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Selection of odour removal technologies in wastewater treatment plants: A guideline based on Life Cycle Assessment
Journal of Environmental Management 149 (2015) 77e84
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Journal of Environmental Management journal homepage: www.elsevier.com/locate/jenvman
Selection of odour removal technologies in wastewater treatment plants: A guideline based on Life Cycle Assessment M. Estrada b, c, Raúl Mun ~ oz b, Carolina Alfonsín a, *, Raquel Lebrero b, Jose N.J.R. (Bart) Kraakman d, e, Gumersindo Feijoo a, Mª Teresa Moreira a a
Department of Chemical Engineering, Institute of Technology, University of Santiago de Compostela, 15782 Santiago de Compostela, Galicia, Spain Department of Chemical Engineering and Environmental Technology, Escuela de Ingenierías Industriales, Sede Dr. Mergelina, University of Valladolid, Dr Mergelina s/n, 47011 Valladolid, Spain c School of Engineering, London South Bank University, UK d Department of Biotechnology, Delft University of Technology, Julianalaan 67, 2628 BC Delft, The Netherlands e CH2M Hill, Level 7, 9 Help Street, Chatswood, NSW 2067, Australia b
a r t i c l e i n f o
a b s t r a c t
Article history: Received 22 July 2014 Received in revised form 9 October 2014 Accepted 13 October 2014 Available online
This paper aims at analysing the environmental benefits and impacts associated with the treatment of malodorous emissions from wastewater treatment plants (WWTPs). The life cycle assessment (LCA) methodology was applied to two biological treatments, namely biofilter (BF) and biotrickling filter (BTF), two physical/chemical alternatives, namely activated carbon tower (AC) and chemical scrubber (CS), and a hybrid combination of BTF þ AC. The assessment provided consistent guidelines for technology selection, not only based on removal efficiencies, but also on the environmental impact associated with the treatment of emissions. The results showed that biological alternatives entailed the lowest impacts. On the contrary, the use of chemicals led to the highest impacts for CS. Energy use was the main contributor to the impact related to BF and BTF, whereas the production of glass fibre used as infrastructure material played an important role in BTF impact. Production of NaClO entailed the highest burdens among the chemicals used in CS, representing ~90% of the impact associated to chemicals. The frequent replacement of packing material in AC was responsible for the highest environmental impacts, granular activated carbon (GAC) production and its final disposal representing more than 50% of the impact in most categories. Finally, the assessment of BTF þ AC showed that the hybrid technology is less recommendable than BF and BTF, but friendlier to the environment than physical/chemical treatments. © 2014 Elsevier Ltd. All rights reserved.
Keywords: Activated carbon Biofiltration Chemical scrubbing Life cycle assessment (LCA) Odour abatement Wastewater treatment plants (WWTPs)
1. Introduction Wastewater treatment plants (WWTPs) are considered important sources of gaseous emissions, including green house gases (GHG) and odorants (Shaw and Koh, 2013). The malodorous emissions associated with treatment processes are considered one of the major concerns of exposed population living in surrounding areas of WWTPs. In this context, the concentration of malodours in the air is often monitored and controlled with the aim of complying with odour regulations while keeping a respectable public image of the emission sources.
* Corresponding author. Tel.: þ34 881 816 739. E-mail addresses: [email protected] (C. Alfonsín), [email protected]. ~ oz), es (R. Lebrero), [email protected] (J.M. Estrada), [email protected] (R. Mun [email protected] (N.J.R.(B. Kraakman), [email protected] (G. Feijoo), [email protected] (M.T. Moreira). http://dx.doi.org/10.1016/j.jenvman.2014.10.011 0301-4797/© 2014 Elsevier Ltd. All rights reserved.
Odour abatement technologies have been widely investigated as cost-efficient and reliable alternatives for the mitigation of odour nuisance (Revah and Morgan-Sagastume, 2005; Schlegelmilch et al., 2005). These technologies are commonly classified into physical/chemical and biological techniques. Physical/chemical technologies have been broadly implemented as a consequence of their rapid start-up, low empty bed residence time (EBRT) and consolidated know-how and experience in design and operation. These techniques are often based on the transfer of odorants from the gas emission to either a solid (adsorption) or liquid (absorption) phase. These pollutants can be further transformed into byproducts according to their reactivity with the chemicals used. However, in the last decades biological systems have been increasingly used due to their ability to efficiently treat malodorous emissions at lower operating costs (Schlegelmilch et al., 2005). The main advantages of bioprocesses compared to their physical/ chemical counterparts derive from their low generation of secondary wastes and low demand of resources, such as chemicals or
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adsorbent media. On the other hand, biological processes often require larger EBRTs (2e120 s vs. 1e5 s) and associated footprint than physical/chemical alternatives at similar odour removal efficiencies. Technologies for odour treatment have been widely reviewed in the literature in order to establish their optimal range of application and performance for the removal of volatile organic compounds (VOCs) and volatile inorganic compounds (VICs). Nowadays, there are enough experimental evidences regarding the capability of biofilters and biotrickling filters to achieve significantly high removal efficiencies of air pollutants at both trace levels (Lebrero et al., 2014, 2012) and industrial concentrations (Balasubramanian et al., 2012; Omri et al., 2013). Estrada et al. (2012) demonstrated that biological techniques were the most cost-efficient alternatives with lower sensitivity to design parameters and lower operating costs than physical/chemical treatments at typical odour concentrations emitted in WWTPs. In terms of sustainability, Estrada et al. (2011) assessed the performance of different physical/chemical and biological odour abatement technologies based on the IChemE Metrics methodology (IChemE, 2002). The analysis focused on major environmental indicators such as resource use, waste production and emission impacts as well as process economics and social impact. This preliminary study showed as compared to physical/chemical technologies, biological treatments presented lower demand of energy, material and chemicals and limited production of hazardous wastes. These systems were the most favourable option despite the required high initial investment costs when analysing investment costs and operational costs over 20 years. Although the aforementioned reports offered valuable information for the selection of the most suitable technology for the treatment of malodorous emissions, the IChemE Metrics methodology is only focused on the target process and does not consider a holistic approach for the assessment. For instance, despite water consumption or material use being considered as environmental indicators, the impact of their production processes is out of the scope of the IChemE Metrics analysis. Nevertheless, the origin and management of the components/consumables of the system under study could also be major contributors to the environmental impact. Previous research papers showed how a complete environmental evaluation helps to select the most adequate treatment technology, considering not only removal efficiencies but also potential environmental impacts (Alfonsín et al., 2013; Hospido et al., 2012). Life cycle assessment (LCA) methodology is a quantitative procedure to assess the environmental burdens associated with products, processes and services, which is commonly used for environmental impact evaluation (Baumann and Tillman, 2004; ISO, 2006b). Although this methodology has been widely and satisfactorily applied to technologies dealing with wastewater treatment (Hospido et al., 2004; Rodriguez-Garcia et al., 2011), there is, to the best of our knowledge, only one application of LCA to the abatement of gaseous emissions, focused on lab-scale biofilters (Alfonsín et al., 2013). This study aimed at enriching the guidelines for the selection of the best technological alternative for the treatment of malodorous emissions under an environmental perspective. In this context, a comprehensive LCA was performed in order to determine the overall environmental impacts of five full-scale odour abatement technologies in a WWTP scenario: biofilter (BF), biotrickling filter (BTF), activated carbon filter (AC), chemical scrubber (CS) and a hybrid technology consisting of a BTF coupled with AC (BTF þ AC). The results obtained for each technology were compared to the environmental impact derived from the direct discharge of the malodorous emission without treatment, named non treatment scenario (NT).
