Journal of Cleaner Production 354 (2022) 131734 Contents lists available at ScienceDirect Journal of Cleaner Production journal homepage: www.elsevier.com/locate/jclepro Increasing the energy production in an urban wastewater treatment plant using a high-rate activated sludge: Pilot plant demonstration and energy balance Julián Carrera a, *, Oriol Carbó a, b, Silvia Doñate b, María Eugenia Suárez-Ojeda a, Julio Pérez a a Departament d’Enginyeria Química, Biològica i Ambiental, Escola d’Enginyeria, Universitat Autònoma de Barcelona, 08193, Bellaterra, Barcelona, Spain b Depuración de Aguas del Mediterráneo (DAM), Av. Benjamín Franklin, 21, 46980, Paterna, Valencia, Spain A R T I C L E I N F O A B S T R A C T Handling Editor: Baoshan Huang The use of high-rate activated sludge (HRAS) reactors for the removal of COD in urban wastewater treatment plants (WWTPs) has been investigated because its potential contribution to energy generating WWTPs. A one- Keywords: year operation period of a pilot plant treating the effluent of the primary settler of a full-scale WWTP was Biomethane potential analyzed. The HRAS pilot plant operated without iron salts addition at temperatures between 12 and 28 ◦ C at an Observed sludge yield coefficient average organic loading rate 2.8 ± 0.5 kg COD m− 3 d− 1 and with an average inflow COD concentration of 330 ± COD mass balance 86 mg O2 L− 1. The influence of sludge retention time (SRT) on COD recovery and biomethane potential of the Partial nitritation Autotrophic nitrogen removal produced sludge was investigated and compared to the full scale WWTP performance. The highest observed sludge yield coefficient and biomethane potential of the sludge were achieved at SRT of 0.6 days. The weak point of the HRAS performance at STR of 0.6 days is the high loss of organic matter in the effluent due to the limited efficiency of the solids separation in the secondary settler. At higher SRT (in the range 1.0–2.1 days), the sec­ ondary settler efficiency and the COD recovery are higher than those achieved at SRT of 0.6 days but part of the inlet ammonium can be nitrified in the HRAS system at temperatures above 20 ◦ C. A detailed energy balance indicated that two-thirds reduction of aeration requirements and one-third increase of biogas production could be achieved in a plant configuration in which HRAS is coupled to autotrophic biological nitrogen removal (BNR) compared to the heterotrophic BNR configuration, yielding a net energy production of ca. 0.1 kWh m− 3 of treated water. 1. Introduction considering an emission of 0.2307 kg CO2/kWh (European Environ­ mental Agency, 2022). Both, WWTP costs and greenhouse emissions are Currently, urban wastewater treatment plants (WWTPs) are net of great concern for stakeholders and the development of new technol­ energy-consumers systems with 8–16 kWh/person/year of electrical ogies able to reduce them is of paramount importance in sustainability consumption depending on the type of treatment (removal of only context (Cardoso et al., 2021; Khalkhali and Mo, 2020). organic matter or removal of organic matter and nutrients and consid­ Urban WWTPs can become zero-energy consuming systems ering the energy recovery in form of biogas from the digestion of sewage combining the chemical energy recovered by anaerobic digestion with sludge) being the conventional nitrification/denitrification the most the thermal energy recovered from the treated wastewater through heat consuming process of the whole plant (Capodaglio and Olsson, 2020; pumps and the solar energy obtained with photovoltaic systems (Liu Kartal et al., 2010). In the EU, with a population of 447 million persons et al., 2021). Also, urban WWTPs can approach to the energy neutrality (Eurostat, 2021a), this would mean an energy consumption of ca. with aeration upgrades and using only the chemical energy recovered by 3576–7152 GWh/year for treating wastewater. Considering an average anaerobic digestion if the sewage sludge is co-digested with other price of 0.1283 €/kWh for non-household consumers in the EU (Euro­ external organic waste (Macintosh et al., 2019; Torretta et al., 2021). stat, 2021b), this energy consumption represents 459–918 M€/year, However, some theoretical studies suggested that the implementation of with an equivalent associated emission of ca. 1–2 Mtons CO2/year the autotrophic biological nitrogen removal (BNR) in the mainstream of * Corresponding author. E-mail address: julian.carrera@uab.es (J. Carrera). https://doi.org/10.1016/j.jclepro.2022.131734 Received 16 February 2022; Received in revised form 25 March 2022; Accepted 10 April 2022 Available online 12 April 2022 0959-6526/© 2022 The Author(s). Published by Elsevier Ltd. This is an open access article under the CC BY-NC license (http://creativecommons.org/licenses/by- nc/4.0/). J. Carrera et al. Journal of Cleaner Production 354 (2022) 131734 urban WWTPs would transform the current net-energy consumer 2. Materials and methods WWTPs into self-sufficient or even net-energy producer facilities only using the chemical energy recovered by anaerobic digestion of the This section is divided in two main blocs. The first one describes the sewage sludge (Jetten et al., 1997; Kartal et al., 2010). This kind of experimental set-up and the analytical methods used in the experimental technology is based in a two-stage biological process. The first stage is a part of this study. The second one describes all the calculations needed high-rate activated sludge (HRAS) system for organic matter redirection. to make the mass and energy balances. A HRAS system is a process modification of a conventional activated sludge (CAS) system. This process modification consists on operating at 2.1. Experimental set-up and analytical methods shorter hydraulic and sludge retention times (HRT and SRT, respec­ tively) than those commonly applied in CAS systems (Tchobanoglous 2.1.1. HRAS pilot plant description et al., 2003). This modification results in lower biomass concentration The HRAS pilot plant was comprised of an aerobic biological reactor and higher volumetric and specific organic loading rates in the HRAS followed by a sedimentation tank for solid-liquid separation (Fig. 1). The system. The biomass produced in this step would have, therefore, a very working volumes were 250 and 470 L, respectively. Mixed liquor from favourable bio-production of methane, greater than the achieved in the biological reactor flowed by gravity to the sedimentation tank, current urban WWTPs (Rahman et al., 2020). The second stage uses where solid-liquid separation took place. Dissolved oxygen (DO) con­ autotrophic BNR for removing nitrogen in the mainstream without the centration was controlled by means a closed control-loop that use a need of organic matter (Juan-Díaz et al., 2021, 2022; Reino et al., 2016, Hach LDO sensor (Hach Lange, Spain) and an automatic control valve to 2018). The application of autotrophic BNR to the mainstream would regulate the airflow rate. Temperature and pH were also monitored on- reduce severely the aeration costs compared to current urban WWTPs line. Influent and recirculation flow rates were also controlled by means where classical heterotrophic BNR (het-BNR) is applied (Kartal et al., a closed control-loop that used mass flowmeters and helical rotor pumps 2010; Pérez et al., 2015). to obtain the desired flow rate. Excess sludge flow rate was controlled by In the HRAS process, a high fraction of the organic matter is removed two automatic on-off valves located in the recirculation line, which through a physical process, i.e. the adsorption of this organic matter on alternated the recirculation flow rate, either to the reactor or to the the sludge, preserving its energy recovery potential. However, the suc­ purge. cessful implementation of HRAS process is poorly understood and it is The pilot plant was located in the urban WWTP of Rubí-Valldoreix linked to the conditions imposed, such as, among others, the addition of (Barcelona, Spain) and