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Chemical Engineering Journal 172 (2011) 654– 664

Contents lists available at ScienceDirect

Chemical Engineering Journal

jo u r n al hom epage: www.elsev ier .com/ locate /ce j

reatment of biorecalcitrant �-methylphenylglycine aqueous solutions with aolar photo-Fenton-aerobic biological coupling: Biodegradability andnvironmental impact assessment

nna Serraa, Enric Brillasb, Xavier Domènecha, José Perala,∗

Departament de Química, Edifici Cn, Universitat Autònoma de Barcelona, E-08193 Bellaterra, Barcelona, SpainLaboratori d’Electroquímica dels Materials i del Medi Ambient, Departament de Química Física, Facultat de Química, Universitat de Barcelona, Martí i Franquès 1-11,8028 Barcelona, Spain

r t i c l e i n f o

rticle history:eceived 8 February 2011eceived in revised form 10 June 2011ccepted 11 June 2011

eywords:-Methyl-phenylglycinedvanced Oxidation Processes (AOP)hemical–biological couplingolar photo-Fenton

a b s t r a c t

The photo-Fenton process under solar irradiation (SPF) was coupled to an aerobic activated sludge bio-logical treatment carried out in a sequencing batch reactor (SBR). The effectiveness of the coupling forthe treatment of aqueous solutions polluted with non-biodegradable �-methyl-phenylglycine (�-MPG)was evaluated. Respirometry was chosen as alternative to other conventional tests (BOD5/COD, Zahn-Wellens, Microtox®) for the quantification of both biodegradability and toxic/inhibitory effect of thepolluted water along the SPF time, the final goal being the determination of the suitable reaction time forchemical–biological coupling. The nitrification of the aqueous solution during the biological treatment, aprocess that might take place at the same time than the oxidation of organic matter and endogenous res-piration, was evaluated along the different SBR cycles. Finally, a comparative life cycle assessment (LCA)

espirometryife cycle assessment (LCA)

of SPF and an integrated chemical–biological coupling (SPF–SBR) was carried out in order to identifythe best alternative, from an environmental point of view, for the treatment of biorecalcitrant pollu-tants. The determination of the main impact factors of the biological treatment was also carried out. Theresults of the bench-scale study indicate that both the SPF and SPF–SBR treatments are environmentallyfriendly and effective alternative strategies for the treatment of water loaded with non-biodegradable

compounds.

. Introduction

Nowadays, the presence of emerging or unregulated pollutants,uch as pharmaceutical and personal care products, disinfectiony-products, gasoline additives, pesticides and their degradationy-products, etc., in water is a cause for environmental concern [1].requently, many of such pollutants are of bio-recalcitrant nature,nd they cannot be degraded by the activated sludge of a conven-ional municipal wastewater treatment plant (WWTP), ending upn the environment and leading to probable toxic effects on thecosystems [2,3]. Therefore, a current critical issue for the scien-ific community is to develop and improve treatment technologiesnd procedures to remove those pollutants from the aqueousedium, improving water quality and enhancing its possible

euse.

Since the 90s, an increasing number of chemical processes are

eing modified attending to environmental criteria, and takingnto account several principles of the Green Chemistry [4]. Set in

∗ Corresponding author.E-mail address: jose.peral@uab.cat (J. Peral).

385-8947/$ – see front matter © 2011 Elsevier B.V. All rights reserved.oi:10.1016/j.cej.2011.06.026

© 2011 Elsevier B.V. All rights reserved.

this context, Advanced Oxidation Processes (AOPs) are presentedas environmentally friendly physico-chemical based procedures.They generate some reactive radical species like hydroxyl radical(•OH), a powerful oxidant (E◦(•OH/H2O) = 2.80 V vs. SHE) that non-selectively oxidizes most organics [5]. Among several AOPs (Fenton,photo-Fenton, electro-Fenton, heterogeneous photocatalysis withTiO2, ozonation, etc.), some studies [6–9] have demonstratedthat Fenton reaction (Eq. (1)) based processes present highmineralization efficiencies and minor environmental impacts,especially when solar light is used to assist the iron cycling and,thus, the catalytic character of the reactive system.

H2O2(aq) + Fe(aq)2+ → Fe(aq)

3+ + OH(aq)− + •OH(aq) (1)

However, AOP applicability depends on both water COD con-centration and water flow rate [10]. They are not effective for highvalues of those two variables because large energy and reagentquantities that increase the operational costs would be involved.To solve this disadvantage, AOPs have been coupled with activated

sludge biological treatments [11–13].

In the integrated chemical–biological treatments, an AOP shouldbe carried out until the treated water is readily biodegradableand presents no toxic/inhibitory effects on the activated sludge,

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A. Serra et al. / Chemical Engin

hus allowing complete pollutants removal with a biological treat-ent. Therefore, biodegradability/toxicity tests should be carried

ut along with the AOP to assess those parameters evolution.Commonly, Biological Oxygen Demand in 5 days (BOD5) and Zahn-

ellens are used as biodegradability tests. In both procedures, auitable quantity of biomass from a WWTP is introduced within

polluted sample for 5 or 28 d, respectively, under temperaturend pH controlled conditions. Contrary to BOD5 test, Zahn-Wellenss carried out under continuous aeration [14,15]. However, oper-tional conditions of those tests are clearly different from thosef an aerobic treatment in a WWTP where hydraulic retentionimes (HRT) frequently lower than 40 h are set [16]. This fact cannvolve questionable (overestimated) test results, especially whenhe influent contains toxic/inhibitory substances, because in 5 or8 d slowly biodegradable organic matter might be assimilated byhe activated sludge, whereas that would not be the case for HRTower than 40 h. Something similar happens with routine toxic-ty tests like Microtox®, a test that usually overestimate toxicityecause the Vibrio Fischeri bacteria used is highly sensitive to toxictimulus and lives in marine environments that are very differentrom that of a WWTP biological system [17,18].