2. Materials and methods The ISO 14040 (ISO, 2006a) standard determines four basic stages for LCA studies: goal and scope definition, life cycle inventory (LCI), life cycle impact assessment (LCIA) and finally, interpretation of results (Baumann and Tillman, 2004). Goal and scope definition constitutes the first phase of an LCA and aims at defining the boundaries of the study and the quality of data used. A functional unit (FU), which represents the function of the system under study, must be also established in this phase. Then, LCI is performed, which involves data collection and interpretation of inputs and outputs. The allocation procedure is also conducted during the LCI phase, which consists on distributing input and output flows among the products of the process. LCIA represents the third phase and its purpose is to convert LCI data into potential impacts associated to products and processes. LCIA includes two mandatory steps (i.e. classification and characterization) and other optional elements, such as normalization and weighting. Finally, the interpretation of the results allows identifying the hot spots of the process as well as recommending options to reduce the environmental burdens. 2.1. Goal and scope definition This study was conducted to evaluate the environmental impacts associated with the performance of five of the most commonly implemented odour abatement technologies in WWTPs. The functional unit chosen in this study was 1 m3 of treated air, which is consistent with the approach used in a previous study evaluating lab-scale gas-phase bioreactors (Alfonsín et al., 2013). 2.2. System boundaries The system boundaries determine the units and elements of the process included in the analysis (ISO, 2006a). The assessment here conducted considered the incoming polluted emissions, which are directly discharged to the atmosphere in the NT scenario, the output treated emission, material used in the infrastructure, consumables (e.g. packing material), potable or secondary plant effluent water, energy use, chemicals and transportation and final disposal of all wastes. Water input is required in all systems except in AC and potable water instead of WWTP secondary effluent water is used in CS. The secondary effluent water of the WWTP is considered a product related to wastewater treatment and no impact is attributed when used for irrigating the packing material of the biofilter or the biotrickling filter. The leachate collected in each technology (except in AC) was not considered in the analysis since it is returned to the WWTP headworks with a negligible flow in comparison with the net flow treated in the WWTP. The transportation from the manufacturing industry to the odour treatment facility of consumables like packing materials and chemicals was included. 2.3. Description of odour treatment technologies Five odour abatement technologies were assessed: two biological (BF and BTF), two physical/chemical (CS and AC) and a hybrid technology (BTF þ AC). The model malodorous stream of 50000 m3h1, representing a typical emission from WWTPs, included 30 different VOCs (Zarra et al., 2008), methanethiol (1.97 mg m3) and hydrogen sulphide (20.9 mg m3) (Table 1). The odorants were classified into groups of high, medium and low hydrophobicity, which largely determine their removal efficiencies in the technologies evaluated (Estrada et al., 2011). A typical release of CO2 of 0.75 g CeCO2 (g-Coxidized)1 due to the microbial oxidation of odorants and a biomass yield of 0.25 g C-biomass (g-Coxidized)1 were considered. Water was supplied to the BF, BTF and CS in order
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Table 1 Summary of the inventory data for the non treatment scenario and for the five abatement technologies evaluated. The units are referred to the FU: 1 m3 of treated air. NT
BF
BTF
AC
BTF þ AC
CS
Discharge of air 1-Ethyl-2-methylbenzene (mg) 1,2,4-Trimethylbenzene (mg) 1,3,5-Trimethylbenzene (mg) 2-Butoxyethanol (mg) a-Pinene (mg) Acetic acid (mg) Acetone (mg) Acetophenone (mg) Benzaldehyde (mg) Benzene (mg) Butanone (mg) Butyric acid (mg) D-limonene (mg) Decane (mg) Dimethyl disulfide (mg) Dodecane (mg) Ethylbenzene (mg) Hydrogen sulfide (mg) Limonene (mg) m-Xylene (mg) Methanethiol (mg) Methylcyclohexane (mg) Nonane (mg) o-Xylene (mg) Octane (mg) p-Xylene (mg) Propanoic acid (mg) Tetrachloroethylene (mg) Tetradecane (mg) Toluene (mg) Tridecane (mg) Undecane (mg) Carbon dioxide (mg)
1.31 1.06 8.36 9.63 1.31 1.05 4.62 5.88 6.70 2.19 4.54 2.09 2.30 7.08 2.13 3.41 1.47 20.91 1.15 1.77 1.97 1.26 1.07 1.77 1.07 1.77 1.13 1.07 4.76 5.09 4.04 5.37 0
Infrastructure material Concrete (m3) Glass fibre (kg)
e e
9.33 109 e
e 1.48 105
e 3.23 105
e 2.13 105
e 4.83 105
Packing material Compost (kg) Perlite (kg) Polyurethane foam (kg) Activated carbon (kg) Intalox saddles (kg)
e e e e e
1.31 104 4.36 105 e e e
e e 4.76 106 e e
e e e 8.18 105 e
e e e e 1.27 106
e e 2.85 106 2.05 105 e
Chemicals NaOH solution 50% (kg) NaClO solution 25% (kg) HCl solution 5% (kg)
e e e
e e e
e e e
e e e
9.99 105 6.35 104 1.82 105
e e e
Water From WWTP secondary effluent Potable water
e e
2.68 102 e
5.23 102 e
e e
e 1.25 102
5.23 102 e
Electricity From the grid (kWh)
e
3.95 104
2.40 104
6.80 104
5.23 104
9.00 104
Transport Lorry 3.5e7.5 t (kg km)
e
3.72 102
3.52 103
6.05 102
2.74 104
6.27 102
Final disposal Composting (kg) Landfill deposits (kg)
e e
1.74 104 e
e 4.76 106
e 8.18 105
e 8.25 106
e 2.33 105
102 102 103 102 102 101 101 101 102 102 102 102 103 101 103 102 101 102 102 102 102 102 102 102 102 103 101 103 103
6.55 5.28 4.18 9.63 3.29 1.05 4.62 5.88 6.70 1.10 4.54 2.09 5.76 1.77 1.06 8.53 7.36 2.09 2.87 8.84 1.97 3.14 2.68 8.84 2.68 8.84 1.13 2.66 1.19 2.55 1.01 1.34 14.19
104 104 104 104 103 103 103 103 104 103 102 104 103 103 102 104 104 101 102 104 102 103 103 104 103 104 104 103 103 102 103 103
to maintain the moisture conditions required and to compensate for evaporation losses. The pressure drop associated with each technology determined the energy required for gas circulation according to the Eq. (1):
E ¼ Qgas $DP$BE1
(1)
where E is the energy needed (kWh), Qgas is the gas flow (m3s1), DP is the pressure drop in the reactor (kPa) and BE is the blower
1.31 1.06 8.36 9.63 6.57 1.05 4.62 5.88 6.70 2.19 4.54 2.09 1.15 3.54 2.13 1.71 1.47 2.09 5.73 1.77 1.97 6.29 5.36 1.77 5.36 1.77 1.13 5.33 2.38 5.09 2.02 2.69 13.97
103 103 104 104 103 103 103 103 104 103 102 104 102 103 102 103 103 101 102 103 102 103 103 103 103 103 104 103 103 102 103 103
2.62 2.11 1.67 9.63 1.31 1.05 4.62 5.88 6.70 4.38 4.54 2.09 2.30 7.08 4.25 3.41 2.94 2.09 1.15 3.53 1.97 1.26 1.07 3.53 1.07 3.53 1.13 1.07 4.76 1.02 4.04 5.37 0
104 104 104 103 105 102 102 102 103 104 101 103 105 106 103 106 104 101 104 104 101 105 105 104 105 104 103 105 106 102 106 106
1.31 1.06 8.36 9.63 6.57 1.05 4.62 5.88 6.70 2.19 4.54 2.09 1.15 3.54 2.13 1.71 1.47 2.09 5.73 1.77 1.97 6.29 5.36 1.77 5.36 1.77 1.13 5.33 2.38 5.09 2.02 2.69 18.63
103 103 104 104 103 103 103 103 104 103 102 104 102 103 102 103 103 101 102 103 102 103 103 103 103 103 104 103 103 102 103 103
2.62 2.11 1.67 9.63 6.57 1.05 4.62 5.88 6.70 4.38 4.54 2.09 1.15 3.54 4.25 1.71 2.94 2.09 5.73 3.53 1.97 6.29 5.36 3.53 5.36 3.53 1.13 5.33 2.38 1.02 2.02 2.69 13.97
105 105 105 105 106 104 104 104 105 105 103 105 105 106 104 106 105 101 105 105 103 106 106 105 106 105 105 106 106 103 106 106
efficiency, in which case a typical value of 0.7 was assumed (Estrada et al., 2011). 2.3.1. Biofilter (BF) This configuration consisted of a reactor packed with a mixture of organic and inorganic materials: 75% compost and 25% perlite (v/ v) (Lebrero et al., 2010). The BF was built in standard concrete with a volume of 694 m3 (1 m height) and operated at an EBRT of 50 s. Irrigation water (129 kg h1) from the secondary effluent of the WWTP was used. A standard pressure drop of 1000 Pa was
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considered. The malodorous air was moistened in a humidifier prior to biofiltration with additional water (1211 kg h1). The leachate produced from the BF and the excess water from the humidifier were returned to the WWTP headworks for further treatment. The mixture of compost and perlite was replaced after 2 years of operation and processed in a composting facility, since its characteristics are similar to those of an average organic waste like sludge. 2.3.2. Biotrickling filter (BTF) A two-stage biotrickling filter of 208 m3 with a total EBRT of 15 s was selected. The glass fibre BTF was packed with inert plastic supports, which exhibited a density of 100 kg m3 and a total pressure drop of 500 Pa (Dorado et al., 2009). In the first stage, the pH was maintained at 2 for the efficient removal of H2S and in a second stage a neutral pH was kept to favour the removal of the other odorants. Water was supplied by spray nozzles and recirculated at a rate of 7.2 L m3 min1 using a renewal ratio of 2.5 L of water added per gram of H2S removed. Landfills were selected as the disposal scenario for the packed-bed after 10 years of lifespan.
2.3.3. Chemical scrubber (CS) A two-stage NaOHeNaClO scrubber (2 m) packed with Intalox Saddles (density of 100 kg m3) was considered, operated with a total pressure drop of 500 Pa. The glass fibre reactor was operated at an EBRT of 4 s (2 s per stage), which entailed a reactor volume of 55.6 m3. The liquid phase was recirculated at a rate of 180 L m3 min1 via spray nozzles (with a pressure drop of 50 kPa) (Gabriel and Deshusses, 2004) while 15040 L of additional water were added daily. The pH was maintained at 10e12 in the first stage, whereas the second stage was operated at a slightly lower pH (8.5e9.5). NaClO (25% w/w) at a flow rate of 28.85 L h1 and NaOH (50% w/w) at a flow rate of 3.26 L h1 were continuously supplied. The packing material was washed once a year with HCl (5% w/w) in order to avoid the accumulation of precipitates in the medium. The packing material was disposed of in a landfill after 10 years of operation. 2.3.4. Activated carbon filter (AC) Two glass fibre towers of 42 m3 packed with granular impregnated activated carbon (density of 430 kg m3) and operating at an EBRT of 3 s and a pressure drop of 1750 Pa were used as a model adsorption filter. One filter was in operation while the second one was in stand-by for the periodic renewal of the activated carbon every 6 months, which was later disposed of in landfills. Regeneration of the activated carbon was not considered according to the current operation of most full scale facilities (Estrada et al., 2011, 2012). 2.3.5. Biotrickling filter coupled with an activated carbon tower (BTF þ AC) The hybrid technology consisted of a biotrickling filter coupled with an activated carbon unit. The biotrickling filter, packed with 2 m of inert plastic support (density of 100 kg m3 and pressure drop of 500 Pa), was operated at an EBRT of 9 s. The 42 m3 AC tower was operated at an EBRT of 3 s and pressure drop of 1750 Pa. The main difference between the AC coupled to a biotrickling filter and the AC operated stand-alone was the lifetime of the activated carbon in the combined configuration: 2 years, due to the lower odorant concentrations in the inlet stream. The packing materials were identical to the previous configurations and consequently, landfills were considered for disposal.