it was inoculated with activated sludge from the iron salts for improvement of organic matter removal and sludge coag­ aerobic reactors of this facility. The mainstream line of the Rubí-Vall­ ulation/flocculation (de Graaff et al., 2016; de Vrieze et al., 2013). doreix WWTP is comprised by a pre-treatment, a primary sedimentation One straight ahead way of retrofitting an existing CAS urban WWTP and a CAS process with secondary sedimentation. The full-scale CAS for organic matter redirection using the HRAS process would be to treat process was operated at SRT, HRT and DO concentration of 14 days, 8 h the effluent of the primary settler in a biological reactor in which the and 2 mg O2 L− 1, respectively, throughout the period considered in this HRT and SRT applied are short. Indeed, it is generally assumed that study. The sidestream of the Rubí-Valldoreix WWTP includes a meso­ HRAS process is achieved by imposing HRT and SRT in the range philic anaerobic digester where primary and secondary sludge are 1.5–3.0 h and 0.5–2.0 days, respectively (Tchobanoglous et al., 2003). digested together. When HRAS process is intended to be connected to an autotrophic BNR, The temperature was not controlled on the HRAS pilot plant, and it the achievement of high organic matter removal is of paramount changed from 12 to 30 ◦ C depending on the season. In the HRAS pilot importance since it might compromise the anammox process itself (Li plant, the operation started with the highest SRT (2.1 ± 0.6 days) and et al., 2020; Pijuan et al., 2020) and because organic matter removal is finished with the lowest one (0.6 ± 0.2 days). To avoid the effect of uniquely performed in the HRAS process. other operational parameters on the performance of the HRAS pilot Here, an experimental study with a HRAS pilot plant, operated on- plant, HRT and DO concentration were maintained throughout the study site a full scale urban WWTP treating the effluent of the primary at 3 h and 2 mg O2 L.1, respectively. Iron salts were not added settler of this facility, is presented. The energy recovery with the HRAS throughout the experimental period. process has been assessed through the biomethane potential of the ob­ tained sludge and compared to the current energy production of the full scale WWTP. In this sense, the aim of this study is to compare the performance of a pilot HRAS, treating the effluent of a primary settler, with the perfor­ mance of an existing full-scale urban WWTP. Therefore the basic assumption is the retrofitting of existing WWTPs where the first settling step has been kept in the treatment scheme. The comparison of the re­ sults of a HRAS pilot plant with a full-scale CAS system treating the same influent is of interest since this comparison is not usually carried out in this kind of studies. Moreover, the use of the effluent of a primary settler as influent for the HRAS system is a significant difference with previous studies where the inlet of the HRAS system directly came from the pre- treatment of an urban WWTP (Cao et al., 2020; Ge et al., 2017; Jimenez et al., 2015; Kinya et al., 2017; Liang et al., 2022; Taboada-Santos et al., 2020). The main objectives of this study are: (i) to determine experimentally the influence of the SRT imposed in a HRAS pilot plant over the percent of COD removal and COD recovery, the sludge settleability and the nitrification process. (ii) To carry out a complete analysis of the imple­ mentation of a HRAS system in the mainstream of an urban WWTP combining the experimental data of the performance of the HRAS pilot Fig. 1. Process diagram of the HRAS pilot plant: 1. DO, pH and temperature plant, the fate of the organic matter and an energy balance. probes; 2. Automatic control valve; 3. Helical rotor pump; 4. Automatic on-off valve; PLC is a programmable logic controller. 2 J. Carrera et al. Journal of Cleaner Production 354 (2022) 131734 2.1.2. Sampling and analysis was periodically determined by using gas chromatography. Accumu­ Samples were collected at least once a week from the influent of the lated volumetric gas production was calculated from the pressure in­ pilot plant, the biological reactor, the sludge purge and the secondary crease in the headspace volume and expressed under standard effluent. Mixed liquor total suspended solids (TSS), mixed liquor volatile conditions (25 ◦ C, 1 atm) so that the results were able to be compared suspended solids (VSS) and sludge volumetric index (SVI) were with other studies. measured following the recommendations and procedures published in Standard Methods (APHA, 1999). Total Chemical Oxygen Demand 2.2. Calculations (tCOD) and ammonia nitrogen (N–NH4+) analyses were conducted using the Hach Test-in-Tube (TNT) kits (Hach Lange, Spain). The tCOD was 2.2.1. Calculation of the SRT, SVI, efficiency of the secondary settler and assumed to be composed by particulate (pCOD), colloidal (cCOD) and observed sludge yield coefficient soluble (sCOD) fractions. The sCOD fraction was measured using the SRT was measured by dividing the amount of VSS in the HRAS flocculation-filtration (ZnSO4 as flocculant, 0.45 μm) method of Mamais reactor by the VSS washed out with the effluent and the purge (Equation et al. (1993). The cCOD was measured as the COD fraction that passed (1)). through a 1.5 μm filter, from which the sCOD fraction was subtracted. Vreactor ⋅[VSS]reactor The pCOD was calculated by subtracting COD filtered through 1.5 μm SRT = (1) Qpurge ⋅[VSS]purge + Qeffluent ⋅[VSS]effluent from tCOD. Nitrite and nitrate analyses were conducted using the ionic chro­ where SRT = sludge retention time (d); Vreactor = reactor volume (L); matography ICS-2000 Integrated Reagent-Free IC system (DIONEX [VSS]reactor = VSS concentration in the reactor (g VSS L− 1), Qpurge = Corporation, USA). DO concentration, pH and temperature data of the purge flow rate (L d− 1), [VSS]purge = VSS concentration in the excess HRAS system were online recorded throughout the whole experimental sludge (g VSS L− 1), Qeffluent = effluent flow rate (L d− 1) and [VSS]effluent period. = VSS concentration in the effluent (g VSS L− 1). Sludge Volumetric Index (SVI) was measured as described in Equa­ 2.1.3. HRAS influent characteristics tion (2): The HRAS pilot plant was fed with the effluent of the primary settlers of the full-scale WWTP. The average composition of the HRAS influent SV30 SVI = (2) wastewater was as described in Table 1. [VSS]reactor 2.1.4. Determination of biomethane potential at standard temperature and where SV30 = settled sludge volume at 30 min (mL L− 1) and [VSS]reactor pressure (BMP at STP) = VSS concentration in the reactor (g VSS L− 1). BMP at STP tests, from now on, only BMP, were used to assess the Secondary settler efficiency (EFSS) was measured as described in amount of biogas produced by the anaerobic digestion of sludge pro­ Equation (3): duced in the HRAS pilot plant and the full-scale CAS system. The BMP [VSS]effluent tests were carried out by following the procedure proposed by Angel­ EFSS = ⋅100 (3) [VSS]reactor idaki et al. (2009). Control experiments (i.e. without added substrate) were included to recalculate the BMP accounting for residual biogas where [VSS]reactor = VSS concentration in the reactor (g VSS L− 1) and production. Control experiments were performed in triplicate, whereas [VSS]effluent = VSS concentration in the effluent (g VSS L− 1). substrate added tests in quintuplicate. Identical 160 mL serum bottles Purge flow rate from the line of return activated sludge (RAS) was were used in all the tests. These bottles were filled up to 125 mL so that daily regulated to obtain the desired SRT. HRAS reactor walls were the headspace was 35 mL. Substrates used for the BMP tests were sec­ scraped twice per week to avoid the biofilm formation and the possible ondary sludge from the HRAS pilot plant operated at different SRTs and alteration of the applied SRT. The secondary settler was operated with a from the full-scale CAS system. The inoculum consisted of mesophilic