Consequently, taking into account an aerobic activated sludgeiological post-treatment, biodegradability and toxicity tests work-

ng with the same type of heterogeneous bacteria and operatingnder conditions similar to the ones in oxidation steps of a WWTPhould be made available. In this sense, respirometry has been sug-ested as an alternative test [13,18]. This is based in the measurend interpretation of the biomass aerobic respiration rate after theddition of a pulse sample. In this study, respirometric assays haveeen selected to determine the biodegradability of organic matter

n the polluted waters and to evaluate the toxic/inhibitory effect ofhe sample on the activated sludge.

Environmental impact assessment associated to polluted waterreatment is also needed in order to determinate those factors withegative impacts to the environment, thereby helping to avoidew environmental problems and/or to evaluate opportunities forrocedures/technology improvement. Life cycle assessment (LCA)ppears as the most accepted, objective and rigorous tool for thisurpose. LCA is based on the assessment of the overall (life cycle)nvironmental impact of a process [19], i.e., it defines the envi-onmental burdens of a product, process or activity by identifyingnd quantifying inputs and outputs of the system and subsystemsenergy, reagents and materials used; emissions to air, water andand of the several processes involved in the life cycle; etc.). LCA

ay be applied to integrated chemical–biological treatments inrder to identify those steps involving the largest environmentalmpacts.

The aim of this work is to treat aqueous solutions pollutedith �-methyl-phenylglycine (�-MPG), a non-biodegradable andighly soluble substance used by the pharmaceutical industry, byoupling an AOP (solar photo-Fenton process) with a biologicalreatment (activated sludge in a sequencing batch reactor), carry-ng out biodegradability and toxicity tests based on respirometricssays, and also evaluating environmental impacts of those labora-ory bench-scale processes.

. Materials and methods

.1. Chemicals and activated sludge

Pure �-MPG from DSM Pharma Chemicals Venlo B.V. was used

s received. Ferrous sulphate heptahydrate (99.5%), used as cata-yst, and sodium sulphite anhydrous (98%), used as oxidant agent,

ere supplied by Merck and Panreac, respectively. Reagent grade3% (w/v) H2O2 was obtained from Panreac. Nutrients used for acti-

Journal 172 (2011) 654– 664 655

vated sludge: magnesium sulphate heptahydrate, calcium chloridedihydrate and ferric chloride anhydrous, were of analytical gradeand supplied by Panreac and Fluka. Buffer solution was obtainedusing ammonium chloride, sodium and potassium hydrogenousphosphate anhydrous and sodium dihydrogen phosphate dihy-drate from Sigma–Aldrich and Panreac.

Solutions were prepared with high-purity water from a Mil-lipore Milli-Q system (resistivity > 18 M� cm at 25 ◦C). The pHwas controlled with analytical grade sulphuric acid or sodiumhydroxide both from Merck. Organic solvents and other chemicalsemployed were either HPLC or analytical grade from Fluka, Panreacand Sigma–Aldrich.

Respirometer and sequencing batch reactor (SBR) were seededwith non-acclimated activated sludge coming from the recircula-tion stage of municipal WWTP from Manresa (Catalonia, Spain),a facility that operates whit single-sludge step. The average ofvolatile suspended solids (VSS) was 2400 ± 200 mg L−1, HRT was20.0 ± 0.8 h and cellular retention time (CRT) was 32 ± 4 d (valuesprovided by Manresa WWTP).

2.2. Analytical techniques and procedures

Treated water mineralization was monitored by measuring totalorganic carbon (TOC, mg L−1 C), using a SHIMADZU TOC-VCSH/CSNanalyzer, previous solution filtration with a Whatman 0.45 �mPTFE filters.

Unicam UV/Vis UV4 spectrophotometer, thermostated at 25 ◦C,was used to determine H2O2 concentration from the light absorp-tion at 408 nm of the titanic-hydrogen peroxide coloured complex[20].

Released NH4+ and NO3

− ions from treated solutionswere quantified by ion chromatography (SHIMADZU VP SeriesHPLC chromatograph coupled with a SHIMADZU CDD 10Avpconductivity detector). NO3

− was determined with a Shim-Pack IC-A1S anion column (100 mm × 4.6 mm (i.d.)) usingan aqueous mobile phase with 2.5 mM phtalic acid and2.4 mM tris(hydroxymethyl)aminomethane (TRIS) (pH ≈ 4) at1.5 mL min−1; the chromatogram displayed the correspondingpeak at tr = 4.0 min. For NH4

+ measurement, the chromato-graph was fitted with a Shodex IC KY-421 cation column(125 mm × 4.6 mm (i.d.)) at 40 ◦C, using a solution of 5 mM tartaricacid, 2 mM dipicolinic acid, and 24.2 mM boric acid as mobilephase at 1.0 mL min−1; the NH4

+ peak appeared at tr = 4.4 min. Thenitrogen removed is calculated by subtracting the N detected insolution in the form of NH4

+ or NO3− from the total N content of

the initial 500 mg L−1 of �-MPG.Total suspended solids (TSS, mg L−1) present in the solution

coming from the Manresa WWTP were quantified following theStandard Methods 2540 D [21]. Volatile suspended solids (VSS,mg L−1), assimilated to biomass quantity, were determined follow-ing Standard Methods 2540 E [22]. Chemical oxygen demand (COD,mg L−1 O2) was determined by the Close Reflux Colorimetric Methodof Standard Methods 5220 D [23]. Dissolved oxygen (DO, mg L−1 O2)in the solution was measured by an EXTECH 407510 sensor.

2.3. Polluted water treatment by AOP

The SPF experiments were carried out as follows. 250 mL ofsynthetic non-biodegradable wastewater containing 500 mg L−1 �-MPG (327 mg L−1 of TOC) were treated at laboratory bench-scale ina wide base open Pyrex glass beaker of 400 mL. All experimentswere performed with 10 mg L−1 Fe2+, vigorously stirring the solu-

tion with a magnetic bar to ensure its homogenization, and keepingconstant both pH (2.9 ± 0.1) and temperature (25.0 ± 0.1 ◦C).