2.4. Life cycle inventory, data quality and simplifications The LCI included the emission data reported by Zarra et al. (2008) as well as theoretical and literature data for each technology concerning the odorant concentrations in the outlet emission, infrastructure material, packing bed material, energy requirements, chemicals use, cleaning agents, water consumption, transport distances and final disposal of wastes (Table 1). The inventory data for auxiliary processes were taken from the Ecoinvent database and completed with recent bibliography (Table 2). The Spanish profile for energy was adapted with data of the electricity production/importation system for Spain in 2011 (RedElectrica, 2011). The electricity transmission network, emissions of sulphur hexafluoride to air, and losses during the medium voltage transmission and transformation from high to medium voltage were also considered (Dones et al., 2007). The long term stability of the technologies was indirectly estimated by the lifespan of the packing materials throughout the operation of the systems, which was applied to calculate packing material requirements per FU. The production of Intalox Saddles for the CS configuration was modelled according to the process of polyvinyl chloride (PVC) production (Hischier, 2007). Granular activated carbon (GAC) from coal was considered in the AC system ~ oz, 2006). The amount of infrastructure material (concrete or (Mun glass fibre) was calculated assuming a wall thickness of 15 cm of concrete and 15 mm of glass fibre. The production processes of NaOH, NaClO and HCl were taken from the Ecoinvent database and adapted to the concentrations and consumptions used in this case study (Althaus et al., 2007). Diaphragm cells were considered for the production of NaOH (50%w/w), since it is the most common process applied (Althaus et al., 2007). The production processes of NaClO (25%w/w) and HCl (5%w/w) were estimated from process data of 15%w/w and 30%w/w solutions, respectively (Althaus et al., 2007). Finally, the production of 1 kg of compost from 1 kg of packing material was assumed within the composting process. Inventory data considered in the LCI were related to the steadystate operation of the systems. Therefore, the evaluation of the
Table 2 Summary of data sources considered in this study. Inventory inputs Energy Infrastructure material
Packing material
Chemicals
Water Transport Waste treatment
Data sources Electricity (Spanish electricity profile) Concrete
Ecoinvent database (Dones et al., 2007) Ecoinvent database (Kellenberger et al., 2007) Glass fibre Ecoinvent database (Kellenberger et al., 2007) Compost Ecoinvent database (Nemecek and K€ agi, 2007) Expanded perlite Ecoinvent database (Kellenberger et al., 2007) ~ oz (2006) Granular activated carbon Mun Polyurethane, rigid foam Ecoinvent database (Hischier, 2007) PVC Ecoinvent database (Hischier, 2007) Sodium hydroxide (50%) Ecoinvent database (Althaus et al., 2007) Sodium hypochlorite (15%) Ecoinvent database (Althaus et al., 2007) Hydrochloric acid (30%) Ecoinvent database (Althaus et al., 2007) Tap water Ecoinvent database (Althaus et al., 2007) Truck 3.5e7.5 t, EURO 4 Ecoinvent database (Spielmann et al., 2007) Landfill deposits Ecoinvent database (Doka, 2007)
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robustness of the technologies towards process fluctuations and operational upsets was out of the scope of the manuscript. For a detailed evaluation of the robustness of the most common odour treatment technologies see Estrada et al. (2012). 2.5. Impact categories In the LCIA phase, the role of each element of the LCI was quantitatively assessed in terms of its contribution to nine impact categories: i) Cumulative Energy Demand (CED), which accounts for direct and indirect energy use throughout the entire life cycle of a product or process (Huijbregts et al., 2010); ii) malodorous air, defined as the quantity of contaminated air exceeding the odour threshold value established for each compound (Heijungs et al., 1992) and evaluated using the methodology developed by the Centre of Environmental Science of Leiden University (CML, 2001) e et al., 2002), and; iii) climate change, freshwater eutro(Guine phication, photochemical oxidant formation, human toxicity, and marine, terrestrial and freshwater ecotoxicity assessed with the ReCiPe Midpoint method using the Hierarchist Perspective (H) version 1.05 (Goedkoop et al., 2009). The software SimaPro 8 was used for the computational implementation of all the inventories (http://www.pre.nl/simapro). 3. Results 3.1. Malodorous air The impact category of malodorous air is of major relevance in this research study. The impact of malodorous air in the NT scenario was very high: 55-times higher than the impact related to the worst case scenario (BF): 48728 m3 m3 for NT vs. 894 m3 m3 for BF. The treatment of packing material in composting facilities entailed the highest impacts for BF in this category. At this point, it must be stressed that the potential of malodorous air was reduced by 98% by BF and 99% with other alternative technologies. Only the results related to the five technological alternatives were plotted in Fig. 1 based on the large differences between the NT scenario and the impacts associated with the different technologies. The main impact associated with the technology performance was that corresponding to the treated effluent released to the atmosphere. The treated gas accounted for values of malodorous air ~95% in BTF, AC, CS and BTF þ AC, whereas odorants associated with the packing material production and disposal contributed to ~50% of the impact in BF (Fig. 1). Among the odorants released to the atmosphere, H2S was found the main responsible of the malodorous air impact related to direct emissions.
Fig. 1. Malodours air impact for the five technologies evaluated. Malodorous air of NT: 48728 m3 per FU.
Fig. 2. Cumulative energy demand (CED) for the five technologies evaluated and NT scenario.
3.2. Cumulative energy demand (CED) No cumulative energy demand was computed in the NT scenario. Therefore, the potential information that this indicator provides is only useful when identifying the most energy efficient technology. The lowest CED impacts were obtained for BF and BTF, where this parameter was mainly associated with the use of energy for air and water pumping (Fig. 2). The impact associated with AC was mainly due to the packing material production, accounting for 76% of the overall impact. Both GAC production and energy use represented the impact associated with the hybrid technology, higher than those of biological techniques but ~50% lower than AC potential impact. Finally, the production of chemicals used in the CS contributed with 72% to the global CED impact. 3.3. Climate change The odorants present in the target malodorous emission do not contribute directly to this category and, consequently, climate change impact in the NT scenario was considered negligible. Moreover, although the oxidation of carbon-based odorants implies the emission of CO2, this contribution was deemed negligible when compared to other CO2 sources such as the indirect emissions derived from packing material or chemicals use (Fig. 3). Biological treatments exhibited the lowest climate change impacts. In fact, BTF and BF presented impacts ~91% and 84% lower than the impact associated with AC (Fig. 3). Process energy requirements in BTF and BF were the main contributors to climate change, while GAC production and transportation was the main responsible of its climate change potential in AC. The use of chemicals in CS accounted for
Fig. 3. Climate change impact for the five technologies evaluated and NT scenario.
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approximately 80% of the total impact, with NaClO presenting the largest contribution. The hybrid technology presented lower impact than physical/chemical technologies but higher than biological alternatives. 3.4. Freshwater eutrophication This impact category is not directly applicable to the NT scenario as the odorants present in the raw emission do not contribute to water eutrophication. CS presented the highest eutrophication impact due to the eutrophying emissions related to the production of chemicals required (Fig. 4). In this particular case, NaClO production accounted for 88% of the impact associated with chemicals. Biological treatments presented the lowest impacts, with electricity use as the main contributor in BF (91%) and BTF (55%). Similarly, energy use accounted for up to 50% of the eutrophication impact in the hybrid system. Finally, GAC production in AC represented 63% of the eutrophication potential. 3.5. Photochemical oxidant formation A total impact of 3.78 106 kg NMVOC per FU was calculated in the NT scenario as a result of the direct release of VOCs to the atmosphere (Fig. 5). The most important impacts were observed for physical/chemical alternatives, followed by BTF þ AC and finally, biological technologies. Packing material accounted for 56% of the total impact in AC whereas chemicals production entailed remarkable impact in CS. Significantly lower impacts compared to the NT scenario of 38%, 77% and 80% were estimated in BTF þ AC, BF and BTF, respectively. The photochemical oxidant formation impact related to the production of energy was considerable in all technologies evaluated. Finally, it is worth noting that the use of glass fibre as a building material resulted in higher impact compared to the use of concrete. 3.6. Ecotoxicity and human toxicity The toxicity associated with the treatment of the target malodorous emission was assessed in freshwater, marine and terrestrial environments and humans (Fig. 6 aed). Biological treatments presented the most environmentally friendly performance, with energy use appearing as the main responsible of the impact associated with BF. Production of glass fibre and disposal in landfills entailed the largest impacts in BTF. The chemicals added to CS (especially NaClO) induced the largest impacts regardless of the toxicity category. The contributors to toxicity-associated impacts in
Fig. 4. Freshwater eutrophication impact for the five technologies evaluated and NT scenario.
Fig. 5. Photochemical oxidant formation of the five technologies evaluated and NT scenario.
AC and BTF þ AC depended to a great extent on the categories analyzed, disposal of packing material, energy use and glass fibre production contributing significantly to freshwater and marine ecotoxicity, terrestrial ecotoxicity and human toxicity, respectively. Thus, although direct odorant emissions caused a negative impact on the receiving environment and on human health, this impact was considered negligible when compared to the indirect impacts related to energy consumption, the production of the packing material or the bed material disposal in landfills. 4. Discussion Overall, biological processes exhibited an improved environmental performance, with the exception of the malodorous air category, where emissions from the composting process led to more significant impacts in BF. In this context, it must be stressed that although the conversion of organic wastes to compost reduces their potential impacts to the environment, composting entails negative environmental impacts, such as the emission of H2S €gi, 2007). The main contributor to malodorous air (Nemecek and Ka associated with the technologies were direct emissions, H2S appearing as the most significant odorant compound. Firstly, the elevated H2S concentration in the raw emission (2.09 101 mg H2S m3) resulted in a significantly high concentration of H2S in the outlet stream despite the great H2S removal efficiency (99%) achieved in all the technologies assessed. Secondly, this compound presents a characterisation factor of 2.33 109 m3 per kg of H2S, which is calculated as the inverse of the Odour Threshold Value (OTV). H2S exhibits a significantly low OTV (4.3 104 mg m3), which entailed negative impacts even at trace level concentrations. Treatment of odorants entailed a reduction of 98e99% in the malodorous impact category. The two biological treatments were found to be the friendliest alternatives to the environment, with diverse outcomes in the different impact categories. BF was the most suitable technology in terms of freshwater eutrophication, human toxicity, freshwater ecotoxicity and marine ecotoxicity, whereas the BTF exhibited the lowest impacts in terms of CED, climate change, photochemical oxidant formation and terrestrial ecotoxicity. Energy use was found to be the largest contributor to the impact in the BF scenario. More specifically, this consumption was directly related to the pressure drop associated with the packing material. On the other hand, waste disposal in landfills and glass fibre used in BTF significantly influenced its overall environmental impact, whereas concrete production was almost negligible in BF. CS presented the highest impacts in freshwater eutrophication, photochemical oxidant formation, human toxicity and ecotoxicity
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Fig. 6. Ecotoxicity in (a) freshwater, (b) marine and (c) terrestrial environments and (d) human toxicity associated with the five technologies evaluated and NT scenario.
