing materials and carbon. It is important to mention that, as sludge-based activated carbon was not available, commercial activated carbon and sludge-based (non-activated) carbon were used. It is expected that the results would be very similar if sludge-based activated carbon had been employed. Volatile Organic Compounds (VOCs) representative of several odor sources such as solid waste treatment facilities or wastewater treatment plants were used as model compounds. The present report includes the results obtained during the development of Tasks 16.1 (Redesign and start-up of an existing biofiltration setup) and 16.2 (Evaluation of biofilter´s performance parameters) of the REMOVALS project. The partner involved in this deliverable was Universitat Autònoma de Barcelona (UAB, partner 3). The objectives included in these tasks have been fully accomplished. 3. Analytical Methods

Gas samples were collected from sampling ports and stored using Tedlar® bags prior to their analysis. Toluene concentration was measured by gas chromatography in a GC 6890N (Agilent Tech., Spain) equipped with a HP-5 capillary column and a flame ionization detector (FID). Method development and calibration for a series of model VOCs was performed according to Prado et al. (2002). Leachate samples were periodically collected from the bottom of each reactor. Conductivity and pH of these samples were measured respectively by means of microCM 2100 and MicropH 2001lab probes (Crison, Spain). Nitrite, nitrate and phosphate of liquid samples were determined by capillary electrophoresis in a Quanta 4000E unit (Waters, Spain). Ammonia concentration was measured by means of a continuous flow analyzer (Baeza et al., 1999). Packing materials porosity and specific surface area were determined in an external laboratory (Serveis Científico-Tècnics, UB) by BET absorption isotherms in an ASAP2000 porosimeter (Micromeritics, U.S.A.). Materials densities were measured in a helium picnometer. Elementary analysis was performed by combustion under standard conditions using sulphanilamide as standard (EA-1108

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ThermoFisher Scientific). Extractable phosphorus was determined by ICP in a Thermo Jarell-Ash model 61E multichannel analyser in standard conditions (Polyscan, U.S.A.). The specific surface area and material densities were determined by the BET technique in a Tristar 3000 apparatus (Micromeritics, U.S.A.). Filter media moisture, organic matter content and pH were determined according to Standard Methods (APHA, 1980). Water holding capacity was measured by keeping the material wet (sparkling constantly tap water for 100 min) 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. 4. Results

4.1. Redesign of an existing biofiltration setup

A biofiltration setup that had been previously employed for the depuration of toluene-containing vapours and consisting of two parallel sets of two in-series biofilters was redesigned and restructured in order to fit the requirements of the REMOVALS project. This plant consisted of four PVC columns (8.8 cm internal diameter) with a packing height of 50 cm, giving a total bed volume of 3 L per reactor (see Figures 1 and 2). Each biofilter was fed with pre-humidified air regulated with calibrated flowmeters as primary gas flow. A secondary air stream was pumped by a peristaltic pump into a glass bubbler unit containing pure liquid toluene. Both gaseous streams were mixed and the resulting gas mixture was fed from the bottom of the reactor. Each reactor had four gas sampling ports located, respectively, at 0, 15, 30, and 45 cm from the gas inlet. Tap water or a nutrient-enriched solution was manually sprinkled on a daily basis over the biofilter beds to provide additional moisture, the necessary nutrients for the microorganisms and to wash dead cells and the end-products of toluene degradation. The leachate was manually collected at the bottom section. Pressure drop along the bed in each reactor was measured by a glass U-tube manometer.

Once the operation of this plant was stopped, in October 2006, it was adapted at no cost to the REMOVALS project specifications. This process, started the first week of November 2006, took approximately 14 weeks. The main modifications performed were:

- Substitution of the gas bubblers by a PC-controlled Micro-burette.

- Inclusion of four PC-controlled electrovalves (one per reactor), aimed at regulating and automating the water and nutrients supply.

- Addition of an on-line CO2 analyzer.

- Integration of an on-line temperature and air moisture content sensor.

- Inclusion of an on-line pressure drop sensor.

All of these stages required considerable repiping, mainly to allow for inlet and outlet gas analyses in all four reactors. As expected, most of the work done during this stage was aimed at developing the software employed for controlling the electrovalves and the PC-controlled Micro-burette and for acquiring and storing data from the sensors. All the software needed was programmed in Visual Basic 6.

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On the other hand, the humidification columns, manometers and flowmeters remained unaltered with respect to the original system. Figures 1 and 2 show, respectively, a picture and a schematic of the experimental setup.

Figure 1. Picture of the experimental bioreactors setup.

Figure 2. Schematic of the biofiltration system (not to scale). (1): Air inlet; (2): humidification column (x 2); (3): peristaltic pump for ammonia supply; (4): microburette for VOCs supply; (5): mixing chamber; (6): flowmeters;

(7): biofilters; (8): PC-controlled electrovalves; (9): adsorption chamber; (10): air outlet; (11): PC-controlled watering pump; (12): water reservoir.

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After the conversion, each biofilter had a volume of 3 L and included, as before, four gas sampling ports located, respectively, at 0, 15, 30, and 45 cm from the gas inlet. The biofilters were fed with compressed dry air. In order to increase the moisture content of the stream, a humidification column was installed in the gas line, upstream of the reactors. The gas pressure and flow rate were set by means of manometers and flowmeters, respectively. Pure ammonia and hydrogen sulfide, when added, were pumped to the main air stream from 8.1 L TedlarTM bags by means of a peristaltic pump (VC-MS/CA8-6, Ismatec S.A., Spain). Pure toluene or a mixture of VOCs were supplied by means of a microburette (Multiburette 2S, Crison Inst. S.A., Spain). When a mixture of VOCs was fed, it was assured that the same mass of each VOC was supplied to the main air stream. The resulting gas stream was split in four and fed to each biofilter from the bottom side. 100 mL of tap water or a nutrient solution were sprinkled every 12 h over the biofilter beds by means of a peristaltic pump (Primus 208-18, Alldos GmbH, Germany) and a series of electrovalves, in order to provide additional moisture and to wash dead cells and metabolic side-products of ammonia and VOCs degradation. The nutrients solution, adjusted to pH 6.5, contained KH2PO4 (1 g·L-1), K2HPO4 (1 g·L-1), NH4Cl (1 g·L-1), NaCl (1 g·L-1), MgSO4 (0.2 g·L-1), CaCl2 (0.02 g·L-

1) and trace elements (1 mL·L-1). The control of the pump and the electrovalves was carried out by means of a PC, using a home-made Visual BasicTM application. 4.2. Evaluation of organic materials as packing bed in bioreactors