minimal sludge blanket to avoid solids accumulation and the possible anaerobic sludge from the full-scale anaerobic digester treating the alteration of the applied SRT. sludge produced in the full-scale CAS system. The inoculum was The observed sludge yield coefficient (Yobs) was calculated for each degassed at 37 ◦ C for 3–5 days before its use. In order to neglect biogas operational period following Equation (4). production by endogenous respiration, the substrate to inoculum ratio of 0.5 (in terms of volatile solids, VS) was maintained in all the tests. The Qpurge ⋅[VSS]purge + Qeffluent ⋅[VSS]effluent (4) amount of substrate and inoculum used were 1 and 2 g L− 1 of VS Yobs = Qinfluent ⋅tCODinfluent − Qeffluent ⋅tCODeffluent respectively, per bottle. The pH of each bottle was set at 7.0 at the start of the experiments. The bottles were closed tightly with a butyl rubber where Yobs = observed sludge yield coefficient (kg VSS kg− 1 COD); septum and aluminium caps and were incubated under mesophilic Qpurge = volumetric flow rate in the purge (L d− 1), [VSS]purge = VSS conditions (37 ◦ C). Anaerobic conditions were achieved by purging the concentration in the purge (g VS L− 1), Qeffluent = volumetric flow rate in headspace of each bottle with N2 for 2 min. The content of each bottle the effluent (L d− 1); [VSS]effluent = VSS concentration in the effluent (g was mixed once per day. The gas production was estimated by VS L− 1); Qinfluent = volumetric flow rate in the influent (L d− 1); tCO­ measuring the pressure increase in the headspace with a gas pressure Dinfluent = total COD concentration in the influent; tCODeffluent = total meter. In each BMP test, the biogas composition (i.e. CH4, CO2 and H2) COD concentration in the effluent. The calculation considered that the produced sludge can be found by Table 1 adding the mass flow rates of VSS in the purge and effluent, while the HRAS influent characteristics. removed substrate was considered to be the difference between the mass flow rates of COD of the influent and the effluent. Component Units Concentration % of tCOD tCOD mg O2 L− 1 330 ± 86 2.2.2. Calculation of the COD mass balance for the HRAS pilot plant pCOD mg O2 L− 1 155 ± 11 47 cCOD mg O2 L− 1 66 ± 30 20 COD mass balance was calculated using the average values of each sCOD mg O2 L− 1 109 ± 28 33 operational period. Definition and related calculations for each term of N–NH4+ mg N L− 1 51 ± 11 - COD mass balance has been indicated in Table 2. N–NO2- mg N L− 1 1±1 - N–NO3- mg N L− 1 1±1 - 3 J. Carrera et al. Journal of Cleaner Production 354 (2022) 131734 2.2.3. Calculation of the energy balance Qinfluent = volumetric flow rate in the influent (m3 d− 1) [N–NH4+]consumed = ammonium concentration consumed in the Energy balances for two different configurations of the full-scale Assuming an empirical formula for the new biomass as: C5H7NO2, the following ratios can be obtained: 0.12 kg N kg− 1 Qinfluent = volumetric flow rate in the influent (m3 d− 1) tCODinfluent = COD concentration in the influent (kg O2 m− 3) Qinfluent = volumetric flow rate in the influent (m3 d− 1) tCODinfluent = COD concentration in the influent (kg O2 m− 3) Qinfluent = volumetric flow rate in the influent (m3 d− 1) tCODinfluent = COD concentration in the influent (kg O2 m− 3) Qeffluent = volumetric flow rate in the effluent (m3 d− 1) tCODeffluent = COD concentration in the effluent (kg O2 m− 3) WWTP were calculated using the experimental results obtained in this and other previous studies. The first configuration (CAS-hetBNR) is comprised by three process lines: (1) a water treatment line composed by a pre-treatment followed by a primary settler and finally a biological CAS process for heterotrophic BNR (a modified Luzdack-Ettinger pro­ cess with anoxic and aerobic zones); (2) a sludge treatment line, where the primary and secondary sludge are treated together in an anaerobic digester and (3) an energy production line, where the produced biogas is transformed into electrical energy. The second configuration is also comprised by three lines: (1) a water treatment line composed by a HRAS process for organic matter removal and an autotrophic BNR (partial nitritation + anammox in The ammonium oxidized to nitrate is not considered in this calculation. separated reactors) for nitrogen removal (HRAS-autBNR); (2) a sludge mCODADS = COD mass flow of adsorbed organic matter (kg O2 d− 1) treatment line, where the primary and secondary sludge are treated mCODOX = COD mass flow of oxidized organic matter (kg O2 d− 1) mCODOX = COD mass flow of oxidized organic matter (kg O2 d− 1) together in an anaerobic digester and (3) an energy production line, mCODX = COD mass flow of new biomass formed (kg O2 d− 1) tCODeffluent = COD concentration in the effluent (kg O2 m− 3) tCODeffluent = COD concentration in the effluent (kg O2 m− 3) where the produced biogas is transformed into electrical energy. For CODX = COD mass flow of new biomass formed (kg O2 d− 1) this second configuration (HRAS-autBNR), the energy balance was mCODINF = COD mass flow in the influent (kg O2 d− 1) mCODEFF = COD mass flow in the effluent (kg O2 d− 1) calculated for three different operating conditions of the HRAS process, corresponding to the three SRTs applied in the HRAS pilot plant (0.6, 1.0 and 2.1 days). Energy balance for HRAS-autBNR configuration aimed to provide what would be energy recovery in case the application Yobs = observed sludge yield (kg kg− 1) of partial nitritation-anammox assuming this treatment to be feasible, therefore they do not represent any evaluation on the feasibility of the autotrophic BNR. The definition and related calculations for each term of energy balance is presented in Table 3. It should be point out that this VS and 1.42 kg O2 kg− 1 VS HRAS process (kg N m− 3) energy balance was calculated making the following assumptions: (i) effective population equivalent, inflow and influent COD and ammonia nitrogen concentrations were considered the same for all configurations and they were assumed to the ones recorded from the current Rubí- Valldoreix WWTP, (ii) the energy consumption for physical processes: pumping, mixing, etc (Garrido et al., 2003) was the same for all con­ figurations and finally, (iii) the biogas production estimated through BMP of the sludge would be the biogas production achieved in the digester of the full-scale WWTP. mCODADS = (Qinfluent ⋅(tCODinfluent − tCODeffluent )) − mCODOX − mCODX 3. Results and discussion 3.1. COD removal, settleability and ammonium oxidation: comparison 1 kg VS 1.42 kg O2 mCODOX = Qinfluent ⋅(tCODinfluent − tCODeffluent )⋅(1 − Yobs ) 0.12 kg N 1 kg VS mCODINF = mCODX + mCODOX + mCODADS + mCODEFF between the performances of the HRAS pilot plant and the full-scale CAS The HRAS pilot plant was operated during ca. one year treating the ⋅ wastewater coming from the effluent of a primary settler of the full- Definition and calculation of the COD mass balance for HRAS pilot plant. scale WWTP, where a CAS process was implemented for COD 4 consumed ⋅ removal. This is a significant difference with previous studies where the inlet of the HRAS system directly came from the pre-treatment of an mCODINF = Qinfluent ⋅tCODinfluent mCODEFF = Qeffluent ⋅tCODeffluent ⌉ urban WWTP (Cao et al., 2020; Ge et al., 2017; Jimenez et al., 2015; mCODX = Qinfluent ⋅ N − NH+ Kinya et al., 2017; Liang et al., 2022; Taboada-Santos et al., 2020). The whole operational period can be divided in three periods depending on ⌈ the applied SRT (2.1, 1.0 and 0.6 days) as presented in Fig. 2 and Table 4. The average data of the full-scale WWTP operation for a year have been included in Table 4 for easing the comparison. The data were provided by the manager of the facility at the time of the study. Despite the HRT was maintained at a constant value of 3 h in the HRAS pilot plant, the applied organic loading rate (OLR) varied be­ tween the three operational periods since the average influent tCOD was different in each period. As can be seen, the