due to the use of large amounts of chemicals. The significant amount of NaClO required for odorant oxidation had a key role in the impacts associated with the use of chemicals. The negative burdens derived from the use of chemicals in odour abatement technologies were also pointed out by Alfonsín et al. (2013) when assessing the environmental consequences for human toxicity and ecotoxicity of conventional biofilters at laboratory scale. Likewise, this high impact related to the use of chemicals in CS was in accordance with the results obtained by Estrada et al. (2011). Nevertheless, the authors considered the amount of reagents used per year, chemicals production being out of the analysis. The impact associated with AC was similar to that related to NT scenario in photochemical oxidant formation mainly due to packing material production, whereas the use of chemicals caused the highest impact in CS. Therefore, indirect impacts were found more important than the reduction of butanone directly released to the atmosphere, which represented ~75% of the impact associated with the NT scenario. On the other hand, AC showed the worst results for CED and climate change potentials followed by CS in both categories. As already mentioned, the most significant parameter determining the ~ oz, 2006), which AC impact was the GAC production process (Mun was mainly attributed to a large demand of natural gas and its replacement every 6 months. The contribution of the packing material to the environmental impact showed differences with the preliminary analysis conducted by Estrada et al. (2011), who reported that BF was the technology presenting the highest annual packed bed-material requirements. Nevertheless, LCA demonstrated that GAC production associated impacts were more relevant than those related to BF packing materials. Finally, the hybrid technology (BTF þ AC) showed a lower environmental performance than its biological counterparts, but
better results than physical/chemical technologies. BTF þ AC exhibited an improved environmental performance as compared to AC filtration due to the fact that the replacement frequency of GAC was reduced to 2 years instead of 6 months for stand-alone AC. Estrada et al. (2011), using the IChemE metrics, concluded that BTF þ AC was the less favourable option in terms of energy consumption. In this study, energy use in BTF þ AC was found to be significant in categories such as terrestrial ecotoxicity, but in no case was this contribution comparable to the global impact of physical/chemical technologies. Therefore, BTF þ AC could be ranked second after its biological counterparts, its implementation being highly recommended in odour sensitive scenarios due to its more reliable performance. 5. Conclusions Biological treatments exhibited the best environmental performance among the technologies evaluated, except in terms of malodorous air impact where BF showed the highest impact as a result of the contribution of compost production and disposal. Nonetheless, BF achieved a reduction in the malodour air impact of 98% compared to the NT scenario, which was slightly lower than the reduction achieved by the other technologies (99%). Even when the use of a technology implied more significant impacts than the NT scenario, BF and BTF showed an improved environmental performance compared to physical/chemical technologies. In this context, AC and CS presented the highest impacts due to the use of GAC as packing material, chemical use and waste disposal in landfills. Finally, the operation of BTF þ AC was more recommendable from an environmental sustainability perspective than physical/chemical treatments, but slightly worse than biotechnologies. In other words, based on the fact that odour treatment is either compulsory
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to comply with, based on the legislative thresholds, or highly recommended, in order to avoid public complaints, biotechnologies are the best treatment options from an environmental point of view. Acknowledgements The authors acknowledge the financial support of the Spanish Ministry of Economy and Competitiveness (project CTQ2012-34949 and contract BES-2010-030994) and the European Union (LIVEWASTE project: LIFE12 ENV/CY/544). The authors from the University of Santiago de Compostela belong to the Galician Competitive Research Group GRC 2013-032, programme co-funded by FEDER. References Alfonsín, C., Hernandez, J., Omil, F., Prado, O.J., Gabriel, D., Feijoo, G., Moreira, M.T., 2013. Environmental assessment of different biofilters for the treatment of gaseous streams. J. Environ. Manage. 129, 463e470. Althaus, H.J., Chudacoff, M., Hischier, R., Jungbluth, N., Osses, M., Primas, A., 2007. Life Cycle Inventories of Chemicals. Ecoinvent Report no. 8. Swiss Centre for Life Cycle Assessment, Dübendorf. Balasubramanian, P., Philip, L., Murty Bhallamudi, S., 2012. Biotrickling filtration of VOC emissions from pharmaceutical industries. Chem. Eng. J. 209, 102e112. Baumann, H., Tillman, A.M., 2004. The Hitch Hiker's Guide to LCA : an Orientation in Life Cycle Assessment Methodology and Application. Studentlitteratur, Lund. Doka, G., 2007. Life Cycle Inventories of Waste Treatment Services. Ecoinvent report No. 13. Swiss Centre for Life Cycle Inventories, Dübendorf. Dones, R., Bauer, C., Bolliger, R., Burger, B., Faist Emmenegger, M., Frischknecht, R., € der, A., Tuchschmid, M., 2007. Life Cycle Inventories of Heck, T., Jungbluth, N., Ro Energy Systems: Results for Current Systems in Switzerland and Other UCTE Countries. Ecoinvent report no. 5, Paul Scherrer Institut Villigen. Swiss Centre for Life Cycle Assessment, Dübendorf, CH. Dorado, A.D., Rodríguez, G., Ribera, G., Bonsfills, A., Gabriel, D., Lafuente, J., Gamisans, X., 2009. Evaluation of mass transfer coefficients in biotrickling filters: experimental determination and comparison to correlations. Chem. Eng. Technol. 32, 1941e1950. Estrada, J.M., Kraakman, N.J., Munoz, R., Lebrero, R., 2011. A comparative analysis of odour treatment technologies in wastewater treatment plants. Environ. Sci. Technol. 45, 1100e1106. ~ oz, R., 2012. A sensitivity analysis of Estrada, J.M., Kraakman, N.J.R., Lebrero, R., Mun process design parameters, commodity prices and robustness on the economics of odour abatement technologies. Biotechnol. Adv. 30, 1354e1363. Gabriel, D., Deshusses, M.A., 2004. Technical and economical analysis of the conversion of a full-scale scrubber to a biotrickling filter for odor control. Water Sci. Technol. 50, 309e318. Goedkoop, M., Heijungs, R., Huijbregts, M., De Schryver, A., Struijs, J., van Zelm, R., 2009. ReCiPe 2008 e a Life Cycle Impact Assessment Method Which comprises harmonised Category Indicators at the Midpoint and the Endpoint Level. Ministry of Housing, Spatial Planning and Environment (VROM), The Netherlands. e, J.B., Gorre e, M., Heijungs, R., Huppes, G., Kleijn, R., Koning, A.d., Oers, L.v., Guine Wegener Sleeswijk, A., Suh, S., Udo de Haes, H.A., Brujin, H.d., va., D.R., Huijbregts, M.A.J., 2002. Handbook on Life Cycle Assessment. Operational Guide to the ISO Standards. I: LCA in Perspective. IIa: Guide. IIb: Operational Annex. III: Scientific Background. Kluwer Academic Publishers, Dordrecht, ISBN 1-40200228-9. e, J.B., Huppes, G., Lankreijer, R.M., Udo de Haes, H.A., Wegener Heijungs, R., Guine Sleeswijk, A., Ansems, A.M.M., Eggels, P.G., Duin, R.v., Goede, H.P.d., 1992.
Environmental life cycle assessment of products. I: Guide - October 1992. II: Backgrounds - October 1992. CML (NOH Report 9266 þ 9267), ISBN 90-5191064-9 (Leiden). Hischier, R., 2007. Life Cycle Inventories of Packaging and Graphical Papers. Ecoinvent-Report No. 11. Swiss Centre for Life Cycle Inventories, Dübendorf. ndez-Couto, M., Feijoo, G., 2004. Environmental Hospido, A., Moreira, M., Ferna performance of a municipal wastewater treatment plant. Int. J. Life Cycle Assess. 9, 261e271. Hospido, A., Sanchez, I., Rodriguez-Garcia, G., Iglesias, A., Buntner, D., Reif, R., Moreira, M.T., Feijoo, G., 2012. Are all membrane reactors equal from an environmental point of view? Desalination 285, 263e270. Huijbregts, M.A., Hellweg, S., Frischknecht, R., Hendriks, H.W., Hungerbuhler, K., Hendriks, A.J., 2010. Cumulative energy demand as predictor for the environmental burden of commodity production. Environ. Sci. Technol. 44, 2189e2196. IChemE, 2002. The Sustainability Metrics. The Institution of Chemical Engineers, Rugby, UK. ISO, 2006a. ISO 14040: 2006. Environmental Management e Life Cycle Assessment e Principles and Framework. ISO, 2006b. ISO 14044: 2006. Environmental Management - Life Cycle Assessment Requirements and Guidelines. Kellenberger, D., Althaus, H.J., Jungbluth, N., Künniger, T., 2007. Life Cycle Inventories of Building Products. Final report Ecoinvent data v2.0 No. 7. Swiss Centre for Life Cycle Inventories, Dübendorf, Switzerland. Kellenberger, D., Althaus, H.J., Jungbluth, N., Künniger, T., 2007. Life Cycle Inventories of Building Products. Final report Ecoinvent data v2.0 No. 7. Swiss Centre for Life Cycle Inventories, Dübendorf, Switzerland.Lebrero, R., rez, R., García-Encina, P.A., Mun ~ oz, R., 2014. Comparative Gondim, A.C., Pe assessment of a biofilter, a biotrickling filter and a hollow fiber membrane bioreactor for odour treatment in wastewater treatment plants. Water Res. 49, 339e350. ~ oz, R., 2012. Odor Lebrero, R., Rodríguez, E., Estrada, J.M., García-Encina, P.A., Mun abatement in biotrickling filters: effect of the EBRT on methyl mercaptan and hydrophobic VOCs removal. Bioresour. Technol. 109, 38e45. ~ oz, R., 2010. H2S and Lebrero, R., Rodríguez, E., Martin, M., García-Encina, P.A., Mun VOCs abatement robustness in biofilters and air diffusion bioreactors: a comparative study. Water Res. 44, 3905e3914. ~ oz, I., 2006. Life Cycle Assessment as a Tool for Green Chemistry: Application to Mun Different Advanced Oxidation Processes for Wastewater Treatment. Doctoral Thesis. Autonomic University of Barcelona. Nemecek, T., K€ agi, T., 2007. Life Cycle Inventories of Swiss and European Agricultural Production Systems. Final report ecoinvent V2.0 No. 15a. Agroscope Reckenholz-Taenikon Research Station ART, Swiss Centre for Life Cycle Inventories, Zurich and Dübendorf, CH. Omri, I., Aouidi, F., Bouallagui, H., Godon, J.-J., Hamdi, M., 2013. Performance study of biofilter developed to treat H2S from wastewater odour. Saudi J. Biol. Sci. 20, 169e176. ctrica de Espan ~ a. El sistema ele ctrico espan ~ ol (in RedElectrica, 2011. Red Ele Spanish). Revah, S., Morgan-Sagastume, J., 2005. Methods of odor and VOC control. In: Shareefdeen, Z., Singh, A. (Eds.), Biotechnology for Odor and Air Pollution Control. Springer, Berlin Heidelberg, pp. 29e63. Rodriguez-Garcia, G., Molinos-Senante, M., Hospido, A., Hernandez-Sancho, F., Moreira, M.T., Feijoo, G., 2011. Environmental and economic profile of six typologies of wastewater treatment plants. Water Res. 45, 5997e6010. Schlegelmilch, M., Streese, J., Stegmann, R., 2005. Odour management and treatment technologies: an overview. Waste Manage. 25, 928e939. Shaw, A.R., Koh, S.-H., 2013. Gaseous emissions from wastewater facilities. Water Environ. Res. 85, 1298e1307. Spielmann, M., Bauer, C., Dones, R., Tuchschmid, M., 2007. Transport Services. Ecoinvent report No. 14. Swiss Centre for Life Cycle Inventories, Dübendorf. Zarra, T., Naddeo, V., Belgiorno, V., Reiser, M., Kranert, M., 2008. Odour monitoring of small wastewater treatment plant located in sensitive environment. Water Sci. Technol. 58, 89e94.