4.2.1. Physico-chemical characterization of organic packing materials A complete characterization of the packing materials used in the present study was performed prior to their first inoculation. In addition, the four biofilters were characterized for typical parameters such as bed porosity or water content before inoculation. Table 1 shows the results obtained during the characterization step. It must be stressed that the pore size, the specific surface area, the material density and the C, H, N, S and P content are inherent to the material and, therefore, comparable to other materials characterized in the literature (Bohn, 1996). It was shown that the pore size showed a high correlation with the toluene adsorption capacity determined by breakthrough curves performed in the biofilters (data not shown). Adsorption tests performed at an inlet toluene load of 100 g·m-3·h-1 showed that the bed was saturated after 6, 15, 25 and 34 minutes for the coconut fibre, peat, pine leaves and compost biofilters, respectively. In any case, it is important to mention that high surface areas, with deep pores as these exhibited by compost, may be counteracted due to the biomass growth over the surface of the packing material, thus reducing the sorption capabilities to buffer transient inlet loads. Also, highly mineralized materials such as compost and peat showed low organic matter contents compared to other usual biofilter packing materials, which may lead to a higher resistance to long-term deterioration. Coupled with higher nitrogen content, compost seemed a priori the most suitable material out of the materials tested in terms of physical and chemical characteristics.

Table 1. Initial characteristics of the biofilters and of the organic packing materials used in this study.

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Parameter Coconut fiber Compost Peat Pine leaves Water content (%) 78 ± 2 46 ± 0 73 ± 1 68 ± 3 Organic matter (% dry weight) 81 ± 2 38 ± 5 59 ± 5 87 ± 1 Wet bed density (g·L-1) 430 750 660 310 Bed porosity (-) 0.55 0.43 0.51 0.71 C (% dry weight) 47.32 ± 0.12 28.65 ± 1.51 31.62 ± 5.62 46.42 ± 0.62 H (% dry weight) 5.69 ± 0.12 3.29 ± 0.21 3.38 ± 0.59 5.32 ± 0.05 N (% dry weight) 0.52 ± 0.01 2.87 ± 0.33 1.17 ± 0.13 0.57 ± 0.01 S (% dry weight) Not detected 0.52 ± 0.01 0.10 ± 0.01 0.11 ± 0.01 P (% dry weight) 0.23 Not analyzed 0.05 0.02 Material pore size (Å) 109 ± 1 213 ± 1 175 ± 1 205 ± 2 Specific surface area (m2·g-1) 0.75 ± 0.10 5.12 ± 0.10 1.21 ± 0.02 0.23 ± 0.01 Material density (g·cm-3) 2.02 ± 0.01 1.79 ± 0.01 1.46 ± 0.01 1.28 ± 0.01

Other parameters shown in Table 1 depend upon the operating conditions and on how the reactor is packed. The pine leaves biofilter was packed with a much lower amount of material than the other biofilters. This led to higher bed porosity and a notably lower wet bed density, which implies a lower pressure drop across the bed during normal operation and better water trickling across the bed during watering periods. 4.2.2. Evaluation of biofilters´ performance parameters A previous study performed under similar conditions as the ones expected to be employed in the current study showed that unacceptably long start-up periods could happen when the biofilters were inoculated with sludge from an urban wastewater treatment plant (Maestre et al., 2007). In order to avoid this, an innovative start-up strategy was applied. In this study, instead of inoculating the four biofilters with sludge, they were packed with filter bed materials that had been previously employed in another study and that were, accordingly, already colonized by toluene-degrading strains. As, at the time of the beginning of the experiment, no activated carbon was available, it was decided that, for a first stage, four different organic packing materials would be employed. The materials selected were pine leaves, a mixture of compost and pine leaves in a proportion 3:1 (v:v), coconut fibre and peat. 4.2.2.1. Effect of the EBRT Previous works have highlighted the influence of the EBRT on the performance of VOC-degrading bioreactors (Deshusses and Hamer, 1993; Zhou et al., 1998). According to the literature, it can be hypothesised that bioreactors packed with different materials can show different responses to EBRT variations. In order to study this for the materials selected in this study, four conventional biofilters packed with the four abovementioned organic materials were operated at EBRT values ranging from 90 to 5 s, corresponding to inlet air flow rates from 120 to 2000 L/h.

Although few studies have been published regarding the biodegradation of VOCs at low EBRT (Popov et al., 2005), no previous works have been found in the literature on the removal of toluene in bioreactors at EBRT values as low as 5 s. All the filter bed materials present in the new reactors had been working in previous systems under steady-state conditions for more than eight months, reason why no pre-adaptation phase was observed. A toluene inlet concentration of 1.1 ± 0.3 g/m3 was set during all the experiment. The reactors were periodically fed 100 mL of the abovementioned nutrient solution twice per day. Each EBRT change was performed after steady-state conditions had been reached and kept for at least three hours to ensure consistent results. It was observed that a fungal population remained dominant inside the reactor during all the study.

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Figure 3 shows that a decrease in the removal efficiency was found in all four reactors as the EBRT decreased, even though slight differences in their performance could be observed. However, removal efficiencies above 75 % were found in all systems when operated at EBRT values above 60 s. In the reactor packed with pine leaves (Figure 3a), only a small effect of the EBRT on the removal efficiency was found when operating at EBRT values between 20 and 90 s. Under such conditions, the removal efficiency remained between 75 and 85 %. However, at EBRT values as low as 5 s the removal efficiency dropped to 35 %. The biofilter packed with a mixture of compost and pine leaves (Figure 3b) decreased its performance from 77 % at an EBRT of 60 s to below 20% at an EBRT of 5 s. Both biofilters packed with coconut fibre and peat (Figures 3c and 3d, respectively) showed a similar behaviour to this of pine leaves biofilter. It is remarkable that both of them showed efficiencies around 30 % at EBRT values of 5 s. Except for the reactor packed with compost plus pine leaves, all other reactors showed removal efficiencies between 30 and 35 % at EBRTs as low as 5 seconds, showing that the elimination of toluene at EBRT values around 5 s can be performed with moderate efficiencies. In any case, long-term testing should be undertaken in order to assure the stability of the treatment.

Figure 3. Influence of the EBRT on the toluene removal efficiency of four biofilters packed with different organic materials. (a) pine leaves biofilter; (b) compost + pine leaves biofilter; (c) coconut fibre biofilter; (d) peat biofilter.

4.2.2.2. Influence of nutrients supply

The supply of a nutrient solution plays an important role in the biofilter performance, providing the adequate moisture level and the nutrient content needed by the microorganisms. Although it has been stated that organic packing materials usually offer the necessary nutrients to maintain a proper microbial population, an extra-nutrient addition has been demonstrated to be needed when high pollutant loads are being fed (Devinny et al., 1999). Also, this addition is a key factor for pH control in bioreactors, providing the necessary wash out of by-products.