SPF treatments were carried out with H2O2 initial concentra-tions of 36.3 and 72.5 mM, corresponding to half the stoichiometric

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56 A. Serra et al. / Chemical Engin

1/2 st.) and the stoichiometric (st.) concentration of H2O2 neededo completely mineralize �-MPG as shown in reaction (2):

9H11NO2(aq) + 24H2O2(aq)

→ 9CO2(g) + NO3(aq)− + H(aq)

+ + 29H2O(l) (2)

Solar wastewater treatments were performed in sunny and clearays during the months of June–August at the Barcelona laboratory.he details concerning light irradiance can be found elsewhere [9].

.4. Respirometric assays

Treated solutions coming from the SPF processes were condi-ioned before addition to the respirometer. In the conditioning step,he solution was neutralized with NaOH 1 M. Then, 8.0 mL of thisolution were diluted to 10.0 mL in a flask that contained the suit-ble amount of sodium sulphite to oxidize the remaining hydrogeneroxide. Afterwards, exceeding sulphite was oxidized with a syn-hetic air flow to avoid its interference in the respirometric assays.

Biodegradability/toxicity/inhibition assays were carried outsing an off-line and closed LSS (Liquid-Static-Static) respirome-er, i.e., dissolved oxygen concentration was measured in the liquidhase (L) and both liquid and gaseous phases were static (S,S) [24].he setup was run with 300 mL of solution in a reactor of 350 mL ofapacity. VSS concentration was around 3000 mg L−1. The solutionemperature was maintained at 25.0 ± 0.5 ◦C with a water recircu-ation jacket connected to an external thermostat, and pH was seto 7.0 ± 0.5. During each run, the solution was vigorously stirredith a magnetic bar to ensure its homogenization. 6 mL of 1 g L−1

-allylthiourea were added to avoid nitrification process interfer-nces. Moreover, suitable amounts of nutrients and buffer solutionere added to the respirometer [14].

The procedure followed to determine toxic/inhibitory behaviournd biodegradability is a variation of the procedure proposed byuisasola et al. [25] with a LFS (Liquid-Flowing-Static) respirom-ter. Nevertheless, the operational simplicity of the LSS used inhe present study did not allow the distinction between toxic andnhibitory effects and, consequently, no differences between bothoncepts will be considered here.

Once the respirometer solution was loaded with nutrients, sett suitable pH and temperature, and saturated with oxygen, fouronsecutive respirometric tests were carried out for each sample.e-aeration of the solution before every single test was needed tonsure oxygen saturation:

0) Endogenous respiration was determined. The activated sludgeshould be starved overnight to ensure endogenous conditions.Subsequently, and after oxygen flow was stopped, DO temporalevolution was measured. The straight slope obtained (DO vs.time) corresponds to oxygen uptake rate (OUR), in this case,OURendogenous.

1) A pulse of a readily biodegradable substance (acetic acid),as a control substrate, was added to the respirometer. TheCOD added to the respirometer (CODadded) was 5 mg L−1 O2.After determination of OURtotal, OURexogenous corresponding toacetic acid can be obtained by subtracting OURendogenous fromOURtotal.

2) Once the control substrate used in the previous pulse was con-sumed, and under new oxygen saturation conditions, a pulseof polluted solution (CODadded = 5 mg L−1 O2) was added. Thebiodegradability of the solution can be calculated by using Eq.(3):

%Biodegradability

={

1 −[

(OURCH3COOH) − OURpolluted water

(OURCH3COOH)1

]}× 100 (3)

Journal 172 (2011) 654– 664

(3) In the last pulse, the same quantity of the control substrate, i.e.,CODadded (acetic acid) = 5 mg L−1 O2, is added to the solution.The difference of OURexogenous between the two control sub-strate pulses allows to determine the toxic/inhibitory effect onthe biomass of the polluted water pulse (Eq. (4)):

%Toxicity/Inhibition

=[

(OURCH3COOH)1 − (OURCH3COOH)3(OURCH3COOH)1

]× 100 (4)

Each respirometric test was repeated twice with new non-acclimated activated sludge to ensure reproducible results.

2.5. Experimental setup for the biological treatment

Aerobic biological treatment was carried out in two laboratoryscale suspended biomass SBR running in parallel: one of them wasfed with the solution coming from the SPF that used a 36.3 mMhydrogen peroxide initial concentration; and the other one wasfed with the solution from the SPF processes with 72.5 mM ini-tial concentration of that reagent. The operating liquid volume was500 mL made up with 250 mL of the polluted solution and 250 mLof a 600 mg L−1 VSS activated sludge. The mixture was continu-ously stirred at 500 rpm and oxygenated by means of a syntheticair stream that allowed a dissolved oxygen concentration above4 mg L−1 O2. The reactors were kept at room temperature, and pHwas set at 7.0 ± 0.5 by means of a buffer solution. Suitable quantitiesof nutrients were also added to the solution [14].

15 cycles were carried out to complete each SBR experiment.During the 10 first cycles, HRT was 2 d, i.e., the outflow was set to250 mL d−1. For the last 5 cycles, with the activated sludge moreacclimatised, the outflow was increased to 500 mL d−1 (HRT = 1 d).

Each SBR cycle was made up of the following four steps (takinginto account a HRT of 2 d): (1) fill period (5 min), where 250 mLof the polluted solution (previously neutralized), nutrients andbuffer solution were added to 250 mL of activated sludge, andaeration-agitation started; (2) aeration-reaction period (23 h); (3)settle period (55 min) where agitation and aeration were switchedoff and, therefore, activated sludge settled down; (4) draw period(5 min) where 250 mL of treated effluent were evacuated (seeFig. 1).

TOC, ammonium and nitrate were measured in both the influentand effluent. Periodical VSS, DO and pH measurements were carriedout to ensure suitable experimental conditions.

2.6. LCA methodology

Details of the principles and structure of an LCA study can befound elsewhere [26]. The functional unit defined in this work inorder to compare water treatment strategies was the removal of93% of TOC present in 250 mL of water polluted with 500 mg L−1 of�-MPG. Processes that lead to environmental impacts and, there-fore, included in the LCA, as well as the main hypothesis andlimitations assumed in this study, are summarized in Table 1.