Contents lists available at ScienceDirect
Journal of Environmental Management journal homepage: www.elsevier.com/locate/jenvman
Selection of odour removal technologies in wastewater treatment plants: A guideline based on Life Cycle Assessment M. Estrada b, c, Raúl Mun ~ oz b, Carolina Alfonsín a, *, Raquel Lebrero b, Jose N.J.R. (Bart) Kraakman d, e, Gumersindo Feijoo a, Mª Teresa Moreira a a
Department of Chemical Engineering, Institute of Technology, University of Santiago de Compostela, 15782 Santiago de Compostela, Galicia, Spain Department of Chemical Engineering and Environmental Technology, Escuela de Ingenierías Industriales, Sede Dr. Mergelina, University of Valladolid, Dr Mergelina s/n, 47011 Valladolid, Spain c School of Engineering, London South Bank University, UK d Department of Biotechnology, Delft University of Technology, Julianalaan 67, 2628 BC Delft, The Netherlands e CH2M Hill, Level 7, 9 Help Street, Chatswood, NSW 2067, Australia b
a r t i c l e i n f o
a b s t r a c t
Article history: Received 22 July 2014 Received in revised form 9 October 2014 Accepted 13 October 2014 Available online
This paper aims at analysing the environmental benefits and impacts associated with the treatment of malodorous emissions from wastewater treatment plants (WWTPs). The life cycle assessment (LCA) methodology was applied to two biological treatments, namely biofilter (BF) and biotrickling filter (BTF), two physical/chemical alternatives, namely activated carbon tower (AC) and chemical scrubber (CS), and a hybrid combination of BTF þ AC. The assessment provided consistent guidelines for technology selection, not only based on removal efficiencies, but also on the environmental impact associated with the treatment of emissions. The results showed that biological alternatives entailed the lowest impacts. On the contrary, the use of chemicals led to the highest impacts for CS. Energy use was the main contributor to the impact related to BF and BTF, whereas the production of glass fibre used as infrastructure material played an important role in BTF impact. Production of NaClO entailed the highest burdens among the chemicals used in CS, representing ~90% of the impact associated to chemicals. The frequent replacement of packing material in AC was responsible for the highest environmental impacts, granular activated carbon (GAC) production and its final disposal representing more than 50% of the impact in most categories. Finally, the assessment of BTF þ AC showed that the hybrid technology is less recommendable than BF and BTF, but friendlier to the environment than physical/chemical treatments. © 2014 Elsevier Ltd. All rights reserved.
Keywords: Activated carbon Biofiltration Chemical scrubbing Life cycle assessment (LCA) Odour abatement Wastewater treatment plants (WWTPs)
1. Introduction Wastewater treatment plants (WWTPs) are considered important sources of gaseous emissions, including green house gases (GHG) and odorants (Shaw and Koh, 2013). The malodorous emissions associated with treatment processes are considered one of the major concerns of exposed population living in surrounding areas of WWTPs. In this context, the concentration of malodours in the air is often monitored and controlled with the aim of complying with odour regulations while keeping a respectable public image of the emission sources.
* Corresponding author. Tel.: þ34 881 816 739. E-mail addresses: [email protected] (C. Alfonsín), [email protected]. ~ oz), es (R. Lebrero), [email protected] (J.M. Estrada), [email protected] (R. Mun [email protected] (N.J.R.(B. Kraakman), [email protected] (G. Feijoo), [email protected] (M.T. Moreira). http://dx.doi.org/10.1016/j.jenvman.2014.10.011 0301-4797/© 2014 Elsevier Ltd. All rights reserved.
Odour abatement technologies have been widely investigated as cost-efficient and reliable alternatives for the mitigation of odour nuisance (Revah and Morgan-Sagastume, 2005; Schlegelmilch et al., 2005). These technologies are commonly classified into physical/chemical and biological techniques. Physical/chemical technologies have been broadly implemented as a consequence of their rapid start-up, low empty bed residence time (EBRT) and consolidated know-how and experience in design and operation. These techniques are often based on the transfer of odorants from the gas emission to either a solid (adsorption) or liquid (absorption) phase. These pollutants can be further transformed into byproducts according to their reactivity with the chemicals used. However, in the last decades biological systems have been increasingly used due to their ability to efficiently treat malodorous emissions at lower operating costs (Schlegelmilch et al., 2005). The main advantages of bioprocesses compared to their physical/ chemical counterparts derive from their low generation of secondary wastes and low demand of resources, such as chemicals or
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adsorbent media. On the other hand, biological processes often require larger EBRTs (2e120 s vs. 1e5 s) and associated footprint than physical/chemical alternatives at similar odour removal efficiencies. Technologies for odour treatment have been widely reviewed in the literature in order to establish their optimal range of application and performance for the removal of volatile organic compounds (VOCs) and volatile inorganic compounds (VICs). Nowadays, there are enough experimental evidences regarding the capability of biofilters and biotrickling filters to achieve significantly high removal efficiencies of air pollutants at both trace levels (Lebrero et al., 2014, 2012) and industrial concentrations (Balasubramanian et al., 2012; Omri et al., 2013). Estrada et al. (2012) demonstrated that biological techniques were the most cost-efficient alternatives with lower sensitivity to design parameters and lower operating costs than physical/chemical treatments at typical odour concentrations emitted in WWTPs. In terms of sustainability, Estrada et al. (2011) assessed the performance of different physical/chemical and biological odour abatement technologies based on the IChemE Metrics methodology (IChemE, 2002). The analysis focused on major environmental indicators such as resource use, waste production and emission impacts as well as process economics and social impact. This preliminary study showed as compared to physical/chemical technologies, biological treatments presented lower demand of energy, material and chemicals and limited production of hazardous wastes. These systems were the most favourable option despite the required high initial investment costs when analysing investment costs and operational costs over 20 years. Although the aforementioned reports offered valuable information for the selection of the most suitable technology for the treatment of malodorous emissions, the IChemE Metrics methodology is only focused on the target process and does not consider a holistic approach for the assessment. For instance, despite water consumption or material use being considered as environmental indicators, the impact of their production processes is out of the scope of the IChemE Metrics analysis. Nevertheless, the origin and management of the components/consumables of the system under study could also be major contributors to the environmental impact. Previous research papers showed how a complete environmental evaluation helps to select the most adequate treatment technology, considering not only removal efficiencies but also potential environmental impacts (Alfonsín et al., 2013; Hospido et al., 2012). Life cycle assessment (LCA) methodology is a quantitative procedure to assess the environmental burdens associated with products, processes and services, which is commonly used for environmental impact evaluation (Baumann and Tillman, 2004; ISO, 2006b). Although this methodology has been widely and satisfactorily applied to technologies dealing with wastewater treatment (Hospido et al., 2004; Rodriguez-Garcia et al., 2011), there is, to the best of our knowledge, only one application of LCA to the abatement of gaseous emissions, focused on lab-scale biofilters (Alfonsín et al., 2013). This study aimed at enriching the guidelines for the selection of the best technological alternative for the treatment of malodorous emissions under an environmental perspective. In this context, a comprehensive LCA was performed in order to determine the overall environmental impacts of five full-scale odour abatement technologies in a WWTP scenario: biofilter (BF), biotrickling filter (BTF), activated carbon filter (AC), chemical scrubber (CS) and a hybrid technology consisting of a BTF coupled with AC (BTF þ AC). The results obtained for each technology were compared to the environmental impact derived from the direct discharge of the malodorous emission without treatment, named non treatment scenario (NT).