In the present study, the reactor packed with coconut fibre was operated during 100 days under different watering conditions in order to determine the effect of the addition of a nutrient solution on its

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performance. The EBRT was kept at a constant value of 60 s during all the study. Initially, the reactor was fed a nutrient solution under the conditions described in section 3.1. The toluene load was set at 106.8 ± 14.2 g/m3h during the first 25 days operation. An elimination capacity of 37.3 ± 11.1 g/m3h was found during this stage.

As Figure 4 shows, the removal efficiency did not overcome 50 %. On day 26 of operation, the toluene load was decreased to a value of 28.4 ± 5.7 g/m3h with the purpose of increasing the removal efficiency prior to the nutrient limitation stage. Immediately, an increase in the removal efficiency was observed, up to values around 90 %. pH remained slightly below 2 during this stage.

Once a steady-state had been reached on day 39, the nutrient solution was replaced by distilled water. Thus, the biofilter did not receive any external nutrients between days 39 and 88. The toluene load during this stage was 28.0 ± 7.1 g/m3h. Figure 4 shows that the toluene removal efficiency remained above 90 % during more than one month, dropping sharply after day 80. A progressive increase of the leachate pH was observed, which was related to a decrease in the production of acidic metabolites. Even though an increase in pH could lead to the appearance of bacteria overgrowing the fungal population, which, in turn, would decrease the performance of the biofilter, in the present case the biofilter was predominantly fungal-based during all the study.

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Figure 4. Performance of the biofilter under different toluene inlet loads and nutrient supply conditions. The dotted lines point the days when the liquid solution supplied was switched. TL: Toluene load. EC: Elimination capacity.

RE: Removal efficiency.

A previous work performed under similar conditions (Maestre et al., 2007) proved that the performance of a series of fungal bioreactors packed with organic media were very stable at pH values up to 4-5. Hence, it can be concluded that the decrease in the removal efficiency was not due to the pH increase. Subsequently, in order to assure that it was due to a lack of nutrients, on day 88 the nutrient solution feeding was restored. Figure 3 shows that this led to an immediate increase in the removal efficiency, which reached values comparable to the original ones just after a few days. A concomitant pH drop was then observed.

The results presented herein point out that the operation of toluene-degrading biofilters colonized with fungi may be carried out temporarily without any nutrient addition, even though a N-containing nutrient solution must be added when no other nitrogen source is available to keep reactor performance.

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4.2.2.3. Effect of the watering rate As biodegradation processes require of high moisture content in the filter bed material, it is a common practice in bioreactors to supply water in an intermittent manner. This leads to an increase in the pressure drop of the bioreactor. In the present section, an assessment of the pressure drop variations as a function of different operational parameters is given. In order to determine this, a specific lab-scale setup (Figure 5) was constructed and operated. The reactor employed for the development of the study consists in a PVC column with an internal diameter of 4.6 cm and a height of 70 cm. A humidification column sited upstream the reactor supplies the desired amount of moisture. Inlet air pressure and air flow rate are controlled by means of a manometer and a flowmeter, respectively. The air current is supplied in an upflow mode. Tap water is supplied from the top of the reactor, pumped by a peristaltic pump. An optical level sensor is employed to determine the amount of water inside the reactor. Pressure drop is determined by means of two digital pressure sensors (Testo 512-20hPa and Testo 506-200hPa, Testo S.A., Spain). The pressure was determined at different air flow rates, moisture content values and porosities for eight different filter bed materials. These were compost, coconut fibre, pine leaves, peat, polyurethane foam, immature coal, an innovative material composed of an inner core of clay externally covered by compost and volcanic rock.

Figure 5. Schematic of the reactor constructed to evaluate the variations in the pressure drop. 1: humidification

column; 2: fixed bed (for pressure drop studies); 3: fixed bed (for water retentivity studies); 4: flowmeters; 5: manometer; 6: peristaltic pump; 7: digital pressure drop sensor.

The study was performed at seven different air flow rates, five moisture content values and three different material porosities. The air flow rates selected were so as to achieve EBRT values between 5 and 40 s, considered typical in actual bioreactors. Figure 6 shows the results obtained, respectively, for coconut fibre, compost, polyurethane foam and the mixture of clay and compost. It can be seen that, in all four cases, the pressure drop increases when the air flow rate or the water content increase as well or when the porosity decreases. Volcanic rock was the material that led to highest pressure drop values, reaching up to 50 cm water/m column at an air flow rate of 700 L/h, a water content of 8 % and a porosity of 0.65. Conversely, the organic materials led to lower maximum pressure drop values.

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Pressure drops in a fixed bed have been described by different semi-empirical mathematic expressions. Most studies use Ergun´s equation (Ergun, 1952), shown in Eq. 1, to describe the pressure drop.

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εερ

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In Eq. 1, ΔP is the pressure drop in Pa; H is the filter bed height in m; μ is the air viscosity in Pa·s; ν0 is the superficial air velocity in m·s-1; ε is the filter bed porosity; dφ is the spherical equivalent diameter of the particle in m and a and b are, respectively, the first and the second parameters of Ergun´s equation. These parameters are related to the friction factor. Some authors have satisfactorily adjusted experimental data to Ergun´s equation by adapting these parameters by means of a correction factor (Delhoménie et al. 2003). Some other authors have developed specific relationships, due to the heterogeneity of the material and to the difficulty inherent to the modelling of the pressure drop using Ergun´s equation (Comiti and Renaud, 1989).

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Figure 6. Influence of the operating parameters on the pressure drop in a bioreactor packed with (a) coconut fibre; (b) compost; (c) polyurethane foam; (d) clay and compost aggregates.

In the present work, parameters a and b from Ergun´s equation have been adjusted as a function of the material, the porosity and the water content. The pressure drop (∆P/H) is adjusted by linear regression as a function of the gas superficial velocity. In all cases, the correlation coefficient was above 0.990,

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showing a clear linear trend between the operational parameters and the pressure drop. The experimental study incorporates the effect of the water content in the pressure drop, even though this parameter is not present in Ergun´s equation. Table 2 shows the final results of the systematic study aimed at comparing the effect of the water content in each material. The water content of the compost, lava rock and the aggregate of compost and clay shows the most intense effect on parameter a, while the effect of the water content on parameter b is significantly lower. Accordingly, this allows establishing a modified version of Ergun´s equation which incorporates the effect of the water content on the pressure drop for different packing bed materials.

Table 2. Ergun´s equation parameters as a function of the water content in the biofilters.

4.3. Assessment of sludge-based carbon as packing bed in a biofilter

4.3.1. Physico-chemical characterization of commercial activated carbon

In order to complement the physico-chemical characterization performed on the organic packing materials, a preliminary characterization of a number of inert packing materials, including commercial (supplied by Chemviron Carbon Ltd., U.K) and sludge-based (supplied by Université de Nantes, France) carbon was performed. Table 3 summarizes the results obtained during the physico-chemical analyses, which show that the sludge-based carbon has a significantly higher buffer capacity, conductivity and pH than the commercial activated carbon. All the packing materials studied showed a pH commonly close to neutral. Interestingly, the sludge-based carbon showed the highest pH, with an average value above 8.