OLRs were 2- or 3-fold COD in the influent (CODINF) COD in the effluent (CODEFF) COD as new biomass (CODX) greater than the applied in the full-scale CAS process. COD adsorbed (CODADS) The COD removal in the HRAS pilot plant was between 65 and 79% COD oxidized (CODOX) and it was correlated to the applied SRT since the lower the applied SRT, the lower the total COD removal. Liang et al. (2022) found similar COD balance COD removal operating at SRT between 0.5 and 2.0 days but those authors did not found a clear relationship between the applied SRT and Table 2 the achieved COD removal. However, this correlation was also found in other studies (Cao et al., 2020; Ge et al., 2017) but, in general, the 4 J. Carrera et al. Journal of Cleaner Production 354 (2022) 131734 Table 3 Table 3 (continued ) Definition and calculation of the energy balance for CAS-hetBNR and HRAS- Oxygen OCNO3 = oxygen autBNR configurations. consumption for consumption for Mass flow of VSSS = Qinfluent ⋅(tCODinfluent − VSSS = mass flow of complete complete oxidation secondary sludge tCODeffluent )⋅Yobs secondary sludge ammonia of ammonia (VSSS) (kg VS d− 1) nitrogen nitrogen to nitrate Qinfluent = oxidation to (kg O2 d− 1) volumetric flow rate nitrate (OCNO3) NO3 = in the influent (m3 concentration of d− 1) tCODinfluent = ammonia nitrogen COD concentration oxidized to nitrate in the influent (kg (kg N m− 3) O2 m− 3) Stoichiometric ratio tCODeffluent = COD for complete concentration in the ammonia oxidation effluent (kg O2 m− 3) to nitrate Yobs = observed considering biomass sludge yield (kg production: 4.3 kg kg− 1) O2 kg− 1 N Methane CH4PS = VSPS ⋅BMPPS CH4PS = methane Oxygen [TKN]influent OCNO2 = oxygen OCNO2 = Qinfluent ⋅ ⋅1.6 production from production from consumption for 2 consumption for primary sludge primary sludge (m3 partial ammonia partial oxidation of (CH4PS) CH4 d− 1) nitrogen ammonia nitrogen VSPS = mass flow of oxidation to to nitrite (kg O2 d− 1) primary sludge (kg nitrite (OCNO2) Stoichiometric ratio VS d− 1) for partial ammonia BMPPS (biomethane oxidation to nitrite potential of primary considering biomass sludge) = 0.37 m3 production: 1.6 kg CH4 kg− 1 VS ( O2 kg− 1 N Baquerizo et al., Assuming only half 2021) of the influent TKN Methane CH4SS = VSSS ⋅BMPSS CH4SS = methane is oxidized to nitrite production from production from to have a suitable secondary sludge secondary sludge influent for (CH4SS) (m3 CH4 d− 1) anammox reactor VSSS = mass flow of Energy EPHY = Qinfluent ⋅PEef ⋅20 EPHY = energy secondary sludge consumption for consumption for (kg VS d− 1) physical physical processes BMPSS = processes (EPHY) (kWh d− 1) biomethane PEef = 100000 potential of effective population secondary sludge equivalent (inh-eq) (m3 CH4 kg− 1 VS) Ratio of energy Total methane CH4T = CH4PS + CH4SS CH4T = total consumption = 20 production methane production Wh inh-eq− 1 ( (CH4T) (m3 CH4 d− 1) Garrido et al., 2003) Oxygen ( OCCOD = Qinfluent ⋅ tCODinfluent − OCCOD = oxygen Energy ECOD = (OCCOD − SOCCOD )⋅1.0 ECOD = energy consumption for ) consumption for consumption for consumption for tCODeffluent − 1.42⋅VSSS COD oxidation COD oxidation (kg COD oxidation COD oxidation (OCCOD) O2 d− 1) (kWh d− 1) Assuming an Specific energy empirical formula consumption for for the new biomass oxygen transfer = as: C5H7NO2, the 1.0 kWh kg− 1 O2 ( following ratio can Tchobanoglous be obtained: 1.42 kg et al., 2003) O2 kg− 1 VS Energy ENO3 = OCNO3 ⋅1.0 ENO3 = energy Saved oxygen SOCCOD = Qinfluent ⋅([TKN]influent − SOCCOD = saved consumption for consumption for consumption for [NO−3 ]effluent )⋅2.86 oxygen complete complete ammonia COD oxidation consumption for ammonia nitrogen oxidation through COD oxidation nitrogen to nitrate (kWh d− 1) heterotrophic through oxidation to denitrification heterotrophic nitrate (SOCCOD) denitrification (kg Energy ENO2 = OCNO2 ⋅1.0 ENO2 = energy O2 d− 1) consumption for consumption for [TKN]influent = TKN partial ammonia partial ammonia concentration in the nitrogen nitrogen oxidation influent (kg N m− 3) oxidation to to nitrite (kWh d− 1) [NO−3 ]effluent = nitrate nitrite Energy recovery by ECH4 = CH4T ⋅ρCH4 ⋅HCH4 ⋅η ECH4 = energy concentration in the combustion of recovery by effluent (kg N m− 3) biogas (ECH4) combustion of Stoichiometric ratio biogas (kWh d− 1) between COD ρCH4 = 0.657 oxidation with (methane density in oxygen or nitrate: kg CH4 m− 3 CH4) 2.86 kg O2 kg− 1 N HCH4 = 15.425 (heat OCNO3 = Qinfluent ⋅NO3 ⋅4.3 (continued on next page) 5 J. Carrera et al. Journal of Cleaner Production 354 (2022) 131734 Table 3 (continued ) Table 4 of methane Main results of the performance of the HRAS pilot plant throughout the three combustion in kWh different operational periods compared to the urban Rubí-WWTP (full-scale CAS kg− 1 CH4) ( process). The results of the full-scale CAS process are only from the secondary Tchobanoglous biological treatment. Most of the results are reported as an average value and its et al., 2003) standard deviation in an operational period. η = 0.35 (efficiency HRAS Full-scale of fueled electric pilot plant CAS process power generators) ( Garrido et al., 2003) Period I II III year 2015 SRT (d) 2.1 ± 1.0 ± 0.3 0.6 ± 14 ± 1 0.6 0.2 Operation time (d) 101 133 83 365 Temperature (◦ C) 15 ± 2 24 ± 3 20 ± 3 20 ± 5 Qinfluent (m3 d− 1) 2 2 2 20000 HRT (h) 3 3 3 7.6 DO (mg O2 L− 1) 2±1 2±1 2±1 2±1 tCODinfluent (mg O2 L− 1) 394 ± 284 ± 97 326 ± 350 ± 47 96 56 tCODeffluent (mg O2 L− 1) 73 ± 12 67 ± 16 113 ± 50 ± 24 37 COD removal (%) 79 ± 6 76 ± 6 65 ± 10 85 ± 8 OLR (kg COD m− 3 d− 1) 3.2 ± 2.3 ± 0.5 2.6 ± 1.1 ± 0.3 0.8 0.5 3 ORR (kg COD m− d− 1) 2.5 ± 1.7 ± 0.7 1.7 ± 0.9 ± 0.2 0.6 0.6 [N–NH4 ]influent - + − 8±3 − 19 ± 6 -6±4 − 11 ± 6 [N–NH4+]effluent (mg N L− 1) SVI (mL g− 1) 214 ± 293 ± 96 160 ± 150 ± 79 74 89 [TSS]eff (mg L− 1) 28 ± 16 12 ± 5 34 ± 15 25 ± 13 Secondary settler efficiency 99 98 93 99 (%) [VSS]reactor (mg L− 1) 2300 ± 800 ± 200 500 ± 3000 ± 700 800 100 Fig. 2. Performance of the HRAS pilot plant: (i) Temperature in the HRAS [VSS]purge (mg L− 1) 4100 ± 1800 ± 1500 ± 5600 ± 1300 reactor, (ii) influent and effluent total COD concentrations and (iii) influent and 1500 600 600 effluent N-compounds concentrations. reported COD removal efficiencies were lower than the ones achieved in this study. However, COD removal did not show a clear correlation with the temperature of the HRAS pilot plant (Table 4). COD removal can be increased in a HRAS process up to about 90% if iron salts are added to remove phosphate since its addition improves the settleability of the sludge and, consequently the COD removal (Taboada-Santos et al., 2020). In this study, iron salts were not added for a better assessment and comparison with the full-scale CAS process, where iron salts were not added either. Moreover, if COD removal in the HRAS pilot plant is compared with the COD removal achieved in the full-scale CAS process, the correlation between COD removal and applied SRT is clear (Fig. 3). It is interesting to note that the COD removal achieved at SRT of 2.1 days in the HRAS pilot plant was similar to the maximum one achieved in the full-scale CAS process despite the applied OLR was 3-fold higher in the HRAS pilot plant compared to the full-scale CAS process. The reason behind this result could be that the operational conditions in the HRAS pilot plant (lower HRT and SRT than in the full-scale CAS process) promoted the development of an r-strategist culture in contrast to the biomass of the CAS process that could be a k-strategist culture. The se­ Fig. 3. Experimental COD removal in the HRAS pilot plant and the full-scale lection of an r- or k-strategist culture can be achieved by modifying the CAS process as a function of the applied SRT. A correlation between COD operational conditions (Yu et al., 2020), as in this case. An r-strategist removal and applied SRT has been found by curve-fitting of the experi­ culture has a higher substrate consumption rate but a lower substrate mental data. affinity than a