The data used in this LCA study was handled with the SimaPro 7.0software. The Swiss database Ecoinvent version 1.2 [27] was takenas the source of background inventory data. The following datasetshave been considered for chemicals and energy consumed, as wellas exceeding sludge management:

(1) Low voltage electricity, production UCTE at grid [28].(2) Transport, lorry 16 t/RER [29].(3) Hydrogen peroxide, 50% in H2O, at plant/RER [30].(4) Iron (II) sulphate, at plant/RER [28].

A. Serra et al. / Chemical Engineering Journal 172 (2011) 654– 664 657

Table 1Scope for the present LCA study (involved processes and main hypothesis and limitations assumed).

Included processesSPF:1. Production of chemicals consumed in the SPF process, including extraction of resources, production of the corresponding chemicals and the different

transportation steps.2. CO2 release produced by the mineralization processes.SBR:1. Production of chemicals consumed (nutrients and oxygen), including extraction of resources, production of the corresponding chemicals and the different

transportation steps. Total oxygen consumed is due to oxidation (2) and nitrification processes (reaction 4 and 5), as well as endogenous respiration (3).2. Production of electricity for mechanical aeration equipments to obtain suitable oxygen conditions.3. CO2 release produced by the mineralization processes (2) and endogenous respiration (3) in the SBR.4. Release of chemicals in the final treated effluent.Exceeding activated sludge management by controlled landfill:1. Thickening, dewatering and stabilization processes.2. Stabilized sludge transport until the landfill.3. Atmospheric and hydrospheric emissions from sludge management.Main limitations and hypothesis assumed(1) Results refer to experiments made at laboratory scale. Thus, no extrapolation is possible to pilot plant or industrial scale.(2) Infrastructure and equipments have not been optimized and, consequently, they have not been considered in the analysis(3) Chemical pre-treatments (i.e., acidification) and post-treatments (i.e., neutralization) are not taken into account because they are the same for several

treatments.(4) Energy consumed is delivered by the European electrical grid (average of the Union for the Coordination of Transmission of Electricity, UCTE, net).(5) Reagents consumed are produced in Spain, with a transport distance to the consumer of 50 km, and being carried by 16 t-trucks.(6) It is assumed that added nutrients to the influent are incorporated by biomass.(7) Exceeding activated sludge is calculated by means of reaction (2).(8) It is assumed that VSS/TSS = 0.76 ± 0.03 (n = 13, = 0.05).(9) It is assumed that dry matter (DM) of sludge corresponds to TSS.

) is C8

, are c

(

(

(

(

(

(10) It is assumed that stabilized fresh sludge composition (dry matter and water(11) Emissions from landfill, due to anaerobic digestion of stabilized fresh sludge

model

5) Oxygen is supplied by mechanical aeration taking into account1 kWh consumed by 1.5 kg O2 produced by air diffusers [31].

6) Exceeding sludge management: thickening, dewatering andstabilization processes [32]. In the thickening step, 4 kg of acry-lonitrile polymer and 50 kWh are consumed per ton of drymatter (DM) and in the dewatering one, 5 kg of acrylonitrilepolymer and 40 kWh are needed per ton of DM, achieving a26% DM content in the sludge. In the stabilization step, 200 kgof CaO and 5 kWh are consumed per ton of DM, the final DMcontent of the stabilized fresh sludge being 31%.

7) ORWARE theoretical model [33] assumes that stabilized freshsludge (C8H15O4N) is anaerobically digested producing biogasand leachate according to reaction (5):

2C8H15O4N(s) + 6H2O(l) → 9CH4(g) + 7CO2(g) + 2NH3(g) (5)

With this model, data of air/water/land distribution for sev-eral compounds present in a landfill are obtained. Temporalscale considered for digested sludge biochemical stabilizationis 100 years. The model considers:- 50% biogas capture efficacy [34]. Biogas is burned in a torch,

being the CH4 totally oxidized to CO2. Each kg of CH4 burnedreleases 7.5 g of N-NOx, in form of NO2.

- 90% leachate capture efficacy [35]. Leachate is treated in aWWTP with 90% of yield for COD and BOD7 [36], and 80% ofyield for nitrification process [34] where 50% of the nitrateproduced is denitrificated and, the rest, is emitted to hydro-sphere.

- Non captured biogas and leachate are emitted to atmosphereand hydrosphere, respectively.

8) Electricity consumption for biogas pumping: BUWAL 250database [37]. It is calculated that 1.35 kWh per ton of freshstabilized sludge produce 200 m3 of biogas.

9) Gasoil: Diesel BUWAL 250 database [37]. Diesel is con-

sumed by landfill machinery that lay out and compact theresidues. It is calculated 1.8 L of diesel consumption per tonof residue (i.e., fresh sludge, acrylonitrile polymer and CaO)[36].

H15O4N [7].alculated by means of ORWARE (Organic Waste Research) theoretical

The inventoried data were classified, i.e., allocated to the fol-lowing impact categories: global warming potential (GWP), ozonedepletion potential (ODP), aquatic eutrophication potential (AEP),acidification potential (AP), human toxicity potential (HTP), freshwater aquatic ecotoxicity potential (FAEP), marine aquatic eco-toxicity potential (MAEP), terrestrial ecotoxicity potential (TEP),photochemical oxidation potential (POP) and abiotic resourcedepletion potential (ADP).

Afterwards, the classified data were characterized, i.e., quanti-fied using equivalency factors from literature [38–40]: (a) kg CO2stands for the equivalent mass of the representative greenhousegas CO2 in GWP; (b) kg of CFC-11 stands for the equivalent massof the most aggressive CFC towards ozone in ODP; (c) kg of PO4

3−

stands for the equivalent mass of the representative eutrophica-tion forming substance PO4

3− in AEP; (d) kg of SO2 stands for theequivalent mass of the main acid rain forming substance SO2 inAP; (e) kg of 1,4-DB stands for the equivalent mass of the repre-sentative toxic compound 1,4-dichlorbenzene in HTP, FAEP, MAEPand TEP; (f) kg of C2H4 stands for the equivalent mass of the rep-resentative smog forming hydrocarbon C2H4 in POP; and (g) kg ofSb stands for the equivalent mass of a scarce resource as antimonyin ADP.