Table 3. Some of the main physico-chemical parameters determined for a series of inert packing materials, including commercial and wastewater treatment plant sludge-based GAC.

ε a ord. a slope b ord. b slope 0.70 12.634 -0.069 0.250 0.003 0.76 132.370 0.720 0.595 0.009 0.79 432.090 4.077 1.199 0.004 0.94 0.626 0.004 0.062 0.001 0.96 2.145 0.019 0.090 0.000 0.99 9.130 0.750 0.333 -0.011 0.73 75.398 0.340 0.512 0.005 0.76 115.150 0.384 0.531 0.001 0.77 234.900 2.276 0.766 0.002 0.58 14.618 0.212 0.176 0.001 0.63 35.182 0.387 0.248 0.002 0.64 100.830 0.270 0.466 0.007 0.91 1.885 0.014 1.146 0.058 0.92 2.524 0.016 2.517 0.144 0.96 9.378 0.059 11.196 0.000

Clay/compost aggr. 0.65 12.342 1.458 0.160 0.009

Lava rock

Lignite

Pine leaves

Compost

Coconut fibre

Surfac ar a e2 e

-1 (m ·g ) Humidity (%)Water holding

capacity (g·g -1 )Water retentivity

(% day -1 )Conductivit y

(μS·cm-1) pH Buffer capacity (mL SO 4

2- ·L -1 )

Coconut fiber 1,68 6,62 3,90 192,24 315 5,93 33

Pine leaves 0,50 7,79 1,51 422,78 216 6,90 120

Peat with heather 1,43 6,97 1,80 66,38 338 5,13 20

Compost 2,82 7,83 0,68 57,89 470 7,24 128

Clay/compost mixture 0,76 37,62 0,58 41,90 226 5,72 13

Lava rock 0,62 0,06 0,18 23,33 33 7,21 33

Commercial carbon 950 4,81 0,39 - 40 6,90 43

Sludge-based carbon 85,6 0,90 0,34 - 87 8,20 78

Lignite 5,99 4,85 0,28 41,62 205 6,51 45

Polyurethane foam 0,02 - - 416,45 - - -


This is probably a consequence of the presence of alkaline forms of phosphorus and nitrogen (Table 4). The buffer capacity of the packing materials was always below 130 mL SO4

-2·L-1. The buffer capacity of the sludge-based carbon was around twice the value of the commercial activated carbon. This makes the sludge-based carbon more interesting as support in bioreactors, being able to act as a pH controller in a more efficient way, and reducing, accordingly, possible negative effects of sudden pH variations on the microbial population. The conductivity of the leachate extracted from the materials was very variable, ranging from 33 μS·cm-1 (lava rock) to 226 μS·cm-1 (mixture compost/clay). Again, the conductivity of the sludge-based carbon was significantly higher than that of commercial carbon, as a consequence of its higher content in soluble forms of nitrogen and phosphorus. The water holding capacity showed only slight differences, with values always below 0.6 g(water)·g-1(material) for all the materials considered. It is worth noticing the surface area found for the commercial carbon (Table 3) compared to the sludge-based carbon and other packing materials. GAC materials may be considered good adsorbents compared to other packing materials in Table 3, even if a larger adsorption capacity is expected for the Chemviron carbon compared with the sludge-based carbon. As the elemental composition of the sludge-based activated carbon was expected not to differ greatly from that of sludge-based carbon, these specific assays were not undertaken on this material.

Table 4. Elemental composition of the packing materials considered in the study.

Nitrogen (%) Carbon (%) Hydrogen (%) Sulphur (%)Phosphorus

(μg·g -1 )Organic matter

(%)Coconut fiber 1,17 45,05 6,18 0,12 256 91,62

Pine leaves 0,56 45,18 6,10 0,05 191 86,71

Peat 1,26 21,99 2,56 0,15 455 66,23

Compost 2,68 33,86 4,63 0,63 14487 53,56

Clay/compost mixture 0,34 2,45 0,18 0,19 1259 2,57

Lava rock 0,00 0,40 0,00 0,00 1821 0,63

Commercial carbon 0,47 85,81 0,62 0,26 400 85,58

Sludge-based carbon 3,39 38,54 0,49 0,00 66000 45,44

Lignite 0,85 44,37 4,06 8,81 98 79,69

Additionally, sorption capacity was determined for both dry and wet materials in order to obtain information regarding the interactions between the contaminant, the packing materials and the aqueous phase. Toluene adsorption on packing materials was evaluated by means of a frontal analysis based on toluene measurements at the inlet and outlet of a fixed-bed according to the staircase method (Vente et al., 2005). Isotherms are determined from the breakthrough times of step changes in the feed concentration. Adsorption capacities were evaluated for toluene on compost and SBC. The experimental quantities of the toluene adsorbed are presented in Figure 7 for compost, SBC and SBAC, respectively. In all cases, the behaviours of the materials are analysed in dry and wet conditions. Adsorption capacities of dry materials are evaluated to describe the behaviour in the non-colonized patches in an operated biofilter or to characterize the use of the materials as a buffer to adsorb intermittent pollutant loads using the material previously to the inlet of a biofilter. Adsorption capacities of wet material describe the ability to absorb intermittent pollutant loads when the media supports are used in steady conditions. In Figure 7, experimental quantities of toluene adsorbed on all three porous materials at

13


22ºC in dry and wet conditions are fitted to the well-known isotherms of Freundlich and Langmuir. Characteristic parameters of the mentioned equations were obtained by non-linear regression of the experimental data. Freundlich isotherms are obtained assuming that the energy of interaction between pollutant and adsorbent material is distributed on the surface, grouping into one patch those having the same adsorption energy. On each patch, a pollutant molecule only adsorbs onto only one adsorption site of the material (Do, 1998). On the contrary, the Langmuir model assumes that the surface is homogenous, thus, the energy of interaction between pollutant and adsorbent material is constant over all sites. The adsorption on surface is localised, so each site can accommodate only one molecule (Ruthven, 1984). Freundlich equation (Eq. 2), popularly used in gas phase systems having heterogeneous surfaces, is one of the earliest empirical equations used to describe equilibria data. Langmuir equation (Eq. 3) is the simplest theoretical model for monolayer adsorption.

ninf CKq ·= Eq. 2

in

in

CKCKq+

=2

1· Eq. 3

14

Inlet toluene concentration [mg·m-3]

0 1000 2000 3000 4000 5000 6000

Adso

rbed

tolu

ene

[mg·

g-1 m

ater

ial]

0,0

0,2

0,4

0,6

0,8

Dry (experimental)Wet (experimetnal)Freunlich isotherm Langmuir isotherm

where q is the concentration of the adsorbed specie in the material in mg·g-1 of bed; Kf is the first Freundlich parameter in m3·g-1; Cin is the inlet concentration of the specie in the gas phase in mg·m-3; n is the second Freundlich parameter; K1 is the maximum adsorbed capacity of the material in mg·g-1 and K2 is the inverse adsorption equilibrium constant in mg·m-3.