k-strategist culture (Martín-Hernández et al., 2009) and this fact would explain the higher organic removal rate (ORR) in the worst SVI values were measured at the highest temperatures (>25 ◦ C) in HRAS pilot plant compared to the full-scale CAS process (Table 4). Period II. However, SVI did not show any clear correlation with SRT Sludge settleability is directly related to the performance of the (Table 4). EFSS values were high (98–99%) for SRT of 1.0 and 2.1 days HRAS process since it largely determines the amount of organic matter but it decreased up to 93% for SRT of 0.6 days. Therefore, the operation in the effluent. Settleability of the HRAS pilot plant sludge was estimated at SRT lower than 1 day clearly affected the EFSS, causing a worsening of using two parameters: (i) SVI and (ii) EFSS. SVI values of the sludge of the the effluent quality by the presence of solids. HRAS pilot plant were between 160 and 294 mL g− 1, indicating an If the HRAS process is designed to be connected to an autotrophic acceptable settleability of the activated sludge throughout the study and BNR process, a key point is to prevent nitrification from occurring in the in all the operational conditions despite no iron salts were added. The 6 J. Carrera et al. Journal of Cleaner Production 354 (2022) 131734 HRAS process. At low temperatures (<20 ◦ C), nitrification was pre­ 40–46% at SRT 0.3 days and 44–55% at SRT 0.6 days in Kinya et al. vented at all applied SRT values since no nitrate formation was recorded (2017)) than those found in this study (51% at SRT 0.6 days and 54% at and the ammonium removal matched with ammonium consumption for SRT 1.0 and 2.1 days). These differences on the COD recovery could be heterotrophic growth. At high temperatures (>20 ◦ C), nitrification took due to several factors. First, the influent of these HRAS pilot plants was place at SRT of 1.0 day and a significant concentration of nitrate was different from that of the pilot plant of this study. The influent of measured in the HRAS pilot plant effluent (Fig. 2). However, the nitri­ Jimenez et al. (2015) and Kinya et al. (2017) was the effluent of a fication was completely stopped when SRT was lowered to 0.6 days in pre-treatment without primary sedimentation whereas the influent of spite of working at temperatures higher than 20 ◦ C. this study was the effluent of a primary settler. Therefore, the influent COD fractionation was significantly different, having our influent a 3.2. Fate of COD in the HRAS pilot plant compared to the full-scale CAS higher percentage of sCOD (33% vs 10–23%) while the influent of the process other studies had a higher percentage of pCOD (60–66% vs 47%). Sec­ ond, HRT in our pilot plant was 3 h and the HRTs in the other studies The fate of the organic matter in the HRAS pilot plant and the full- were 0.5–1.0 h. A HRAS treating the effluent of a pre-treatment probably scale CAS process was evaluated through the COD-mass balance pre­ has better COD removal percentages at low HRTs than a HRAS treating sented in Table 2 and calculated for each operational period. The the effluent of a primary settler because what it removes is, basically, comparison of the results of a HRAS pilot plant with a full-scale CAS colloidal organic matter by adsorption and particulate organic matter by system treating the same influent is of interest since this comparison is sedimentation (in the secondary settler, not in the HRAS). A HRAS not usually carried out in this kind of studies (Cao et al., 2020; Ge et al., treating the effluent of a primary settler has to operate at higher HRT 2017; Jimenez et al., 2015; Kinya et al., 2017; Liang et al., 2022; because the percentage of particulate and colloidal fractions that re­ Taboada-Santos et al., 2020). ceives is lower, since these fractions are mainly recovered in the primary The observed sludge yield coefficient (Yobs) was calculated for each settler. However, the combination of primary settler plus HRAS is able to operational period as explained in section 2.2.1. Yobs was found to be achieve a global higher COD recovery than a single HRAS system. correlated with the applied SRT (Fig. 4A). Yobs increased by 52, 61 and Hence, this comparison suggests that the organic matter recovery in a 77% compared to the value of a SRT of 14 days (full-scale CAS system) HRAS process would increase if the urban wastewater is previously for SRTs of 2.1, 1.0 and 0.6, respectively (Fig. 4A). treated in a primary settler and if the HRT is in the upper part of the The results of the COD mass balance are shown in Fig. 4. In the full- range applied in HRAS processes. scale CAS process, only 38% of the COD was recovered as the sum of The applied SRT had a limited effect on the total recovered COD in organic matter adsorbed over the activated sludge (CODADS) and new the HRAS pilot plant, as all of them were between 51 and 54%, irre­ biomass formed (CODX). The lost organic matter (62%) was the sum of spective the SRT applied. Nevertheless, there was a difference in the the oxidized organic matter (CODOX) and the organic matter lost with distribution of CODADS and CODX for the SRTs of 0.6 and 2.1 compared the effluent (CODEFF). The percentage of recovered COD increased up to to the SRT of 1.0 day. At SRT 0.6 and 2.1 days, the most important 53%, on average, in the HRAS pilot plant. That is, the HRAS pilot plant fraction of the recovered COD was the CODADS fraction, whereas at SRT recovers, on average, 40% more organic matter than the full-scale CAS of 1.0 day, the most important fraction was the CODX fraction. Probably, process. Regarding the composition of the recovered COD in the full- this difference was related to the average temperature of each opera­ scale CAS process, only 37% was CODADS and 63% was CODX whereas tional period rather than the applied SRT. At SRT 0.6 and 2.1 days, the in the HRAS pilot plant, the CODADS represented, on average, 53% and average temperatures were 20 ± 3 ◦ C and 15 ± 2 ◦ C, respectively but at CODX was 47%. Regarding the composition of the removed COD in the SRT 1.0 day, the average temperature was 24 ± 3 ◦ C. This means that, at full-scale CAS process, 77% was CODOX and only 23% was CODEFF. high temperatures (>20 ◦ C), the produced biomass increased at the cost These percentages were, on average, 45% of CODOX and 55% of CODEFF of a reduction in organic matter adsorbed in the sludge. Jia et al. (2020) for the HRAS pilot plant. Therefore, the COD recovery was higher in the predicted, in a model-based study, the decrease of the COD recovered in HRAS process than in the full-scale CAS process because of the increase a HRAS process when temperature increase over 20 ◦ C but they do not of the COD adsorption in the sludge and the significant reduction of the split the recovered COD in the CODADS and CODX fractions as we did. COD oxidation. The results of our study seem to contradict their model-based pre­ Our results are different than those reported in HRAS pilot plants dictions since the total COD recovered did not significantly change with operated at similar SRTs and temperatures (20-25 ◦ C). Jimenez et al. the increase of the temperature. Nevertheless, it seems the temperature (2015) and Kinya et al. (2017) reported lower COD recoveries (45% at was only affecting the distribution of CODADS and CODX fractions and SRT 0.5 days and 25% at SRT 1 and 2 days in Jimenez et al. (2015) and not the total recovered COD. Fig. 4. (A) Observed sludge yield coefficient (Yobs) in the HRAS pilot plant and full-scale CAS process as a function of the applied SRT. (B) Biomethane potential at standard temperature and pressure (BMP at STP) as a function of the SRT imposed in the HRAS pilot plant compared to that in full-scale CAS process. 