3. Results and discussion

Due to the non-biodegradable nature of �-MPG, a highly watersoluble molecule, the removal of this pollutant was carried outby coupling and AOP like photo-Fenton with an aerobic biologi-cal treatment. The success of the coupling depends on the choiceof an appropriate time of reaction for the first step. A long timeof chemical reaction would waste an excess of chemical reagentsproducing an aqueous solution with such a low organic load thatit could hardly be used to feed a biological reactor. Conversely,a short reaction time would produce a non-biodegradable inter-

mediate solution that could damage the biological reactor sludge.Thus, the detection of the lower reaction time needed to obtain aneasily biodegradable intermediate solution is critical. In this senserespirometry, a fast, simple and versatile technique was chosen to

658 A. Serra et al. / Chemical Engineering Journal 172 (2011) 654– 664

p of t

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vtotccnlt

(stsarTnrr

Fr

Fig. 1. Experimental setu

onitor the changes of solution biodegradability along chemicaleaction.

.1. Respirometric assays

In respirometric assays, the response of non-acclimated acti-ated sludge in front of an external substrate pulse depends onhe type and concentration of the substrate. The reproducibilityf the technique was previously evaluated by means of additiono the same sludge solution of consecutive pulses of the sameoncentration of a readily biodegradable substrate. In this sense,onsecutive pulses of acetic acid (5 mg COD L−1) into 3000 mg L−1 ofon-acclimated sludge gave an exogenous OURacetic acid that oscil-

ated around 0.51 ± 0.03 mg O2 L−1 min−1 with CV < 6%. The facthat CV is lower than 10% indicates acceptable reproducibility.

On the other hand, respirometric tests with several compoundspropionic acid, hydroquinone, catechol, p-benzoquinone, etc.)howed good agreement with previously published data [13]. Onhis basis, the biodegradability of �-MPG (target pollutant in thistudy) was evaluated. Fig. 2 exemplifies the followed procedure:ssessment of respiration after no substrate addition (endogenousespiration), acetic acid pulse, �-MPG pulse, and acetic acid pulse.

he OUR obtained with those pulses is also shown. It is worth toote that DO is measured after 60 s of the hermetic closing of theespirometer (estimated as the time needed for activated sludgeesponse).

ig. 2. DO time course in the 300 mL LSS respirometer with 3000 mg L−1 of VSS, at pH = 7.0espiration; (�) acetic acid pulse, CODadded = 5 mg L−1 O2; (©) �-MPG pulse, CODadded = 5 m

he biological treatment.

As shown in Fig. 2 the good linear relation found between DO andtime for each one of the plots allows the assessment of OURexogenous.As can be seen, after correction of the OURs by subtraction of theOURendogenous, the data clearly indicate that acetic acid (1st pulse),a readily biodegradable substance, gives a high OUR value (0.5 mgO2 L−1 min−1), whereas �-MPG (2nd pulse), a recalcitrant sub-stance, presents an OUR value close to zero. The OUR obtained foracetic acid was similar for both the 2nd and 4th pulse (0.5 mg O2L−1 min−1), indicating that �-MPG had non toxic/inhibitory effecton the activated sludge, at least within the time of the respirometricassay.

The biodegradability/toxicity/inhibitor behaviour of water gen-erated during two different SPF processes (36.3 mM and 72.5 mMH2O2 initial concentrations) was evaluated along the treatmenttime, i.e., every 30 min a sample of the treated water (COD adjustedto 5 mg L−1 O2 by suitable dilutions) was added as substrate of thesecond pulse of the corresponding respirometric assay. In Table 2,the intermediate effluents produced from both SPF processes areclassified as readily biodegradable, inert or toxic/inhibitory. Forthose respirometric tests with lower OUR in the third pulse thanin the first one, a toxic/inhibitory behaviour of the wastewaterexists, while a non-altered OUR means no (or low) toxic/inhibitory

effect. The intermediate polluted water biodegradability is evalu-ated comparing the OUR obtained in the second pulse with thatfound in the first one: an OURpolluted water similar to OURacetic acidindicates that the treated water is readily biodegradable, while an

± 0.5 and 25.0 ± 0.5 ◦C when the following compounds are added: ( ) endogenousg L−1 O2; ( ) acetic acid pulse, CODadded = 5 mg L−1 O2.

A. Serra et al. / Chemical Engineering

Fig. 3. Time course of (�) biodegradability and (�) toxic/inhibitory percentage for2 −1 ◦

p7

Os

twcsaascbfgaaattr1bct15i

50 mL of a 500 mg L �-MPG solution, at pH = 2.9 ± 0.1 and 25.0 ± 0.5 C, after solarhoto-Fenton treatment with 10 mg L−1 Fe2+ and [H2O2]initial: (a) 36.3 mM and (b)2.5 mM.

URpolluted water close to zero corresponds to a slowly biodegradableubstrate.

Taking into account Eqs. (3) and (4), Fig. 3 displays the evolu-ion of biodegradability and toxic/inhibitory behaviour of treatedaters along the SPF process time. The data of the SPF pro-

ess carried out with H2O2 initial concentration of 36.3 mMhow that, although the water initially polluted with �-MPG hadn inert behaviour, some of the intermediate solutions gener-ted along the treatment had toxic/inhibitory effect on activatedludge, especially during the first hour when percentages of toxi-ity/inhibition were close to 80% (30 min) and 30% (60 min), andiodegradability was closed to 0%. This fact could be related toormation of aromatic by-products like catechol or hydroquinone,enerated early in the �-MPG oxidation sequence as a result of 2-mminopropionic group release and aromatic ring hydroxylation,s postulated by Serra et al. [8]. It has been previously reported that,ccording to their concentration, those intermediates could haveoxic/inhibitory behaviour on the biological system [13]. Duringhe second hour of treatment, the toxic/inhibitory character waseduced to 10–15%; however, biodegradability remained around0–15%. Therefore, at that time the treated solution had an inertehaviour and it was not yet suitable for biological treatment. Inontrast, at 150 min, biodegradability increased up to 85%, although

he toxic/inhibitory character was around 20%, and 30 min later (at80 min), biodegradability and toxicity/inhibition drop to 10% and%, respectively. The biodegradability percentage at 150 min could

ndicate the presence of intermediates such as some short car-

Journal 172 (2011) 654– 664 659

boxylic acids like fumaric, succinic, formic and oxamic [8] that couldbe readily biodegraded by the activated sludge. Simultaneously, thefact that biodegradability fell off 30 min later could indicate thatthose by-products were also rapidly oxidized by photo-Fenton pro-cess. Therefore, the aqueous solution generated after 150 min of SPFtreatment ([H2O2]initial = 36.3 mM), when TOC = 112 mg L−1, was asuitable influent for biological treatment.