2. Materials and methods The ISO 14040 (ISO, 2006a) standard determines four basic stages for LCA studies: goal and scope definition, life cycle inventory (LCI), life cycle impact assessment (LCIA) and finally, interpretation of results (Baumann and Tillman, 2004). Goal and scope definition constitutes the first phase of an LCA and aims at defining the boundaries of the study and the quality of data used. A functional unit (FU), which represents the function of the system under study, must be also established in this phase. Then, LCI is performed, which involves data collection and interpretation of inputs and outputs. The allocation procedure is also conducted during the LCI phase, which consists on distributing input and output flows among the products of the process. LCIA represents the third phase and its purpose is to convert LCI data into potential impacts associated to products and processes. LCIA includes two mandatory steps (i.e. classification and characterization) and other optional elements, such as normalization and weighting. Finally, the interpretation of the results allows identifying the hot spots of the process as well as recommending options to reduce the environmental burdens. 2.1. Goal and scope definition This study was conducted to evaluate the environmental impacts associated with the performance of five of the most commonly implemented odour abatement technologies in WWTPs. The functional unit chosen in this study was 1 m3 of treated air, which is consistent with the approach used in a previous study evaluating lab-scale gas-phase bioreactors (Alfonsín et al., 2013). 2.2. System boundaries The system boundaries determine the units and elements of the process included in the analysis (ISO, 2006a). The assessment here conducted considered the incoming polluted emissions, which are directly discharged to the atmosphere in the NT scenario, the output treated emission, material used in the infrastructure, consumables (e.g. packing material), potable or secondary plant effluent water, energy use, chemicals and transportation and final disposal of all wastes. Water input is required in all systems except in AC and potable water instead of WWTP secondary effluent water is used in CS. The secondary effluent water of the WWTP is considered a product related to wastewater treatment and no impact is attributed when used for irrigating the packing material of the biofilter or the biotrickling filter. The leachate collected in each technology (except in AC) was not considered in the analysis since it is returned to the WWTP headworks with a negligible flow in comparison with the net flow treated in the WWTP. The transportation from the manufacturing industry to the odour treatment facility of consumables like packing materials and chemicals was included. 2.3. Description of odour treatment technologies Five odour abatement technologies were assessed: two biological (BF and BTF), two physical/chemical (CS and AC) and a hybrid technology (BTF þ AC). The model malodorous stream of 50000 m3h1, representing a typical emission from WWTPs, included 30 different VOCs (Zarra et al., 2008), methanethiol (1.97 mg m3) and hydrogen sulphide (20.9 mg m3) (Table 1). The odorants were classified into groups of high, medium and low hydrophobicity, which largely determine their removal efficiencies in the technologies evaluated (Estrada et al., 2011). A typical release of CO2 of 0.75 g CeCO2 (g-Coxidized)1 due to the microbial oxidation of odorants and a biomass yield of 0.25 g C-biomass (g-Coxidized)1 were considered. Water was supplied to the BF, BTF and CS in order
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79
Table 1 Summary of the inventory data for the non treatment scenario and for the five abatement technologies evaluated. The units are referred to the FU: 1 m3 of treated air. NT
BF
BTF
AC
BTF þ AC
CS
Discharge of air 1-Ethyl-2-methylbenzene (mg) 1,2,4-Trimethylbenzene (mg) 1,3,5-Trimethylbenzene (mg) 2-Butoxyethanol (mg) a-Pinene (mg) Acetic acid (mg) Acetone (mg) Acetophenone (mg) Benzaldehyde (mg) Benzene (mg) Butanone (mg) Butyric acid (mg) D-limonene (mg) Decane (mg) Dimethyl disulfide (mg) Dodecane (mg) Ethylbenzene (mg) Hydrogen sulfide (mg) Limonene (mg) m-Xylene (mg) Methanethiol (mg) Methylcyclohexane (mg) Nonane (mg) o-Xylene (mg) Octane (mg) p-Xylene (mg) Propanoic acid (mg) Tetrachloroethylene (mg) Tetradecane (mg) Toluene (mg) Tridecane (mg) Undecane (mg) Carbon dioxide (mg)
1.31 1.06 8.36 9.63 1.31 1.05 4.62 5.88 6.70 2.19 4.54 2.09 2.30 7.08 2.13 3.41 1.47 20.91 1.15 1.77 1.97 1.26 1.07 1.77 1.07 1.77 1.13 1.07 4.76 5.09 4.04 5.37 0
Infrastructure material Concrete (m3) Glass fibre (kg)
e e
9.33 109 e
e 1.48 105
e 3.23 105
e 2.13 105
e 4.83 105
Packing material Compost (kg) Perlite (kg) Polyurethane foam (kg) Activated carbon (kg) Intalox saddles (kg)
e e e e e
1.31 104 4.36 105 e e e
e e 4.76 106 e e
e e e 8.18 105 e
e e e e 1.27 106
e e 2.85 106 2.05 105 e
Chemicals NaOH solution 50% (kg) NaClO solution 25% (kg) HCl solution 5% (kg)
e e e
e e e
e e e
e e e
9.99 105 6.35 104 1.82 105
e e e
Water From WWTP secondary effluent Potable water
e e
2.68 102 e
5.23 102 e
e e
e 1.25 102
5.23 102 e
Electricity From the grid (kWh)
e
3.95 104
2.40 104
6.80 104
5.23 104
9.00 104
Transport Lorry 3.5e7.5 t (kg km)
e
3.72 102
3.52 103
6.05 102
2.74 104
6.27 102
Final disposal Composting (kg) Landfill deposits (kg)
e e
1.74 104 e
e 4.76 106
e 8.18 105
e 8.25 106
e 2.33 105
102 102 103 102 102 101 101 101 102 102 102 102 103 101 103 102 101 102 102 102 102 102 102 102 102 103 101 103 103
6.55 5.28 4.18 9.63 3.29 1.05 4.62 5.88 6.70 1.10 4.54 2.09 5.76 1.77 1.06 8.53 7.36 2.09 2.87 8.84 1.97 3.14 2.68 8.84 2.68 8.84 1.13 2.66 1.19 2.55 1.01 1.34 14.19
104 104 104 104 103 103 103 103 104 103 102 104 103 103 102 104 104 101 102 104 102 103 103 104 103 104 104 103 103 102 103 103
to maintain the moisture conditions required and to compensate for evaporation losses. The pressure drop associated with each technology determined the energy required for gas circulation according to the Eq. (1):
E ¼ Qgas $DP$BE1
(1)
where E is the energy needed (kWh), Qgas is the gas flow (m3s1), DP is the pressure drop in the reactor (kPa) and BE is the blower
1.31 1.06 8.36 9.63 6.57 1.05 4.62 5.88 6.70 2.19 4.54 2.09 1.15 3.54 2.13 1.71 1.47 2.09 5.73 1.77 1.97 6.29 5.36 1.77 5.36 1.77 1.13 5.33 2.38 5.09 2.02 2.69 13.97
103 103 104 104 103 103 103 103 104 103 102 104 102 103 102 103 103 101 102 103 102 103 103 103 103 103 104 103 103 102 103 103
2.62 2.11 1.67 9.63 1.31 1.05 4.62 5.88 6.70 4.38 4.54 2.09 2.30 7.08 4.25 3.41 2.94 2.09 1.15 3.53 1.97 1.26 1.07 3.53 1.07 3.53 1.13 1.07 4.76 1.02 4.04 5.37 0
104 104 104 103 105 102 102 102 103 104 101 103 105 106 103 106 104 101 104 104 101 105 105 104 105 104 103 105 106 102 106 106
1.31 1.06 8.36 9.63 6.57 1.05 4.62 5.88 6.70 2.19 4.54 2.09 1.15 3.54 2.13 1.71 1.47 2.09 5.73 1.77 1.97 6.29 5.36 1.77 5.36 1.77 1.13 5.33 2.38 5.09 2.02 2.69 18.63
103 103 104 104 103 103 103 103 104 103 102 104 102 103 102 103 103 101 102 103 102 103 103 103 103 103 104 103 103 102 103 103
2.62 2.11 1.67 9.63 6.57 1.05 4.62 5.88 6.70 4.38 4.54 2.09 1.15 3.54 4.25 1.71 2.94 2.09 5.73 3.53 1.97 6.29 5.36 3.53 5.36 3.53 1.13 5.33 2.38 1.02 2.02 2.69 13.97
105 105 105 105 106 104 104 104 105 105 103 105 105 106 104 106 105 101 105 105 103 106 106 105 106 105 105 106 106 103 106 106
efficiency, in which case a typical value of 0.7 was assumed (Estrada et al., 2011). 2.3.1. Biofilter (BF) This configuration consisted of a reactor packed with a mixture of organic and inorganic materials: 75% compost and 25% perlite (v/ v) (Lebrero et al., 2010). The BF was built in standard concrete with a volume of 694 m3 (1 m height) and operated at an EBRT of 50 s. Irrigation water (129 kg h1) from the secondary effluent of the WWTP was used. A standard pressure drop of 1000 Pa was
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considered. The malodorous air was moistened in a humidifier prior to biofiltration with additional water (1211 kg h1). The leachate produced from the BF and the excess water from the humidifier were returned to the WWTP headworks for further treatment. The mixture of compost and perlite was replaced after 2 years of operation and processed in a composting facility, since its characteristics are similar to those of an average organic waste like sludge. 2.3.2. Biotrickling filter (BTF) A two-stage biotrickling filter of 208 m3 with a total EBRT of 15 s was selected. The glass fibre BTF was packed with inert plastic supports, which exhibited a density of 100 kg m3 and a total pressure drop of 500 Pa (Dorado et al., 2009). In the first stage, the pH was maintained at 2 for the efficient removal of H2S and in a second stage a neutral pH was kept to favour the removal of the other odorants. Water was supplied by spray nozzles and recirculated at a rate of 7.2 L m3 min1 using a renewal ratio of 2.5 L of water added per gram of H2S removed. Landfills were selected as the disposal scenario for the packed-bed after 10 years of lifespan.
2.3.3. Chemical scrubber (CS) A two-stage NaOHeNaClO scrubber (2 m) packed with Intalox Saddles (density of 100 kg m3) was considered, operated with a total pressure drop of 500 Pa. The glass fibre reactor was operated at an EBRT of 4 s (2 s per stage), which entailed a reactor volume of 55.6 m3. The liquid phase was recirculated at a rate of 180 L m3 min1 via spray nozzles (with a pressure drop of 50 kPa) (Gabriel and Deshusses, 2004) while 15040 L of additional water were added daily. The pH was maintained at 10e12 in the first stage, whereas the second stage was operated at a slightly lower pH (8.5e9.5). NaClO (25% w/w) at a flow rate of 28.85 L h1 and NaOH (50% w/w) at a flow rate of 3.26 L h1 were continuously supplied. The packing material was washed once a year with HCl (5% w/w) in order to avoid the accumulation of precipitates in the medium. The packing material was disposed of in a landfill after 10 years of operation. 2.3.4. Activated carbon filter (AC) Two glass fibre towers of 42 m3 packed with granular impregnated activated carbon (density of 430 kg m3) and operating at an EBRT of 3 s and a pressure drop of 1750 Pa were used as a model adsorption filter. One filter was in operation while the second one was in stand-by for the periodic renewal of the activated carbon every 6 months, which was later disposed of in landfills. Regeneration of the activated carbon was not considered according to the current operation of most full scale facilities (Estrada et al., 2011, 2012). 2.3.5. Biotrickling filter coupled with an activated carbon tower (BTF þ AC) The hybrid technology consisted of a biotrickling filter coupled with an activated carbon unit. The biotrickling filter, packed with 2 m of inert plastic support (density of 100 kg m3 and pressure drop of 500 Pa), was operated at an EBRT of 9 s. The 42 m3 AC tower was operated at an EBRT of 3 s and pressure drop of 1750 Pa. The main difference between the AC coupled to a biotrickling filter and the AC operated stand-alone was the lifetime of the activated carbon in the combined configuration: 2 years, due to the lower odorant concentrations in the inlet stream. The packing materials were identical to the previous configurations and consequently, landfills were considered for disposal.