0,0

0,2

0,4

0,6

0,8

Inlet toluene concentration [mg·m-3]

Dry (experimental)

Wet (experimetnal)Freunlich isotherm

Ads

orbe

d to

luen

e [m

g·g-1

mat

eria

l]

Langmuir isotherm

0 1000 2000 3000 4000 5000

0

50

100

150

200

0 1000 3000 4000 5000

Inlet oluene concentration (mg·m-3)

adso

rbed

tolu

ene

(mg·

g-1 m

ater

ial

■ Dry (experimental) ♦ Wet (experimental)

)

2000

t


Figure 7. Adsorption isotherms (experimental, Langmuir, and Freundlich) of toluene on compost (a), SBC (b) and SBAC (c).

In the interval of toluene concentration studied, the highest adsorption capacity corresponds to the SBAC, being two orders of magnitude higher that those of SBC and compost. Additionally, the adsorption capacities of SBC are higher than the quantity of pollutant adsorbed on compost, being twice at high gas-phase concentrations and increasing the capacities up to six times in low concentrations. The higher adsorption capacity of SBC is in part explained by the high surface area detected in the characterization of materials in comparison to compost (Table 3). Figure 7 shows that the adsorption capacities for all materials decrease drastically when materials are wet. The water film on materials represents a high resistance for the mass transfer of a hydrophobic compound as toluene. Thus, pollutant concentrations on the liquid-solid interface of wet materials are lower than the concentrations on the gas-solid interface in dry conditions. As liquid-phase diffusion is much slower than gas-phase diffusion, the toluene is hardly adsorbed at short contact times. Moreover, previous works affirm that water competes for adsorptive sites when a biofilter is put into operation (Loy et al., 1997). Present results show that the advantages of materials with high adsorption capacity in biofiltration disappear when materials are completely wet. However, when using a separate carbon column placed before a biofilter, the presence of water is prevented and a stable pollutant concentration is buffered to degrade in the Results shown in Figure 7 are in concordance with the thermodynamic and kinetic interpretation of estimated parameters for Freundlich equations and Langmuir model, which are presented in Table 5. Regarding Freundlich parameters, estimated values of n are lower in SBC than in compost, that is, the adsorption isotherm behaviour deviates further away from the linear isotherm, so it is approaching a rectangular isotherm or irreversible isotherm. The concentration needs to go down to an extremely low value before adsorbate molecules desorb from the surface. The value of this parameter, related to the affinity between pollutant and material, is similar in each material independently of the water presence.

Table 5. Parameters of the Freundlich and Langmuir isotherm for compost, SBC and SBAC. Compost SBC SBAC Dry Wet Dry Wet Dry Wet

Kf (m3·g-1) 0.0006 0.000006 0.0514 0.0161 29.0393 0.0507 n 0.83 1.11 0.33 0.31 0.2177 0.8324 Freundlich

fval 0.0369 0.0108 0.0227 0.0216 20.4942 10.6931 K1 (mg·g-1) 2.10 0.19 0.83 0.24 187.04 146.3 K2 (mg·m-3) 10000 10303 655 659 359.93 797.16 Langmuir

fval 0.0450 0.0183 0.0616 0.0008 7.1427 9.5277 Results presented in Table 5 for Langmuir data fitting show that the affinity between pollutant and material (K2 value) is more important in SBC and SBAC than in compost. This affinity is independent on whether the material is wet or dry in the case of SBC and compost. However, the effect of the activation in SBAC is less pronounced in terms of adsorption capacity when the material is wet. Regarding the value of K1, which is related to the maximum adsorption capacity of the material, it is much higher in SBAC that in compost and SBC. Also, the operational parameters influencing the pressure drop were characterized for SBC and compost (as it was expected that the pressure drop generated in SBAC would have great resemblance to that in SBC). Considering that the adsorption capacities of pre-selected materials are similar under the typical operating conditions of biofilters, i.e. with a certain amount of water in the biofilter, the pressure drop analysis of SBC in comparison to compost was selected as the criteria to determine if materials were

15


suitable to pack a biofilter. The comparative pressure drop study was performed as a function of the main operational parameters that generally influence the process. Gas superficial velocity was increased up to 350 m·h-1 according to common operation conditions in biofilters (Devinny et al., 1999). The water content measured in each support material was achieved by continuously watering the material at different water velocities avoiding flooding episodes according to their own properties. Different porosities were obtained for compost, by packing the bed at different particle sizes. A porosity of 0.93 was considered for SBC. Results were represented in surface plots to observe simultaneously the influence of all the parameters on the pressure drop (Figure 8). Considering a wide range of conditions where materials are operative in a biofilter, pressure drop measured are sensitively higher in SBC than in compost, what is in concordance with the shape and the specific surface area of materials. Results shown in Figure 8 demonstrate that the pressure drop for SBC is more influenced by the gas flow rate than for compost, even though the porosity of the bed for SBC is higher than for compost. Similarly, the water content influences SBC to a larger extent, even though the possible range of water content is also higher because, in compost, the bed is flooded at water contents superior to 25 % in bed volume due to the lower porosity of the bed. The strong dependence on water content is more important at high flow rates, where the resistance of the flow towards the bed increases. Regarding the porosity effect in compost, results do not show significant differences in the interval tested though three different sizes of particles were tested to achieve a wide range of operation.

0

100

200

300

400

0

5

10

15

20

25

050

100150

200250

300350

Pres

sure

dro

p (m

mH

2O·m

-1)

Wat

er co

nten

t (%

)

Superficial velocity (m·h -1)

0 100 200 300 400

0

100

200

300

400

010

2030

4050

6070

80

050

100150

200250

300350

Pres

sure

dro

p (m

mH

20·m

-1)

Wat

er co

nten

t (%

)

Superficial velocity (m·h -1)

0 100 200 300 400

ε = 0.93

ε = 0.80

ε = 0.76

ε = 0.70

Figure 8. Influence of operational parameters on pressure drop for compost (a) and SBC (b).