7 J. Carrera et al. Journal of Cleaner Production 354 (2022) 131734 However, the SRT effect was much higher on the distribution of the producing an effluent with a high COD concentration, since this COD lost COD because CODOX significantly increased (up to 93% from 0.6 to would negatively affect the subsequent autotrophic BNR (Li et al., 2020; 2.1 days) with the increase of the SRT. On the contrary, CODEFF signif­ Pijuan et al., 2020). icantly decreased (up to 84% from 0.6 to 2.1 days) with the increase of Overall the data gathered from the experimental campaign indicate the SRT due to the improvement of the secondary settler efficiency, from that the SRT applied should be regulated as a function of temperature. 93% (SRT = 0.6 d) to 99% (SRT = 2.1 d). The model-based study of Jia For low temperatures, nitrification can be easily circumvented, and good et al. (2020) also predicted the high influence of secondary settler effi­ effluent quality (even without iron salts addition) can be obtained at ciency on CODEFF in a HRAS process, with similar results to these SRT of 2 days. Unexpectedly, high temperatures complicated the oper­ experimentally obtained in this study. ation of the HRAS reactor since to avoiding nitrification required of a Consequently, working at SRTs between 0.6 and 2.1 days in the short SRT (lower than 1.0 day), and at such short SRT the sludge set­ HRAS process would have not a noteworthy effect on the recovery of tleability is compromising the effluent quality. organic matter and it has only a noticeable effect on the distribution of the lost organic matter. In any case, considering that HRAS process would be connected to a subsequent autotrophic BNR process, it seems 3.3. Biomethane potential results to be a better option to lose organic matter by oxidation rather than Biomethane potentials of the secondary sludge of the HRAS pilot Fig. 5. COD-mass balance in the HRAS pilot plant and the full-scale CAS process. 8 J. Carrera et al. Journal of Cleaner Production 354 (2022) 131734 plant and the full-scale CAS process were measured using the BMP batch aeration for COD removal and for nitrogen (N) removal. The tests described in section 2.1.4 (Fig. 4B). BMP of the secondary sludge of energy consumption for COD removal is 2- to 4-fold higher for the CAS process (165 mL CH4 g− 1 VS at SRT 14 days) was close to the HRAS-autBNR configurations than for CAS-hetBNR configuration values reported for secondary sludge of other full-scale CAS systems due to the savings in oxygen consumption through heterotrophic operated at similar SRT (150 mL CH4 g− 1 VS at SRT 15 days; Bolzonella denitrification. However, the energy consumption for N-removal et al., 2005). However, the BMPs of the sludge from the HRAS system is 5-fold lower for HRAS-autBNR configurations than that of CAS- (267–317 mL CH4 g− 1 VS) were significantly higher than the values hetBNR configuration due to the greater savings in oxygen con­ reported for CAS systems, regardless of the SRT applied. The maximum sumption for oxidation of half of ammonium to nitrite instead of BMP (317 mL CH4 g− 1 VS) was obtained at SRT 0.6 days. In fact, the oxidation of all of the ammonium to nitrate. BMP of the secondary sludge produced in the HRAS system is much (ii) Energy consumed in the physical processes. This energy con­ closer to the BMP of primary sludge (250–370 mL CH4 g− 1 VS; Elbesh­ sumption was the same for both configurations but makes up to bishy et al., 2012; Baquerizo et al., 2021) than that of secondary sludge 26% of total energy consumption for CAS-hetBNR configuration from CAS systems and even higher than the BMP of the organic fraction and 42–50% for HRAS-autBNR configurations. of municipal waste (212 mL CH4 g− 1 VS; Elbeshbishy et al., 2012). (iii) Energy recovered from biogas produced by anaerobic digestion of According to van Loosdrecht et al. (1997), the lower the applied SRT sludge. The energy recovery is 29–36% higher for the HRAS- in an activated sludge system, the higher the BMP of the produced autBNR configurations than for CAS-hetBNR configuration. sludge. However, the BMP achieved at SRT 1.0 day slightly disturb the expected trend since it was similar to the BMP achieved at SRT 2.1 days All the above results are based on the calculations presented in (Fig. 4B). This is probably due to the composition of the recovered COD Table 3. One of the main items of that energy balance is the estimation of at each SRT (Fig. 5). As explained before, the CODX fraction was the the total methane production from primary and secondary sludge. That predominant one at SRT 1.0 day while the CODADS was the dominant estimation is based on the calculation of the sludge produced and the fraction at SRT 0.6 and 2.1 days. It is well known that the CODADS BMP of each sludge type. The data of real methane production in the fraction has higher BMP than the CODX fraction (Baquerizo et al., 2021) full-scale CAS system were used to check the consistency of that esti­ and this would explain the fact that BMP achieved at SRT 1.0 day is mation. The measured total methane production on the full-scale CAS almost the same than the achieved at SRT 2.1 days. system was 1360 m3 d− 1 while the estimation through our procedure was 1354 m3 d− 1, less than 1% of the difference (see Table 5). Never­ theless, some values of the energy balance calculations could slightly 3.4. Energy balance: comparison between the CAS and the HRAS vary, depending on if they were directly obtained from the operation of processes pilot- or full-scale devices (for example, an anaerobic digestor or an aeration system for N-removal). The main results obtained from the energy balances calculated for The CAS-hetBNR configuration is a net energy consumer process two different WWTP configurations (as described in section 2.2.3) are while all the HRAS-autBNR configurations are net energy producer presented in Table 5 and can be summarized as follows: processes (a direct comparison of the total energy consumed and recovered for each configuration is depicted in Fig. 6). CAS-hetBNR (i) Energy consumed in the aeration of the biological processes. The configuration only recovers 63% of its energy consumption, meaning a total amount of energy consumption is constituted by the Table 5 Main results of the energy balance for CAS-hetBNR and HRAS-autBNR configurations. Parameter Acronym Units CAS- HRAS-autBNR hetBNR SRT = 14 d SRT = 2.1 d SRT = 1.0 d SRT = 0.6 d Influent flow Qinfluent m3 d− 1 20000 (a) 20000 (b) 20000 (b) 20000 (b) Influent COD in the secondary treatment tCODinfliuent kg COD m− 3 0.35 (a) 0.35 (b) 0.35 (b) 0.35 (b) Removal efficiency %RE % 86 (a) 79 (a) 76 (a) 65 (a) Effluent COD secondary treatment tCODeffluent kg COD m− 3 0.05 (a) 0.074 (c) 0.084 (c) 0.122 (c) Observed growth yield Yobs kg VSS kg− 1 0.31 (a) 0.47 (a) 0.50 (a) 0.55 (a) COD VS in the secondary sludge VSSS kg VS d− 1 1880 (c) 2594 (c) 2660 (c) 2508 (c) VS in the primary sludge VSPS kg VS d− 1 2820 (c) 2820 (b) 2820 (b) 2820 (b) Proportion of secondary sludge %SS % 40 (a) 48 (c) 49 (c) 47 (c) VS influent anaerobic digester VST kg VS d− 1 4700 (a) 5414 (c) 5480 (c) 5328 (c) Biochemical Methane Potential for secondary sludge BMPSS m3 CH4 kg− 1 VS 0.165 (a) 0.270 (a) 0.268 (a) 0.317 (a) Methane production from primary sludge CH4PS m3 CH4 d− 1 1044 (c) 1044 (b) 1044 (b) 1044 (b) Methane production from secondary sludge CH4SS m3 CH4 d− 1 310 (c) 702 (c) 710 (c) 793 (c) Total methane production CH4T m3 CH4 d− 1 1354 (c) 1745 (c) 1754 (c) 1837 (c) 1360 (a) 1 Oxygen consumption for COD oxidation OCCOD kg O2 d− 3330 (c) 1837 (c) 1543 (c) 999 (c) 1 Saved oxygen consumption for COD oxidation through heterotrophic SOCCOD kg O2 d− 2860 (c) - - - denitrification Oxygen consumption for complete ammonia nitrogen oxidation to nitrate OCNO3 kg O2 d− 1 5160 (c) - - - Oxygen consumption for partial ammonia nitrogen oxidation to nitrite OCNO2 kg O2 d− 1 - 960 (c) 960 (c) 960 (c) Energy consumption for physical processes EPHY kWh d− 1 2000 (c) 2000 (c) 2000 (c) 2000 (c) Energy consumption for COD oxidation ECOD kWh d− 1 470 (c) 1834 (c) 1543 (c) 999 (c) Energy consumption for complete ammonia nitrogen oxidation to nitrate ENO3 kWh d− 1 5160 (c) - - - Energy consumption for partial ammonia nitrogen oxidation to nitrite ENO2 kWh d− 1 - 960 (c) 960 (c) 960 (c) Energy recovery by combustion of biogas ECH4 kWh d− 1 4801 (c) 6185 (c) 6229 (c) 6521 (c) a Experimental value. b Assumed value. c Calculated value. 