2.4. Life cycle inventory, data quality and simplifications The LCI included the emission data reported by Zarra et al. (2008) as well as theoretical and literature data for each technology concerning the odorant concentrations in the outlet emission, infrastructure material, packing bed material, energy requirements, chemicals use, cleaning agents, water consumption, transport distances and final disposal of wastes (Table 1). The inventory data for auxiliary processes were taken from the Ecoinvent database and completed with recent bibliography (Table 2). The Spanish profile for energy was adapted with data of the electricity production/importation system for Spain in 2011 (RedElectrica, 2011). The electricity transmission network, emissions of sulphur hexafluoride to air, and losses during the medium voltage transmission and transformation from high to medium voltage were also considered (Dones et al., 2007). The long term stability of the technologies was indirectly estimated by the lifespan of the packing materials throughout the operation of the systems, which was applied to calculate packing material requirements per FU. The production of Intalox Saddles for the CS configuration was modelled according to the process of polyvinyl chloride (PVC) production (Hischier, 2007). Granular activated carbon (GAC) from coal was considered in the AC system ~ oz, 2006). The amount of infrastructure material (concrete or (Mun glass fibre) was calculated assuming a wall thickness of 15 cm of concrete and 15 mm of glass fibre. The production processes of NaOH, NaClO and HCl were taken from the Ecoinvent database and adapted to the concentrations and consumptions used in this case study (Althaus et al., 2007). Diaphragm cells were considered for the production of NaOH (50%w/w), since it is the most common process applied (Althaus et al., 2007). The production processes of NaClO (25%w/w) and HCl (5%w/w) were estimated from process data of 15%w/w and 30%w/w solutions, respectively (Althaus et al., 2007). Finally, the production of 1 kg of compost from 1 kg of packing material was assumed within the composting process. Inventory data considered in the LCI were related to the steadystate operation of the systems. Therefore, the evaluation of the
Table 2 Summary of data sources considered in this study. Inventory inputs Energy Infrastructure material
Packing material
Chemicals
Water Transport Waste treatment
Data sources Electricity (Spanish electricity profile) Concrete
Ecoinvent database (Dones et al., 2007) Ecoinvent database (Kellenberger et al., 2007) Glass fibre Ecoinvent database (Kellenberger et al., 2007) Compost Ecoinvent database (Nemecek and K€ agi, 2007) Expanded perlite Ecoinvent database (Kellenberger et al., 2007) ~ oz (2006) Granular activated carbon Mun Polyurethane, rigid foam Ecoinvent database (Hischier, 2007) PVC Ecoinvent database (Hischier, 2007) Sodium hydroxide (50%) Ecoinvent database (Althaus et al., 2007) Sodium hypochlorite (15%) Ecoinvent database (Althaus et al., 2007) Hydrochloric acid (30%) Ecoinvent database (Althaus et al., 2007) Tap water Ecoinvent database (Althaus et al., 2007) Truck 3.5e7.5 t, EURO 4 Ecoinvent database (Spielmann et al., 2007) Landfill deposits Ecoinvent database (Doka, 2007)
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robustness of the technologies towards process fluctuations and operational upsets was out of the scope of the manuscript. For a detailed evaluation of the robustness of the most common odour treatment technologies see Estrada et al. (2012). 2.5. Impact categories In the LCIA phase, the role of each element of the LCI was quantitatively assessed in terms of its contribution to nine impact categories: i) Cumulative Energy Demand (CED), which accounts for direct and indirect energy use throughout the entire life cycle of a product or process (Huijbregts et al., 2010); ii) malodorous air, defined as the quantity of contaminated air exceeding the odour threshold value established for each compound (Heijungs et al., 1992) and evaluated using the methodology developed by the Centre of Environmental Science of Leiden University (CML, 2001) e et al., 2002), and; iii) climate change, freshwater eutro(Guine phication, photochemical oxidant formation, human toxicity, and marine, terrestrial and freshwater ecotoxicity assessed with the ReCiPe Midpoint method using the Hierarchist Perspective (H) version 1.05 (Goedkoop et al., 2009). The software SimaPro 8 was used for the computational implementation of all the inventories (http://www.pre.nl/simapro). 3. Results 3.1. Malodorous air The impact category of malodorous air is of major relevance in this research study. The impact of malodorous air in the NT scenario was very high: 55-times higher than the impact related to the worst case scenario (BF): 48728 m3 m3 for NT vs. 894 m3 m3 for BF. The treatment of packing material in composting facilities entailed the highest impacts for BF in this category. At this point, it must be stressed that the potential of malodorous air was reduced by 98% by BF and 99% with other alternative technologies. Only the results related to the five technological alternatives were plotted in Fig. 1 based on the large differences between the NT scenario and the impacts associated with the different technologies. The main impact associated with the technology performance was that corresponding to the treated effluent released to the atmosphere. The treated gas accounted for values of malodorous air ~95% in BTF, AC, CS and BTF þ AC, whereas odorants associated with the packing material production and disposal contributed to ~50% of the impact in BF (Fig. 1). Among the odorants released to the atmosphere, H2S was found the main responsible of the malodorous air impact related to direct emissions.
Fig. 1. Malodours air impact for the five technologies evaluated. Malodorous air of NT: 48728 m3 per FU.
Fig. 2. Cumulative energy demand (CED) for the five technologies evaluated and NT scenario.
3.2. Cumulative energy demand (CED) No cumulative energy demand was computed in the NT scenario. Therefore, the potential information that this indicator provides is only useful when identifying the most energy efficient technology. The lowest CED impacts were obtained for BF and BTF, where this parameter was mainly associated with the use of energy for air and water pumping (Fig. 2). The impact associated with AC was mainly due to the packing material production, accounting for 76% of the overall impact. Both GAC production and energy use represented the impact associated with the hybrid technology, higher than those of biological techniques but ~50% lower than AC potential impact. Finally, the production of chemicals used in the CS contributed with 72% to the global CED impact. 3.3. Climate change The odorants present in the target malodorous emission do not contribute directly to this category and, consequently, climate change impact in the NT scenario was considered negligible. Moreover, although the oxidation of carbon-based odorants implies the emission of CO2, this contribution was deemed negligible when compared to other CO2 sources such as the indirect emissions derived from packing material or chemicals use (Fig. 3). Biological treatments exhibited the lowest climate change impacts. In fact, BTF and BF presented impacts ~91% and 84% lower than the impact associated with AC (Fig. 3). Process energy requirements in BTF and BF were the main contributors to climate change, while GAC production and transportation was the main responsible of its climate change potential in AC. The use of chemicals in CS accounted for
Fig. 3. Climate change impact for the five technologies evaluated and NT scenario.
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approximately 80% of the total impact, with NaClO presenting the largest contribution. The hybrid technology presented lower impact than physical/chemical technologies but higher than biological alternatives. 3.4. Freshwater eutrophication This impact category is not directly applicable to the NT scenario as the odorants present in the raw emission do not contribute to water eutrophication. CS presented the highest eutrophication impact due to the eutrophying emissions related to the production of chemicals required (Fig. 4). In this particular case, NaClO production accounted for 88% of the impact associated with chemicals. Biological treatments presented the lowest impacts, with electricity use as the main contributor in BF (91%) and BTF (55%). Similarly, energy use accounted for up to 50% of the eutrophication impact in the hybrid system. Finally, GAC production in AC represented 63% of the eutrophication potential. 3.5. Photochemical oxidant formation A total impact of 3.78 106 kg NMVOC per FU was calculated in the NT scenario as a result of the direct release of VOCs to the atmosphere (Fig. 5). The most important impacts were observed for physical/chemical alternatives, followed by BTF þ AC and finally, biological technologies. Packing material accounted for 56% of the total impact in AC whereas chemicals production entailed remarkable impact in CS. Significantly lower impacts compared to the NT scenario of 38%, 77% and 80% were estimated in BTF þ AC, BF and BTF, respectively. The photochemical oxidant formation impact related to the production of energy was considerable in all technologies evaluated. Finally, it is worth noting that the use of glass fibre as a building material resulted in higher impact compared to the use of concrete. 3.6. Ecotoxicity and human toxicity The toxicity associated with the treatment of the target malodorous emission was assessed in freshwater, marine and terrestrial environments and humans (Fig. 6 aed). Biological treatments presented the most environmentally friendly performance, with energy use appearing as the main responsible of the impact associated with BF. Production of glass fibre and disposal in landfills entailed the largest impacts in BTF. The chemicals added to CS (especially NaClO) induced the largest impacts regardless of the toxicity category. The contributors to toxicity-associated impacts in
Fig. 4. Freshwater eutrophication impact for the five technologies evaluated and NT scenario.
Fig. 5. Photochemical oxidant formation of the five technologies evaluated and NT scenario.
AC and BTF þ AC depended to a great extent on the categories analyzed, disposal of packing material, energy use and glass fibre production contributing significantly to freshwater and marine ecotoxicity, terrestrial ecotoxicity and human toxicity, respectively. Thus, although direct odorant emissions caused a negative impact on the receiving environment and on human health, this impact was considered negligible when compared to the indirect impacts related to energy consumption, the production of the packing material or the bed material disposal in landfills. 4. Discussion Overall, biological processes exhibited an improved environmental performance, with the exception of the malodorous air category, where emissions from the composting process led to more significant impacts in BF. In this context, it must be stressed that although the conversion of organic wastes to compost reduces their potential impacts to the environment, composting entails negative environmental impacts, such as the emission of H2S €gi, 2007). The main contributor to malodorous air (Nemecek and Ka associated with the technologies were direct emissions, H2S appearing as the most significant odorant compound. Firstly, the elevated H2S concentration in the raw emission (2.09 101 mg H2S m3) resulted in a significantly high concentration of H2S in the outlet stream despite the great H2S removal efficiency (99%) achieved in all the technologies assessed. Secondly, this compound presents a characterisation factor of 2.33 109 m3 per kg of H2S, which is calculated as the inverse of the Odour Threshold Value (OTV). H2S exhibits a significantly low OTV (4.3 104 mg m3), which entailed negative impacts even at trace level concentrations. Treatment of odorants entailed a reduction of 98e99% in the malodorous impact category. The two biological treatments were found to be the friendliest alternatives to the environment, with diverse outcomes in the different impact categories. BF was the most suitable technology in terms of freshwater eutrophication, human toxicity, freshwater ecotoxicity and marine ecotoxicity, whereas the BTF exhibited the lowest impacts in terms of CED, climate change, photochemical oxidant formation and terrestrial ecotoxicity. Energy use was found to be the largest contributor to the impact in the BF scenario. More specifically, this consumption was directly related to the pressure drop associated with the packing material. On the other hand, waste disposal in landfills and glass fibre used in BTF significantly influenced its overall environmental impact, whereas concrete production was almost negligible in BF. CS presented the highest impacts in freshwater eutrophication, photochemical oxidant formation, human toxicity and ecotoxicity
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Fig. 6. Ecotoxicity in (a) freshwater, (b) marine and (c) terrestrial environments and (d) human toxicity associated with the five technologies evaluated and NT scenario.