4.3.2. Evaluation of sludge-based carbon as filter bed material in a biofilter

After proving that SBC (and, accordingly, SBAC) shows an acceptable pressure drop in a wide range of operation conditions, the behaviour of this material as packing bed was assessed in a biofilter treating a mixture of six organic volatile compounds (VOCs) such as hexanal, butyric acid, limonene, methylisobutylketone (MIBK), dimethylsulfide (DMS) and toluene. The experimental time was divided in 2 phases. In the first phase, the inlet concentration was 5.8 ± 1.9 ppmv per VOC, corresponding an inlet loading rate of 3.2 ± 1.0 g m-3h-1, while in the second phase the VOCs inlet concentration was doubled to 10.1 ± 1.0 ppmv, corresponding to an inlet loading rate of 5.6 ± 0.6 g m-3h-1. The biofilter was packed with a mixture of coconut fiber plus sludge-based carbon from a waste water treatment plant (1:1, v:v), supplied by Partner 14 (Imperial College of London). The coconut fiber was added as bulking agent

16


in order to decrease the pressure drop in the biofilter (McNevin and Barford, 2001; Burgess et al, 2001) and aided to held water content. Growth and immobilization of biomass on SBC were evaluated inoculating the lab-scale plant with activated sludge and studying the abatement of VOCs in the reactor, during 95 days of operation (Figure 9). The gas flow rate was kept at 425 l h-1 (EBRT of 25 s). The bioreactor was sprinkled once a day with tap water.

17


t (d)0 15 30 45 60 75 90

Tolu

ene

Con

cent

ratio

n (p

pmv)

0

5

10

15

20

RE

(%)

0

25

50

75

100

Cin (ppmv) Cout (ppmv) RE (%)

Phase I Phase II Figure 9. Evolution of toluene inlet and outlet concentration and toluene removal efficiency (RE) in SBC-coconut

fiber biofilter.

In Figure 9, the operation and performance of the biofilter in terms of toluene inlet and outlet concentration, and toluene removal efficiency (RE) is represented. The first 10 days correspond to the biofilters start-up phase, in which SBC was basically adsorbing toluene and the immobilized biomass was being acclimatized. Toluene removal efficiencies above 90% were achieved from the first day, corresponding to an EC between 2.5 and 3 g m-3h-1. Results demonstrate that SBC was able to develop and maintain an active biomass on the surface of the material after a short time of acclimatization, due especially the high surface area and the high water holding capacity. In the first phase, while the biomass was adapting to the operation conditions of the reactor, toluene was being adsorbed on the material,. Moreover, the presence of nutrients content in the SBC (Table 6) could contribute to the growth and the maintenance of the biomass immobilized. On day 69, inlet toluene concentration was increased from 5 to 10 ppmv. The RE decreased gradually, down to values under 75%. This shows a relaively poor performance of this packing material (Devinny et al, 1999) in comparison to other studies where roughlier and more hydrophilic packing materials allowed the microorganism to fix more easily (Durham et al, 1994). Nevertheless, as soon as the inlet toluene concentration was decreased again to 5 ppmv, on day 88 of operation, high RE were reached again. Regarding toluene abatement, the RE reported in the analyzed operation time shows that the mixture of SBC-coconut fibre is an acceptable support media for the biomass, but when feeding toluene concentrations above 10 ppmv the removal efficiency decreases drastically.

Table 6. Inlet concentration, inlet loading rate, removal efficiencies and elimination capacities for each VOC and

for each phase.

Parameter Toluene DMS MIBK Butyric Acid

Limonene Hexanal

Cin (ppmv) 5.8 ± 1.9 EE (%) 92.6 ± 11.7 75.3 ± 9.2 79.8 ± 17.9 69.2 ± 21.2 72.1 ± 14.0 81.1 ± 16.1

CT (gm-3·h-1) 3.2 ± 1.0

Ph. I

CE (gm-3·h-1) 3.0 ± 0.8 1.8 ± 0.9 3.2 ± 1.6 2.4 ± 1.7 3.6 ± 1.4 3.1 ± 1.4 Cin (ppmv) 10.1 ± 1.0

EE (%) 84.9 ± 11.9 85.5 ± 4.5 81.2 ± 7.8 82.9 ± 7.8 81.3 ± 12.7 82.5 ± 5.0 CT (gm-3·h-1) 5.6 ± 0.6

Ph. II

CE (gm-3·h-1) 4.4 ± 0.8 3.2 ± 0.2 4.9 ± 0.5 4.4 ± 0.4 6.6 ± 1.0 4.9 ± 0.3

18


19

t (d)0 15 30 45 60 75 90

Pres

sure

dro

p (m

m H

2O /

m)

0

5

10

15

20

25

30Phase I Phase II

t (d)0 15 30 45 60 75 90

Cin

, Cou

t CO

2 (p

pmv)

200

300

400

500

600

700

800

900Cin (ppmv) Cout (ppmv)

Phase IPhase II

Inlet concentration, inlet loading rate, removal efficiency and elimination capacity are presented for each phase and for each VOC (Table 6). Removal efficiencies for all six VOCs were maintained above 70 % during the whole experimental time. These results can be considered acceptable, although different studies performed by other authors report similar or better performances when using activated carbons (Abumaizar et al, 1998; Moe and Li, 2005; Mathur et al, 2007). Inlet and outlet CO2 concentration were measured during the performance time (Figure 10). In biofiltration, the production of CO2 is an important parameter that allows evaluating the extent of pollutant degradation (Mathur et al, 2007). According to Figure 10, an increasing trend in the difference between the inlet and outlet concentration can be clearly seen, demonstrating an increasing biomass activity in the biofilter. Biomass growth, probably fungi due to its white colour, was clearly distinguished in the first stage of the biofilter where more pollutant concentration was available.

Figure 10. Evolution of CO2 inlet and outlet concentration in the SBC-coconut fiber biofilter.

The evolution of the bed pressure drop is related to the accumulation of biomass in the biofilter. Up to 2.5 mm H2O/m packing material were measured during the first phase (Figure 11), which can be considered a typical value (Devinny et al, 1999). Increasing the VOCs inlet concentration up to values around 10 ppmv led to a concomitant increase in the pressure drop, which reached values of 26 mm H2O/m bed. No compaction of the filter bed was observed. The high pressure drop values achieved at the end of the performance time were caused by the biological flora growth and by the accumulation of water.

Figure 11. Evolution of the pressure drop in the SBC-coconut fiber biofilter.