A previous work by Oller et al. [41] treated the same initial con-centration �-MPG aqueous solutions, at pilot plant scale, with SPFprocesses under different conditions, and they used Zahn-Wellenstest to evaluate water biodegradability, and carried out no-test fortoxicity. Those authors found that a 93 mg L−1 TOC solution (slightlylower than in the present study) was suitable for the coupling witha biological treatment.

On the other hand, the SPF treatment that used a H2O2 ini-tial concentration of 72.5 mM showed a toxic/inhibitory increaseduring the first hour of treatment, reaching 60% at 60 min, whileat that time biodegradability was around 25%. This seems toindicate that the toxic/inhibitory aromatic by-products generatedfrom �-MPG oxidation were more rapidly generated and oxidizedunder those more powerful SPF conditions. From 60 to 90 min,both biodegradability and toxic/inhibitory effect fell off to val-ues close to 0% and 10%, respectively, indicating inert behaviourof the treated aqueous solution at 90 min. From 120 to 180 minbiodegradability increased up to 100%, i.e., similar to the biodegrad-ability of acetic acid, the substance that has been taken as readilybiodegradable reference. Moreover, the toxic/inhibitory behaviourwas reduced from 10% at 120 min to 0% at 180 min. That allowsconcluding that the aqueous solution generated after 120 min ofSPF treatment ([H2O2]initial = 72.5 mM), when TOC = 43 mg L−1, wasa suitable influent for biological treatment. This result is also inagreement with Oller et al. [41], who found that biodegradabilityreached 80% in the sample with TOC = 47 mg L−1 in only 2 d of theZahn-Wellens test.

Moreover, it is worth to note that the biodegradability increasetakes place after the removal of oxalic acid, one of the more concen-trated carboxylic acids detected in those waters [8], a compoundthat has been classified like inert in previous studies [13].

Since both SPF processes seem to generate suitable influentsfor biological treatment, activated sludge SBR experiments werecarried out to verify it.

3.2. SPF-biological coupling at laboratory bench-scale

250 mL of the �-MPG solutions pre-treated with the SPF processwere the influent (after neutralization) of two 500 mL biologicalSBR running in parallel, and following the procedure and conditionsindicated in Section 2.5. Both SBR were run uninterruptedly during15 cycles. HRT of the first 10 cycles was 2 d, while for the last 5cycles, when the biomass was better acclimatised, it was reducedto 1 d.

Previous to the SPF–SBR coupling, blank experiments werecarried out in order to evaluate residual TOC corresponding tometabolites produced from cellular activity (600 mg L−1 VSS in bothSBR). Hence, two 250 mL acetic acid solutions were prepared, onewith a TOC of 40 mg L−1 and the other one with 110 mg L−1, thus,simulating influent TOC values. TOC resulting for HRT of 1 and 2 dwere 7 ± 1 and 6 ± 1 mg L−1, respectively.

Table 3 presents the data of TOC, ammonium and nitrate ionsmeasured for each SBR influent and effluent, as well as % TOCand % nitrogen removal for global integrated chemical–biologicaltreatments. Values are the arithmetic average of the three last

cycles (n = 3, = 0.05). As can be seen in Table 3, both SBR achievedsignificant TOC removal (higher than 75%). However, the resid-ual TOC of the SBR that worked with influent from SPF with[H2O2]initial = 36.3 mM (25 ± 3 mg L−1) was always higher than TOC

660 A. Serra et al. / Chemical Engineering Journal 172 (2011) 654– 664

Table 2Transient biodegradability and toxic/inhibitory behaviour of the SPF treatment of 250 mL of a 500 mg L−1 �-MPG solution, at pH = 2.9 ± 0.1 and 25.0 ± 0.5 ◦C. Treatmentscarried out with 10 mg L−1 Fe2+, and two different initial H2O2 concentrations: 36.3 and 72.5 mM.

SPF process with [H2O2]initial = 36.3 mMReaction time/min 0 30 60 90 120 150 180

OUR/mg O2 L−1

min−1

1st acetic acid pulse (COD = 5 mg O2 L−1) 0.5 0.2 0.4 0.4 0.4 0.1 0.6Sample pulse (COD = 5 mg O2 L−1) 0.0 0.0 0.0 0.0 0.0 0.1 0.12nd acetic acid pulse (COD = 5 mg O2 L−1) 0.5 0.0 0.3 0.4 0.4 0.1 0.6

Effect on the activated sludge Inert Toxic/Inhib. Toxic/Inhib. Inert Inert Biodegrad. Inert

SPF process with [H2O2]initial = 72.5 mMReaction time/min 0 30 60 90 120 150 180

OUR/mg O2 L−1

min−1

1st acetic acid pulse (COD = 5 mg O2 L−1) 0.5 0.6 0.5 0.3 0.2 0.6 0.6Sample pulse (COD = 5 mg O2 L−1) 0.0 0.2 0.1 0.0 0.2 0.7 0.62nd acetic acid pulse (COD = 5 mg O2 L−1) 0.5 0.4 0.2 0.3 0.2 0.6 0.6

Effect on the activated sludge Inert Toxic/Inhib. Toxic/Inhib. Inert Biodegrad. Biodegrad. Biodegrad.

Table 3TOC, ammonium, and nitrate concentration, and their elimination percentage, in the influent and effluent of the two SBR of 500 mL of capacity, 600 mg L−1 VSS, and HRT = 1 d.Results are provided as an average of the three last runs (n = 3; = 0.05).