due to the use of large amounts of chemicals. The significant amount of NaClO required for odorant oxidation had a key role in the impacts associated with the use of chemicals. The negative burdens derived from the use of chemicals in odour abatement technologies were also pointed out by Alfonsín et al. (2013) when assessing the environmental consequences for human toxicity and ecotoxicity of conventional biofilters at laboratory scale. Likewise, this high impact related to the use of chemicals in CS was in accordance with the results obtained by Estrada et al. (2011). Nevertheless, the authors considered the amount of reagents used per year, chemicals production being out of the analysis. The impact associated with AC was similar to that related to NT scenario in photochemical oxidant formation mainly due to packing material production, whereas the use of chemicals caused the highest impact in CS. Therefore, indirect impacts were found more important than the reduction of butanone directly released to the atmosphere, which represented ~75% of the impact associated with the NT scenario. On the other hand, AC showed the worst results for CED and climate change potentials followed by CS in both categories. As already mentioned, the most significant parameter determining the ~ oz, 2006), which AC impact was the GAC production process (Mun was mainly attributed to a large demand of natural gas and its replacement every 6 months. The contribution of the packing material to the environmental impact showed differences with the preliminary analysis conducted by Estrada et al. (2011), who reported that BF was the technology presenting the highest annual packed bed-material requirements. Nevertheless, LCA demonstrated that GAC production associated impacts were more relevant than those related to BF packing materials. Finally, the hybrid technology (BTF þ AC) showed a lower environmental performance than its biological counterparts, but
better results than physical/chemical technologies. BTF þ AC exhibited an improved environmental performance as compared to AC filtration due to the fact that the replacement frequency of GAC was reduced to 2 years instead of 6 months for stand-alone AC. Estrada et al. (2011), using the IChemE metrics, concluded that BTF þ AC was the less favourable option in terms of energy consumption. In this study, energy use in BTF þ AC was found to be significant in categories such as terrestrial ecotoxicity, but in no case was this contribution comparable to the global impact of physical/chemical technologies. Therefore, BTF þ AC could be ranked second after its biological counterparts, its implementation being highly recommended in odour sensitive scenarios due to its more reliable performance. 5. Conclusions Biological treatments exhibited the best environmental performance among the technologies evaluated, except in terms of malodorous air impact where BF showed the highest impact as a result of the contribution of compost production and disposal. Nonetheless, BF achieved a reduction in the malodour air impact of 98% compared to the NT scenario, which was slightly lower than the reduction achieved by the other technologies (99%). Even when the use of a technology implied more significant impacts than the NT scenario, BF and BTF showed an improved environmental performance compared to physical/chemical technologies. In this context, AC and CS presented the highest impacts due to the use of GAC as packing material, chemical use and waste disposal in landfills. Finally, the operation of BTF þ AC was more recommendable from an environmental sustainability perspective than physical/chemical treatments, but slightly worse than biotechnologies. In other words, based on the fact that odour treatment is either compulsory
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to comply with, based on the legislative thresholds, or highly recommended, in order to avoid public complaints, biotechnologies are the best treatment options from an environmental point of view. Acknowledgements The authors acknowledge the financial support of the Spanish Ministry of Economy and Competitiveness (project CTQ2012-34949 and contract BES-2010-030994) and the European Union (LIVEWASTE project: LIFE12 ENV/CY/544). The authors from the University of Santiago de Compostela belong to the Galician Competitive Research Group GRC 2013-032, programme co-funded by FEDER. References Alfonsín, C., Hernandez, J., Omil, F., Prado, O.J., Gabriel, D., Feijoo, G., Moreira, M.T., 2013. Environmental assessment of different biofilters for the treatment of gaseous streams. J. Environ. Manage. 129, 463e470. Althaus, H.J., Chudacoff, M., Hischier, R., Jungbluth, N., Osses, M., Primas, A., 2007. Life Cycle Inventories of Chemicals. Ecoinvent Report no. 8. Swiss Centre for Life Cycle Assessment, Dübendorf. Balasubramanian, P., Philip, L., Murty Bhallamudi, S., 2012. Biotrickling filtration of VOC emissions from pharmaceutical industries. Chem. Eng. J. 209, 102e112. Baumann, H., Tillman, A.M., 2004. The Hitch Hiker's Guide to LCA : an Orientation in Life Cycle Assessment Methodology and Application. Studentlitteratur, Lund. Doka, G., 2007. Life Cycle Inventories of Waste Treatment Services. Ecoinvent report No. 13. Swiss Centre for Life Cycle Inventories, Dübendorf. Dones, R., Bauer, C., Bolliger, R., Burger, B., Faist Emmenegger, M., Frischknecht, R., € der, A., Tuchschmid, M., 2007. Life Cycle Inventories of Heck, T., Jungbluth, N., Ro Energy Systems: Results for Current Systems in Switzerland and Other UCTE Countries. Ecoinvent report no. 5, Paul Scherrer Institut Villigen. Swiss Centre for Life Cycle Assessment, Dübendorf, CH. Dorado, A.D., Rodríguez, G., Ribera, G., Bonsfills, A., Gabriel, D., Lafuente, J., Gamisans, X., 2009. Evaluation of mass transfer coefficients in biotrickling filters: experimental determination and comparison to correlations. Chem. Eng. Technol. 32, 1941e1950. Estrada, J.M., Kraakman, N.J., Munoz, R., Lebrero, R., 2011. A comparative analysis of odour treatment technologies in wastewater treatment plants. Environ. Sci. Technol. 45, 1100e1106. ~ oz, R., 2012. A sensitivity analysis of Estrada, J.M., Kraakman, N.J.R., Lebrero, R., Mun process design parameters, commodity prices and robustness on the economics of odour abatement technologies. Biotechnol. Adv. 30, 1354e1363. Gabriel, D., Deshusses, M.A., 2004. Technical and economical analysis of the conversion of a full-scale scrubber to a biotrickling filter for odor control. Water Sci. Technol. 50, 309e318. Goedkoop, M., Heijungs, R., Huijbregts, M., De Schryver, A., Struijs, J., van Zelm, R., 2009. ReCiPe 2008 e a Life Cycle Impact Assessment Method Which comprises harmonised Category Indicators at the Midpoint and the Endpoint Level. Ministry of Housing, Spatial Planning and Environment (VROM), The Netherlands. e, J.B., Gorre e, M., Heijungs, R., Huppes, G., Kleijn, R., Koning, A.d., Oers, L.v., Guine Wegener Sleeswijk, A., Suh, S., Udo de Haes, H.A., Brujin, H.d., va., D.R., Huijbregts, M.A.J., 2002. Handbook on Life Cycle Assessment. Operational Guide to the ISO Standards. I: LCA in Perspective. IIa: Guide. IIb: Operational Annex. III: Scientific Background. Kluwer Academic Publishers, Dordrecht, ISBN 1-40200228-9. e, J.B., Huppes, G., Lankreijer, R.M., Udo de Haes, H.A., Wegener Heijungs, R., Guine Sleeswijk, A., Ansems, A.M.M., Eggels, P.G., Duin, R.v., Goede, H.P.d., 1992.
Environmental life cycle assessment of products. I: Guide - October 1992. II: Backgrounds - October 1992. CML (NOH Report 9266 þ 9267), ISBN 90-5191064-9 (Leiden). Hischier, R., 2007. Life Cycle Inventories of Packaging and Graphical Papers. Ecoinvent-Report No. 11. Swiss Centre for Life Cycle Inventories, Dübendorf. ndez-Couto, M., Feijoo, G., 2004. Environmental Hospido, A., Moreira, M., Ferna performance of a municipal wastewater treatment plant. Int. J. Life Cycle Assess. 9, 261e271. Hospido, A., Sanchez, I., Rodriguez-Garcia, G., Iglesias, A., Buntner, D., Reif, R., Moreira, M.T., Feijoo, G., 2012. Are all membrane reactors equal from an environmental point of view? Desalination 285, 263e270. Huijbregts, M.A., Hellweg, S., Frischknecht, R., Hendriks, H.W., Hungerbuhler, K., Hendriks, A.J., 2010. Cumulative energy demand as predictor for the environmental burden of commodity production. Environ. Sci. Technol. 44, 2189e2196. IChemE, 2002. The Sustainability Metrics. The Institution of Chemical Engineers, Rugby, UK. ISO, 2006a. ISO 14040: 2006. Environmental Management e Life Cycle Assessment e Principles and Framework. ISO, 2006b. ISO 14044: 2006. Environmental Management - Life Cycle Assessment Requirements and Guidelines. Kellenberger, D., Althaus, H.J., Jungbluth, N., Künniger, T., 2007. Life Cycle Inventories of Building Products. Final report Ecoinvent data v2.0 No. 7. Swiss Centre for Life Cycle Inventories, Dübendorf, Switzerland. Kellenberger, D., Althaus, H.J., Jungbluth, N., Künniger, T., 2007. Life Cycle Inventories of Building Products. Final report Ecoinvent data v2.0 No. 7. Swiss Centre for Life Cycle Inventories, Dübendorf, Switzerland.Lebrero, R., rez, R., García-Encina, P.A., Mun ~ oz, R., 2014. Comparative Gondim, A.C., Pe assessment of a biofilter, a biotrickling filter and a hollow fiber membrane bioreactor for odour treatment in wastewater treatment plants. Water Res. 49, 339e350. ~ oz, R., 2012. Odor Lebrero, R., Rodríguez, E., Estrada, J.M., García-Encina, P.A., Mun abatement in biotrickling filters: effect of the EBRT on methyl mercaptan and hydrophobic VOCs removal. Bioresour. Technol. 109, 38e45. ~ oz, R., 2010. H2S and Lebrero, R., Rodríguez, E., Martin, M., García-Encina, P.A., Mun VOCs abatement robustness in biofilters and air diffusion bioreactors: a comparative study. Water Res. 44, 3905e3914. ~ oz, I., 2006. Life Cycle Assessment as a Tool for Green Chemistry: Application to Mun Different Advanced Oxidation Processes for Wastewater Treatment. Doctoral Thesis. Autonomic University of Barcelona. Nemecek, T., K€ agi, T., 2007. Life Cycle Inventories of Swiss and European Agricultural Production Systems. Final report ecoinvent V2.0 No. 15a. Agroscope Reckenholz-Taenikon Research Station ART, Swiss Centre for Life Cycle Inventories, Zurich and Dübendorf, CH. Omri, I., Aouidi, F., Bouallagui, H., Godon, J.-J., Hamdi, M., 2013. Performance study of biofilter developed to treat H2S from wastewater odour. Saudi J. Biol. Sci. 20, 169e176. ctrica de Espan ~ a. El sistema ele ctrico espan ~ ol (in RedElectrica, 2011. Red Ele Spanish). Revah, S., Morgan-Sagastume, J., 2005. Methods of odor and VOC control. In: Shareefdeen, Z., Singh, A. (Eds.), Biotechnology for Odor and Air Pollution Control. Springer, Berlin Heidelberg, pp. 29e63. Rodriguez-Garcia, G., Molinos-Senante, M., Hospido, A., Hernandez-Sancho, F., Moreira, M.T., Feijoo, G., 2011. Environmental and economic profile of six typologies of wastewater treatment plants. Water Res. 45, 5997e6010. Schlegelmilch, M., Streese, J., Stegmann, R., 2005. Odour management and treatment technologies: an overview. Waste Manage. 25, 928e939. Shaw, A.R., Koh, S.-H., 2013. Gaseous emissions from wastewater facilities. Water Environ. Res. 85, 1298e1307. Spielmann, M., Bauer, C., Dones, R., Tuchschmid, M., 2007. Transport Services. Ecoinvent report No. 14. Swiss Centre for Life Cycle Inventories, Dübendorf. Zarra, T., Naddeo, V., Belgiorno, V., Reiser, M., Kranert, M., 2008. Odour monitoring of small wastewater treatment plant located in sensitive environment. Water Sci. Technol. 58, 89e94.