4.3.3. Economical assessment of the use of sludge-based carbon as filter bed in a biofilter

With the purpose of assessing the economical viability of using activated carbon as filter bed in bioreactors, a cost-assessment model which includes the main investment, operating and maintenance costs of a conventional biofilter was developed. Specifications on this model can be found in Prado et al. (2009). The total investment costs have been calculated as the sum of the initial site preparation costs, the packing and support materials costs, the equipment costs, the piping costs, the electrical costs, the equipment installation costs, the engineering design costs, the liner costs and the miscellaneous costs. The annual operating costs were calculated as the sum of the electricity and water annual consumption costs, the labor costs and the overhead costs. Among the medium replacement costs, two actions were considered: the disposal of the old packing material and the addition of the new one. All other eventual costs have been considered negligible compared to these. Although a very simple bioreactor was selected as model reactor, the protocol can be applied to different scenarios (as, e. g., biotrickling filters) after few modifications. No nutrient supply was considered for the economical evaluation. The protocol presented herein was applied to a series of biofilters supposedly packed with different packing materials and operating under the same conditions. A gas flow rate of 20 000 m3/h and an EBRT of 60 s were selected for the biofilters, which were considered to be operating for 15 years. The EBRT value of 60 s was selected as previous studies (see section 4.2.2.1) showed that removal efficiencies above 75 % may be assured at that EBRT irrespective of the packing material. Seven packing materials were selected for the study: compost, agricultural residues, a mixture of pig manure and sawdust, BiosorbensTM and three different types of activated carbon (commercial activated carbon, sludge-based activated carbon and sludge-based non-activated carbon). The cost and the estimated durability of each material are shown in Table 7. All costs presented herein are estimations for Spain, year 2008. Due to the inherent difficulty to quantify many of the parameters included in the protocol, an error around ±20 % was estimated for costs calculated in the present section.

Table 7. Average cost and estimated durability of all seven packing materials considered.

Material Cost (€/m3) Estimated durability (years) Compost 25 2

Agricultural residues 75 1 Pig manure + sawdust 50 1

BiosorbensTM 550 15 Commercial activated carbon 475 10 Sludge-based activated carbon 120* 10

Sludge-based non-activated carbon 80* 5 * Values calculated considering a market cost of 600 €/Tm for SBAC and 400 €/Tm for SBC (as provided by Partner

14) and a carbon density of 0.2 Kg/L in both cases. Table 8 summarizes the main costs implied in all seven cases. It can be clearly seen that the use of

sludge-based carbon represents an economically viable alternative to other inert packing materials, as BiosorbensTM or commercial activated carbon. Total annualized costs of 43360 and 42910 €/year have been estimated for SBAC and SBC, respectively, while, in the case of commercial activated carbon, a total annualized cost of 79710 €/year results from the application of the model. This reduction in the costs is a consequence of the much lower purchase price of sludge-based carbons, which, in turn, results of their reduced production costs. The investment cost for this average-sized system ranges from around 56000 € when selecting compost as packing material to around 380000 €, if BiosorbensTM is the chosen bed. Among these, biofilters packed with sludge based carbons fit in the medium cost range (around 115000 and 90000 € for activated and non-activated carbon-based biofilters, respectively). However, the high durability of these materials causes the annualized medium replacement costs to be low in comparison to the others (4200 and 6600 €/year). Only BiosorbensTM, which has an estimated durability

20


of the same range as the biofilter total operating time, reason why no medium replacement costs need to be considered, has lower annualized medium replacement costs. Regarding operating costs, as all the scenarios contemplate same air flow rate and empty bed residence time, no differences are found. Overall, only compost is a better option from an economical point of view. However, as previously stated, the use of organic packing materials as compost as packing bed in biofilters may lead to operational problems which may decrease performance. Accordingly, there is evidence that sludge-based carbon may be a promising alternative to traditional packing materials both from an operational and an economical point of view.

21

Table 8. Investment, operating and medium replacement costs estimated for a series of average-size conventional biofilters packed with different packing materials,

with Q = 20 000 m3/h, Vb = 400 m3, EBRT = 60 s.

BF1: Compost biofilter. BF2: Agricultural residues biofilter. BF3: Pig re + sawdu ilter. BF4 sorb tivated biofilter. BF6: Sludge-based activated carbon biofilter. BF7: Sludge-based non-activated carbon biofilter.

manu st biof : Bio ensTM biofilter. BF5: Commercial ac carbon

Cost BF1 BF2 BF3 BF4 BF5 BF6 BF7 Total investment costs (€) 56 500 87 260 71 890 379 570 333 420 114 950 90340 Annualized investment costs (€/year) 6 600 10 190 8 400 44 350 38 950 13 430 10 550 Annual operating costs (€/year) 19 340 19 340 19 340 19 340 19 340 19 340 19 340 Annualized operating costs (€/year) 25 750 25 750 25 750 25 750 25 750 25 750 25 750 Packing material replacement costs (€/action) 16 640 36 640 26 640 0 196 640 54 640 38 640 Annualized packing material replacement costs (€/year) 9 820 46 460 33 780 0 15 010 4 170 6 600 Total annualized costs (€/year) 42 170 82 400 67 930 70 100 79 710 43 360 42 910

5. Published, publishable and implementable results

Published and accepted papers: - Prado, Ó. J., Gabriel, D., Lafuente, J. 2009. Economical assessment of the design, construction and operation of open-bed biofilters for waste gas treatment. J. Environ. Manage. doi:10.1016/j.jenvman.2009.01.022 - Dorado, A. D., Hernández, J., Ribera, G., Gabriel, D., Lafuente, J., Gamisans X. (2009). Evaluation of sludge-based carbon as packing material in biofiltration in comparison to classic materials. Water Sci. Tech. (accepted manuscript) Manuscripts presented in congresses, meetings and symposiums: Dorado, A. D., Hernández, J., Ribera, G., Gabriel, D., Lafuente, J., Gamisans X. 2008. Evaluation of sludge-based carbon as packing material in biofiltration in comparison to classic materials. Proceedings of the 3rd IWA Odour and VOCs Conference: Measurement, Regulation and Control Techniques. Barcelona, Spain. Papers in preparation: - Dorado, A. D., Hernández, J., Prado, Ó. J., Lafuente, J., Gamisans, X., Gabriel, D. Technical and economical assessment of the use of sludge-based carbon as packing material in bioreactors. Due date: September 2009. - Dorado A.D., Lafuente F., Gabriel D., GamisansX. A comparative study of the most suitable packing materials according to common situations in biofiltration. Due date: June 2009. According to the technical and economical results presented herein, it appears clear that the use of SBC or SBAC as packing bed in bioreactors will result in similar performances and in reduced costs compared to those of bioreactors packed with traditional filter bed materials. Full-scale studies will have to be undertaken in order to assess the robustness of these systems, even though implementation of such materials can be easily carried out at industrial level. 6. Conclusions

The results obtained so far allow concluding that the biofilters packed with classical, organic filter bed materials can be highly stable and effective for the removal of gas-phase toluene. Fungal strains provide a high robustness against pH drops and variations in the inlet load. The effect of the EBRT on the removal of toluene in organic bed-based biofilters has been characterized, as well as the effect of the supply of a nutrient solution at different rates. A series of inert packing materials commonly used in biofiltration have been characterized and compared for a better knowledge of their advantages and drawbacks. The results obtained show that these materials offer a higher contact surface than traditional, organic, packing materials. Biofiltration has proved to be a useful application to dispose biological sludge from a wastewater treatment plant. Coconut fiber-SBC mixture is an appropriate packing material for biomass growth in a conventional biofilter. Toluene removal efficiency was commonly above 90 % when an inlet concentration around 6 ppmv was fed to the bioreactor. Also, it was proved that the use of sludge-based carbon may represent an economically viable alternative to other commonly employed inert packing materials, as commercial activated carbon. Total costs may be reduced from 40 to 50 % if SBAC or SBC is selected as packing material instead of commercial activated carbon. Comparatively to actual materials


used in biofiltration, the economical study evidences that sludge-based carbon are a promising alternative to traditional packing materials. 7. References