SPF ([H2O2]initial = 36.3 mM) effluent SPF ([H2O2]initial = 72.5 mM) effluent

TOC SBR influent/mg L−1 (112 ± 2) (43 ± 4)TOC SBR effluent/mg L−1 (25 ± 3) (7 ± 2)TOC removed (SBR)/% (78 ± 2) (84 ± 3)TOC removed (SPF + SBR)/% (93 ± 1) (98.0 ± 0.5)NH4

+ SBR influent/mg L−1 (40 ± 1) (43 ± 1)NO3

− SBR influent/mg L−1 – (6.1 ± 0.5)

cTebSttl

tm(frrD

teaopwvs

v

tamisiicp

NH4+ SBR effluent/mg L−1 (34 ± 1)

NO3− SBR effluent/mg L−1 (28 ± 9)

N removed (SPF+SBR)/% (78 ± 5)

orresponding to metabolites of biological activity (7 ± 1 mg L−1).his might be a consequence of the fact that only 85% of this influ-nt was biodegradable and, moreover, it had some toxic/inhibitoryehaviour. At the same time, the SBR loaded with the influent of thePF with [H2O2]initial = 72.5 mM showed a similar residual TOC thathe blank experiment, an expected behaviour taking into accounthat this influent had a similar biodegradability than acetic acid andow toxic/inhibitory effect.

Nevertheless, under the operational conditions ofhis study, both integrated chemical–biological treat-

ents, i.e., SPF ([H2O2]initial = 36.3 mM)–SBR and SPF[H2O2]initial = 72.5 mM)–SBR, are highly effective strategiesor �-MPG removal because global TOC removals of 93% and 98%,espectively, are accomplished with them. Emission of such a lowesidual TOC to the hydrosphere may be acceptable according toirective 91/271/EEC [42].

On the other hand, the possibility of nitrification process duringhe biological treatment was studied. In this sense, both SPF efflu-nts contained high quantities of ammonium (around 40 mg L−1)nd low or negligible values of nitrate (Table 3). Several studies havebserved that nitrification rate of biological nitrogen removal (BNR)rocesses depends on the COD/N relationship [43,44], decreasinghen COD/N increases. This fact has also been observed in acti-

ated sludge biological systems [45]. Nitrification rate has beenatisfactorily predicted by Eq. (6) [46]:

nitrification = Flow([NH4+ − N]affluent − [NH4

+ − N]effluent)VSBR [VSS]SBR

(6)

In this study, and due to the relative elevated COD/N rela-ion of the influents, i.e., COD/N (SPF, [H2O2]initial = 36.3 mM) = 7.5nd COD/N (SPF, [H2O2]initial = 72.5 mM) = 2.5, low nitrification ratesight be expected. However, due to activated sludge heterogene-

ty it is necessary to check the SBR behaviour along all the cyclesequence because the nitrification process can mainly take place

n the early cycles when the sludge is not acclimated and, hence,s more heterogeneous. Therefore, ammonium and nitrate con-entration along the different SBR cycles were quantified andresented in Fig. 4. Table 3 collects the arithmetic average of

(38 ± 2)(14 ± 1)(76 ± 4)

concentration of both nitrogen species in the SBR influent andeffluent, taking into account the three last cycles (n = 3, = 0.05).Nitrite concentration was negligible along all cycles for bothSBR. As shown in Fig. 4, the nitrification process was especiallysignificant during the four first cycles, with a nitrification rateaverage for both SBR between 0.025 and 0.028 mg N mg−1 VSSd−1, and consumed nearly all the initial ammonium. In a bio-logical nitrogen removal process, Carrera et al. [46] obtained anitrification rate stabilized at 0.032 mg N mg−1 VSS d−1 for influ-ent COD/N ratios higher than 2.5. The data of the present study,obtained with non-acclimatised activated sludge biological sys-tems, evidence that nitrification takes place at a rate that iscomparable to the one obtained with a BNR system. This canprobably be due to the heterogeneity of the municipal WWTPsludge, where some nitrification bacteria can exist. As the activatedsludge acclimatise, the nitrification process was progressivelyreduced, achieving negligible nitrification rates in the three lastcycles of both SBR, i.e., 0.011 ± 0.001 mg N mg−1 VSS d−1 with theCOD/N = 7.5 mg COD mg N−1 influent, and 0.005 ± 0.003 mg N mg−1

VSS d−1 with the COD/N = 2.5 mg COD mg N−1 influent. Theseresults might be interpreted by taking into account that bothinfluents could contain quantities of some substances that, inthe first SBR cycles, can inhibit the nitrification process; how-ever, after long exposure times (fifth cycle in this study),inhibitory effects might have been replaced by toxic ones becauseammonium-oxidizing bacteria have been progressively eliminated.Summarizing, under the current conditions, the SBR is not an effec-tive system to combine organic matter oxidation and nitrificationprocesses.

It is worth to note that nitrogen balance was not complete;only around 70–80% of total initial nitrogen was detected for theSBR effluents. However, it should be taken into account that somenitrogen may be incorporated by the biomass. For both effluents,the total quantity of inorganic nitrogen was around 30 mg N L−1.

The emission of those effluents should be carried out in areas nonsensitive to eutrophication process because, according to Directive91/271/EEC [42], in those areas the maximum nitrogen emissionhas to be of 10–15 mg N L−1.

A. Serra et al. / Chemical Engineering Journal 172 (2011) 654– 664 661

Fig. 4. Temporal evolution of ammonium and nitrate in the SBR effluent during 15ca

3

t5s

cottdsmrtttpftp

Fig. 5. Relative environmental impacts (100% is assigned to the highest

of the sources that had environmental impacts, Fig. 5 displaysthe relative impacts (100% is assigned to the highest impact fac-tor in each impact category) of scenario 2 (chemical–biologicalcoupling). As can be seen, in this integrated chemical–biological

ycles. Treatment of 250 mL of influent coming from (�) SPF, [H2O2]initial = 36.3 mM,nd (�) SPF, [H2O2]initial = 72.5 mM.