9 J. Carrera et al. Journal of Cleaner Production 354 (2022) 131734 the range 1.0–2.1 days), because at these SRTs, the secondary settler efficiency and the COD recovery are higher than the those achieved at SRT of 0.6 days (Table 4 and Fig. 5). Additionally, despite the observed sludge yield coefficient and biomethane potential of the sludge are lower than those achieved at SRT of 0.6 days, the energy recovery in form of biogas would be close to the achieved at SRT of 0.6 days (Fig. 6). However, part of the inlet ammonium can be nitrified in the HRAS system at temperatures above 20 ◦ C when it is operated at SRTs from 1.0 to 2.1 days. 4. Conclusions • The removal of COD in the HRAS pilot plant depended on the SRT imposed, ranging from 65% at SRT 0.6 days to 79% at SRT 2.1 days, with temperatures between 12 and 28 ◦ C. Moreover, these COD re­ movals were achieved without iron salts addition for improving settleability in the HRAS system. The maximum COD removal in the HRAS pilot plant was close to the COD removal achieved in the full- scale CAS system used as reference (85% at SRT 14 days). • The observed sludge yield coefficients achieved in the HRAS pilot plant ranged from 0.47 kg VSS kg− 1 COD at SRT 2.1 days to 0.55 kg VSS kg− 1 COD at SRT 0.6 days, being in all cases higher than the achieved in the full-scale CAS system (0.31 kg VSS kg− 1 COD at SRT Fig. 6. Comparison of the total energy consumed and recovered for CAS- 14 days). hetBNR and HRAS-autBNR configurations. The relative errors of each energy • The recovery of COD in the HRAS pilot plant ranged from 51% at SRT component are 5% for energy consumption components and 10% for energy recovery component. 0.6 days to 54% at SRT 2.1 days, being in all cases higher than the achieved in the full-scale CAS system (38% at SRT 14 days). • The performance of the HRAS process was found to be directly linked net energy consumption of 0.14 kWh m− 3. On the contrary, HRAS- to the sludge settleability, which in turn depended on the SRT autBNR configurations recover from 129% up to 165% of their energy imposed. consumptions, being the HRAS-autBNR configuration with a HRAS • Avoiding nitrification in the HRAS system at temperatures above working at SRT = 0.6 days the most favourable process since it has the 20 ◦ C is required. Imposing low SRT at high temperatures easily lowest energy consumption and the highest energy recovery. Depending prevented nitrification (e.g., an SRT of 0.6 days completely sup­ on the SRT of the HRAS process, the net energy recovery of these con­ pressed nitrification at 25 ◦ C). figurations is: 0.07 kWh m− 3 (SRT = 2.1 days), 0.09 kWh m− 3 (SRT = • The biomethane potential of the sludge of the HRAS pilot plant 1.0 day) and 0.13 kWh m− 3 (SRT = 0.6 days). ranged from 270 mL g− 1 VS at SRT 2.1 days to 317 mL g− 1 VS at SRT The average increase in the energy recovery and the decrease of the 0.6 days, being in all cases higher than the achieved with the sec­ energy consumption when passing from CAS-hetBNR to HRAS-autBNR ondary sludge of the full-scale CAS system (165 mL g− 1 VS at SRT 14 configuration are 1510 kWh d− 1 and 3211 kWh d− 1, respectively. This days). means that switching from a net energy consumer to a net energy pro­ • A detailed energy balance indicated that two-thirds reduction of ducer process resulted in two-thirds reduction of aeration requirements aeration requirements and one-third increase of biogas production and one-third increase of biogas production. could be achieved in a WWTP configuration in which an HRAS sys­ tem is coupled to an autotrophic BNR (HRAS-autBNR) process 3.5. Practical implications compared to the CAS-hetBNR configuration. This would yield on average a net energy production of ca. 0.1 kWh per cubic meter of Because the results obtained include a rather wide set of SRT applied wastewater treated in a HRAS-autBNR configuration. (SRTs from 0.6 to 2.1 days) with apparently similar treatment perfor­ mance (recovering 51–54% of the inlet organic matter as energy, Fig. 5), CRediT authorship contribution statement a clear guideline for the recommendations derived from our study would be necessary. In fact the selection of the SRT is a complex decision and it Julián Carrera: Writing – original draft, Supervision. Oriol Carbó: requires for a more detailed definition of the goal of the treatment. Experiments. Silvia Doñate: Idea development. María Eugenia Suárez- If the objective is to achieve the maximum energy recovery, the SRT Ojeda: Idea development, Writing. Julio Pérez: Writing – original draft, should be around 0.6 days (Fig. 6), because at this SRT, the highest Supervision. observed sludge yield coefficient and biomethane potential of the sludge were achieved (Fig. 4). At this SRT, the nitrification of the inlet Declaration of competing interest ammonium is also prevented even at temperatures around 25 ◦ C (Fig. 2). This is a determining point for maximization of energy recovery if the The authors declare that they have not financial and personal re­ desired configuration is a HRAS system followed by an autotrophic BNR lationships that could inappropriately influence (bias) their work. process, in the case that the feasibility of autotrophic BNR at mainstream full-scale conditions would be demonstrated. The weak point of the Acknowledgements HRAS performance at STR of 0.6 days is the high loss of organic matter in the effluent (Fig. 4) due to the limited efficiency of the solids sepa­ This research was funded by the European Commission through the ration in the secondary settler at this SRT (Table 4). This fact could affect LIFE Program: “LIFE SAVING-E: Two-stage autotrophic N-removal for the subsequent autotrophic BNR process, where the presence of organic mainstream sewage treatment” (LIFE14/ENV/ES/000633). The authors matter could compromise both the partial nitritation and the anammox want to thank the Catalan Public Administration (Agència Catalana de processes. In this case, the HRAS should be operated at higher SRT (in l’Aigua) for its help in the development of this research. 10 J. Carrera et al. Journal of Cleaner Production 354 (2022) 131734 References Khalkhali, M., Mo, W., 2020. The energy implication of climate change on urban wastewater systems. J. Clean. Prod. 267, 121905 https://doi.org/10.1016/j. jclepro.2020.121905. Angelidaki, I., Alves, M., Bolzonella, D., Borzacconi, L., Campos, J.L., Guwy, A.J., Kinya, M.N., Elliott, M., Wett, B., Murthy, S., Chandran, K., Bott, C.B., 2017. The role of Kalyuzhnyi, S., Jenicek, P., van Lier, J.B., 2009. Defining the biomethane potential extracellular polymeric substances on carbon capture in a high rate activated sludge (BMP) of solid organic wastes and energy crops: a proposed protocol for batch A-stage system. Chem. Eng. J. 322, 428–434. https://doi.org/10.1016/j. assays. Water Sci. Technol. 59 (5), 927–934. https://doi.org/10.2166/wst.2009.040. cej.2017.04.043. APHA, 1999. Standard Methods for the Examination of Water and Wastewater, twentieth Li, J., Li, J., Peng, Y., Wang, S., Zhang, L., Yang, S., Li, S., 2020. Insight into the impacts ed. American Publishers Health Association, Washington DC, USA. of organics on anammox and their potential linking to system performance of sewage Baquerizo, G., Fiat, J., Buffiere, P., Girault, R., Gillot, S., 2021. Modelling the dynamic partial nitrification-anammox (PN/A): a critical review. Bioresour. Technol. 300, long-term performance of a full-scale digester treating sludge from an urban WRRF 122655 https://doi.org/10.1016/j.biortech.2019.122655. using an extended version of ADM1. Chem. Eng. J. 423, 128870 https://doi.org/ Liang, M., Lu, X., Liu, P., Wu, X., Zan, F., 2022. Tapping the energy potential from 10.1016/j.cej.2021.128870. wastewater by integrating high-rate activated sludge process with anaerobic Bolzonella, D., Pavan, P., Battistoni, P., Cecchi, F., 2005. Mesophilic anaerobic digestion membrane bioreactor. J. Clean. Prod. 333, 130071 https://doi.org/10.1016/j. of waste activated sludge: influence of the solid retention time in the wastewater jclepro.2021.130071. treatment process. Process Biochem. 