Abumaizar, R., Kocher, W., Smith, E. 1998. Biofiltration of BTEX contaminated air streams using compost-activated carbon filter media. J. Hazard. Mater. 60:111-126 Aizpuru, A., Dunat, B., Christen, P., Auria, R., Garcia-Peña, I., Revah, S. 2005. Fungal biofiltration of toluene on ceramic rings. J. Environ. Eng.-ASCE. 131, 396-402. APHA (American Public Health Association), 1980. Standard methods for the examination of water and wastewater. Baeza, J.A., Gabriel D., Lafuente J. 1999. An expert supervisory system for a pilot WWTP, Environ. Modell. Softw. 14, 383–390. Bohn, H.L. 1996. Biofilter media. In: Proceedings of the Air & Waste Management Association, 89th Annual Meeting & Exhibition, Nashville (US). Burgess, J.E., Parsons, S.A., Stuetz, R.M. 2001. Developments in odour control and waste gas treatment biotechnology: a review. Biotechnol. Adv. 19:35-65 Deshusses, M.A., Hamer, G. 1993. The removal of volatile ketone mixtures from air in biofilters. Bioprocess Engineering. 9, 141-146. Devinny, J.S., Deshusses, M.A., Webster, T.S. 1999. Biofiltration for air pollution control. CRC Lewis Publishers. Do, D.D. 1998. Adsorption analysis: equilibria and kinetics. Imperial College Press, UK. Durham, D.R., Marshall, L.C., Miller, J.G., Chmurny, A.B. 1994. Characterization of inorganic biocarriers that moderate system upsets during fixed-film biotreatment processes. Appl. Env. Microbiol. 60:3329-3335. Garcia-Peña, I., Hernández, S., Favela-Torres, E., Auria, R., Revah, S. 2001. Toluene biofiltration by the fungus Scedosporium apiospermum TB1. Biotechnol. Bioeng. 76, 61-69. Kennes, C., Thalasso, F., 1998. Waste gas biotreatment technology. J. Chem. Technol. Biotechnol. 72, 303-319. Kennes, C., Veiga, M.C. 2001. Bioreactors for waste gas treatment. Kluwer Academic Publishers. Kwon, S-H., Cho, D. 2009. A comparative, kinetic study on cork and activated carbon biofilters for VOC degradation. J. Ind. Eng. Chem. doi:10.1016/j.jiec.2008.09.002. Leson, G., Winer, A. M. 1991. Biofiltration - An innovative air-pollution control technology for VOC emissions. J. Air Waste Manage. 41, 1045-1054. Liu, Y. H., Quan, X., Zhao, Y. Z., Chen, S., Zhao, H. M. 2005. Removal of ternary VOCs in air streams at high loads using a compost-based biofilter, Biochem. Eng. J. 23, 85-95.

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Loy, J., Heinrich, K. Egerer, B. 1997. Influence of filter material on the elimination rate in a biotrickling filter bed. Proceedings of the 90th annual meeting exhibition of the air and waste management association. Air and waste management association, USA. Maestre, J.P., Gamisans, X., Gabriel, D., Lafuente, J. 2007. Fungal biofilters for toluene biofiltration: Evaluation of the performance with four packing materials under different operating conditions. Chemosphere 67, 684-692. Mathur, A., Majumder, C.B., Chatterjee, S. 2007. Combined removal of BTEX in air stream by using a mixture of sugar cane bagasse, compost and GAC as biofilter media. J. Haz. Mat.148: 64-74. McNevin, D, Barford, J. 2001. Biofiltration as an odour abatement strategy. Biochem. Eng. J. 5:231-242 Moe, W.M., Li, C. 2005. A design methodology for activated carbon load equalization systems applied to biofilters treating intermittent toluene loading. Chem. Eng. J. 113: 175-185. Oosting, R., Urlings, L.G.C.M., van Riel, P.H., van Driel, C. 1992 Biopur®: alternative packaging for biological systems. In “Biotechniques for air pollution abatement and odour control policies”. Eds. Dragt A.J. and van Ham J. pp. 63-70. Elsevier Science Publishers B. V. Popov, V., Khomenkov, V., Zhukov, V. 2005. Design, construction and long-term performance of novel type of industrial biotrickling filters for VOC and odour control. In “Proceedings of the International Congress Biotechniques for Air Pollution Control”. Eds. Kennes C., Veiga MC. pp. 257-262. La Coruña, Galicia, Spain. Prado, Ó.J., Mendoza, J.A., Veiga, M.C., Kennes, C. 2002. Optimization of nutrient supply in a downflow gas-phase biofilter packed with an inert carrier. Appl. Microbiol. Biotechnol., 59, 567-573. Prado, Ó.J., Veiga, M.C., Kennes, C. 2004. Biofiltration of waste gases containing a mixture of formaldehyde and methanol. Appl. Microbiol. Biotechnol., 65, 235-242. Prado, Ó. J., Gabriel, D., Lafuente, J. 2009. Economical assessment of the design, construction and operation of open-bed biofilters for waste gas treatment. J. Environ. Manage. doi:10.1016/j.jenvman.2009.01.022 Rattanapan, C., Boonsawang, P. Kantachote, D. 2008. Removal of H2S in down-flow GAC biofiltration using sulfide oxidizing bacteria from concentrated latex wastewater. Biores. Technol. 100:125–130. Ruthven, D.M. 1984. Principles of adsorption and adsorption processes. John Wiley & Sons, Inc, USA. van Groenestijn, J. W., Liu, J. X. 2002. Removal of alpha-pinene from gases using biofilters containing fungi. Atmos. Environ. 36, 5501-5508. Vente, J.A., Bosch, H., Haan, A.B. Bussmann P. 2005. Evaluation of sugar sorption isotherm measurement by frontal analysis under industrial processing conditions. Journal of Chromatography A, 1066:71-79. Yani, M., Hiral, M., Shoda, M. 1998. Ammonia gas removal characteristics using biofilter with activated carbon fiber as a carrier. Environ. Technol. 19:709-715.

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Zhou, Q., Huang, Y.L., Tseng, D.H., Shim, H., Yang, S.T. 1998. A trickling fibrous-bed bioreactor for biofiltration of benzene in air. J. Chem. Technol. Biotechnol. 73, 359-368.

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