40, 1453–1460. In: https://doi:10.1016/j. Liu, R., Huang, R., Shen, Z., Wang, H., Xu, J., 2021. Optimizing the recovery pathway of procbio.2004.06.036. a net-zero energy wastewater treatment model by balancing energy recovery and Cao, S., Lu, D., Phuac, K., Yana, W., Le, C., Tao, G., Zhou, Y., 2020. Organics eco-efficiency. Appl. Energy 298, 117157. https://doi.org/10.1016/j. transformation and energy production potential in a high rate A-stage system: a apenergy.2021.117157. demo-scale study. Bioresour. Technol. 295, 122300 https://doi.org/10.1016/j. Macintosh, C., Astals, S., Sembera, C., Ertl, A., Drewes, J., Jensen, P., Koch, K., 2019. biortech.2019.122300. Successful strategies for increasing energy self-sufficiency at Grüneck wastewater Capodaglio, A.G., Olsson, G., 2020. Energy issues in sustainable urban wastewater treatment plant in Germany by food waste co-digestion and improved aeration. Appl. management: use, Demand reduction and recovery in the urban water cycle. Energy 242, 797–808. https://doi.org/10.1016/j.apenergy.2019.03.126. Sustainability 12, 266. https://doi:10.3390/su12010266. Mamais, D., Jenkins, D., Pitt, P., 1993. A rapid physical-chemical method for the Cardoso, B., Rodrigues, E., Gaspar, A.R., Gomes, A., 2021. Energy performance factors in determination of readily biodegradable soluble COD in municipal wastewater. Water wastewater treatment plants: a review. J. Clean. Prod. 322, 129107 https://doi.org/ Res. 27, 195–197. https://doi.org/10.1016/0043-1354(93)90211-Y. 10.1016/j.jclepro.2021.129107. Martín-Hernández, M., Carrera, J., Pérez, J., Suárez-Ojeda, M.E., 2009. Enrichment of a de Graaff, M.S., van den Brand, T.P.H., Roest, K., Zandvoort, M.H., Duin, O., van K-strategist microbial population able to biodegrade p-nitrophenol in a sequencing Loosdrecht, M.C.M., 2016. Full-scale highly-loaded wastewater treatment processes batch reactor. Water Res. 43, 3871–3883. https://doi:10.1016/j.watres.2009.06.00 (A-stage) to increase energy production from wastewater: performance and design 1. guidelines. Environ. Eng. Sci. 33, 571–577. https://doi.org/10.1089/ees.2016.0022. Pérez, J., Isanta, E., Carrera, J., 2015. Would a two-stage N-removal be a suitable de Vrieze, J., de Lathouwer, L., Verstraete, W., Boon, N., 2013. High-rate iron-rich technology to implement at full scale the use of anammox for sewage treatment? activated sludge as stabilizing agent for the anaerobic digestion of kitchen waste. Water Sci. Technol. 72 (6), 858–864. https://doi: 10.2166/wst.2015.281. Water Res. 47, 3732–3741. https://doi.org/10.1016/j.watres.2013.04.020. Pijuan, M., Ribera-Guardia, A., Balcázar, J.L., Micó, M., de la Torre, T., 2020. Effect of Elbeshbishy, E., Nakhla, G., Hafez, H., 2012. Biochemical methane potential (BMP) of COD on mainstream anammox: evaluation of process performance, granule food waste and primary sludge: influence of inoculum pre-incubation and inoculum morphology and nitrous oxide production. Sci. Total Environ. 712, 136372 https:// source. Bioresour. Technol. 110, 18–25. http://doi:10.1016/j.biortech.2012.01.025. doi.org/10.1016/j.scitotenv.2019.136372. European Environmental Agency, 2022. Greenhouse Gas Emission Intensity of Electricity Rahman, A., Hasan, M., Meerburg, F., Jimenez, J.A., Millere, M.W., Bott, C.B., Al- Generation in Europe. https://www.eea.europa.eu/ims/greenhouse-gas-emission-in Omari, A., Murthy, S., Shaw, A., De Clippeleir, H., Riffat, R., 2020. Moving forward tensity-of-1. (Accessed 24 March 2022). with A-stage and high-rate contact-stabilization for energy efficient water resource Eurostat, 2021a. Population and Population Change Statistics. https://ec.europa.eu/eur recovery facility: mechanisms, factors, practical approach, and guidelines. J. Water ostat/data/database. (Accessed 24 March 2022). Proc. Eng. 36, 101329 https://doi.org/10.1016/j.jwpe.2020.101329. Eurostat, 2021b. Electricity Prices for Industrial Consumers - Bi-annual Data (Until Reino, C., Suárez-Ojeda, M.E., Pérez, J., Carrera, J., 2016. Kinetic and microbiological 2007). https://ec.europa.eu/eurostat/data/database. (Accessed 29 April 2021). characterization of aerobic granules performing partial nitritation of a low-strength Garrido, J.M., Fdz-Polanco, M., Fdz-Polanco, F., 2003. Working with energy and mass wastewater at 10 ◦ C. Water Res. 101, 147–156. https://doi.org/10.1016/j. balances: a conceptual framework to understand the limits of municipal wastewater watres.2016.05.059. treatment. Water Sci. Technol. 67 (10), 2294–2301. https://doi.org/10.2166/ Reino, C., Suárez-Ojeda, M.E., Pérez, J., Carrera, J., 2018. Stable long-term operation of wst.2013.124. an upflow anammox sludge bed reactor at mainstream conditions. Water Res. 128, Ge, H., Batstone, D., Mouiche, M., Hu, S., Keller, J., 2017. Nutrient removal and energy 331–340. https://doi.org/10.1016/j.watres.2017.10.058. recovery from high-rate activated sludge processes – impact of sludge age. Bioresour. Taboada-Santos, A., Rivadulla, E., Paredes, L., Carballa, M., Romalde, J., Lema, J.M., Technol. 245, 1155–1161. https://doi.org/10.1016/j.biortech.2017.08.115. 2020. Comprehensive comparison of chemically enhanced primary treatment and Jetten, M.S.M., Horn, S.J., van Loosdrecht, M.C.M., 1997. Towards a more sustainable high-rate activated sludge in novel wastewater treatment plant configurations. municipal wastewater treatment system. Water Sci. Technol. 35, 171–180. https:// Water Res. 169, 115258 https://doi.org/10.1016/j.watres.2019.115258. doi.org/10.1016/S0273-1223(97)00195-9. Tchobanoglous, G., Burton, F.L., Stensel, H.D., 2003. Wastewater Engineering: Jia, M., Solon, K., Vandeplassche, D., Venugopal, H., Volcke, E.I.P., 2020. Model-based Treatment and Reuse, fourth ed. Metcalf & Eddy, Inc, New York, USA. McGraw-Hill. evaluation of an integrated high-rate activated sludge and mainstream anammox Torretta, V., Tolkou, A.K., Katsoyiannis, I.A., Caccamo, F.M., Miino, M.C., Baldi, M., system. Chem. Eng. J. 382, 122858 https://doi.org/10.1016/j.cej.2019.122878. Collivignarelli, M.C., 2021. Enhancement of methanogenic activity in volumetrically Jimenez, J., Miller, M., Bott, C., Murthy, S., de Clippeleir, H., Wett, B., 2015. High-rate undersized reactor by mesophilic Co-digestion of sewage sludge and aqueous activated sludge system for carbon management –Evaluation of crucial pro- cess residue. Sustainability 13, 7728. https://doi.org/10.3390/su13147728. mechanisms and design parameters. Water Res. 87, 476–482. https://doi.org/ van Loosdrecht, M.C.M., Kuba, T., van Veldhuizen, H.M., Brandse, F.A., Heijnen, J.J., 10.1016/j.watres.2015.07.032. 1997. Environmental impacts of nutrient removal processes: a case study. J. Environ. Juan-Díaz, X., Pérez, J., Carrera, J., 2021. Effective dampening of temperature effects in Eng. 123, 33–40. https://doi.org/10.1061/(ASCE)0733-9372(1997)123:1(33). an anammox reactor treating real mainstream wastewater. J. Water Proc. Eng. 40, Yu, L., Chen, S., Chen, W., Wu, J., 2020. Experimental investigation and mathematical 101853 https://doi.org/10.1016/j.jwpe.2020.101853. modeling of the competition among the fast-growing ‘‘r-strategists” and the slow- Juan-Díaz, X., Olmo, L., Pérez, J., Carrera, J., 2022. Coupling anammox and growing ‘‘K-strategists” ammonium-oxidizing bacteria and nitrite-oxidizing bacteria heterotrophic denitrification activity at mainstream conditions in a single reactor in nitrification. Sci. Total Environ. 702, 135049 https://doi.org/10.1016/j. unit. Chem. Eng. J. 431, 134087 https://doi.org/10.1016/j.cej.2021.134087. scitotenv.2019.135049. Kartal, B., Kuenen, J.B., van Loosdrecht, M.C.M., 2010. Sewage treatment with anammox. Science 328, 702–703. https://doi: 10.1126/science.1185941. 11