.3. LCA

Taking into account all the introductory details of Section 2.6,he LCA of the treatment of 250 mL of an aqueous solution of00 mg L−1 of �-MPG was carried out by considering the followingcenarios:

Scenario 1: 93% TOC removal using the solar photo-Fenton processwith [H2O2]initial = 72.5 mM.Scenario 2: 93% TOC removal using the solar photo-Fenton pro-cess ([H2O2]initial = 36.3 mM) coupled to activated sludge biologicaltreatment in a SBR.

In this way, a full chemical scenario is compared with ahemical–biological coupling. Table 4 inventories relevant databtained in those experiments: chemicals and electricity consump-ion, process time, emissions of chemicals to air and water, andransport units (tons of carried chemicals multiplied by travelledistances) per functional unit (93% TOC removal) for both scenarioselected. As can be seen, scenario 1 was the most efficient treat-ent strategy in terms of reaction time, since it needed 3 h to

emove the 93% of TOC. The present LCA study is aimed at iden-ifying the environmental impact of each scenario and, at the sameime, to evaluate the sources (impact factors) that might contributeo the global environmental impact of the chemical–biological cou-ling. In a previous work by the present authors [9], carried out

or EF (electro-Fenton), SPEF (solar photoelectron-Fenton) and SPFreatment of the same pollutant, the obtained results showed thatroduction of hydrogen peroxide consumed as a reagent was the

impact factor in each impact category) of the treatment (93% TOC conver-sion) of 250 mL of water polluted with 500 mg L−1 �-MPG by means of theSPF([H2O2]initial = 36.3 mM)–SBR coupling.

main impact factor for the majority of impact categories in the SPFprocess.

In order to visualize the nature and the relative contribution

Fig. 6. Relative environmental impacts of several impact categories for the two sce-narios of 93% TOC conversion considered (100% is assigned to the highest impactscenario): (�) SPF ([H2O2]initial = 36.3 mM)–SBR; ( ) SPF ([H2O2]initial = 72.5 mM).

662 A. Serra et al. / Chemical Engineering Journal 172 (2011) 654– 664

Table 4Chemical consumption and reaction products, electrical operating parameters, transport requirements and reaction times per functional unit (93% TOC removal) in thedifferent scenarios considered.

93%TOC removal

SPF, [H2O2]initial = 72.5 mM (SPF, [H2O2]initial = 36.3 mM) − biological treatment

SPF, [H2O2]initial = 36.3 mM SBR Sludge management

Inputs/mg

H2O2 6.3 × 102 3.1 × 102 – –FeSO4 6.8 6.8 – –Na2SO4 – – – –O2 – – – –Mg2SO4 – – 23 –CaCl2 – – 36 –Acrylonitrile polymer – – – 0.23CaO – – – 5.2Diesel – – – 0.13

Transport/t km

H2O2 6.3 × 10−5 3.1 × 10−5 – –FeSO4 3.4 × 10−7 3.4 × 10−7 – –Na2SO4 – – – –O2 – – – –Mg2SO4 – – 1.1 × 10−6 –CaCl2 – – 1.8 × 10−6 –Acrylonitrile polymer – – – 4.7 × 10−9

CaO – – – 1.0 × 10−7

Diesel – – – 2.5 × 10−9

Stabilized fresh sludge – – – 1.7 × 10−6

Outputs/mg

Atmosphereemissions

CO2 2.8 × 102 2.0 × 102 3.3 × 102 38CH4 – – – 4.0NH3 – – – 2.1 × 10−2

NOx – – – 0.12

Hydrosphereemissions

TOC 5.7 – 6.1 5.2 × 10−3

NH4+ 11 – 8.5 0.40

NO3− 1.5 – 7.1 2.7

H2O2 – – – –Fetotal 2.5 – – –Na+ – – – –SO4

2− 4.3 – 13 –Cl− – – 18 –

Electricity/kWh Electrolysis – – –Mechanic aeration – – 1.6 × 10−4 –Sludge drying – – – 5.2 × 10−7

Sludge thickened – – – 4.2 × 10−7

Sludge settled – – – 5.2 × 10−8

Landfill machinery – – – 6.4 × 10−7

Biogas pumping – – – 4.3 × 10−8

tStotcpp(pgedipmb(cb

Process time/h 5.0

reatment, the main environmental contributions come from thePF process. In the global treatment, the main impact factor forhe majority of impact categories was attributed to productionf hydrogen peroxide consumed in the AOP. One exception washe aquatic eutrophication potential (AEP), where 85% of theontribution was due to the emission of soluble inorganic N com-ounds in the final SBR effluent. CO2 emissions from the SPFrocess and biological treatment had also a significant weight45%) in global warming potential (GWP), whereas iron (II) sul-hate consumption and transport negligibly contributed to thelobal environmental impact (less than 0.4% for each impact cat-gory). For the activated sludge treatment, after the contributionue to the emissions, production of electricity consumed dur-

ng mechanical aeration with air diffusers (3–42%), followed byroduction of nutrients (2–14%), had a significant weight in theajority of impact categories, except for AEP where the contri-

ution was lower than 1%. In general, sludge excess management26 mg of dry matter) by means of controlled landfill, had a lowontribution to global environmental impact, being the anaero-ic digestion emissions the main contributor impact factor: 22%

2.5 12 –

to the photochemical oxidation potential (POP) due to emissionand reaction of volatiles as CH4 and NOx, generated during theburning in the torch, 12% to GWP as a consequence of CO2 andCH4 emission, 10% to AEP due to N compounds in the leachateemitted to the hydrosphere, and 6% to acidification potential (AP),mainly associated to NOx emission to atmosphere and hydro-sphere.

Fig. 6 illustrates the comparison of the LCA of the two sce-narios considered. As can be seen, integrated chemical–biologicaltreatment is slightly better than the SPF process, the later pre-senting higher environmental impacts for the majority of impactcategories, especially for human toxicity potential (HTP) (48%),fresh water aquatic ecotoxicity potential (FAEP) (43%), ozone deple-tion potential (ODP) (41%) and ADP (33%), higher impacts that arerelated to the production of hydrogen peroxide. On the contrary,chemical–biological coupling had higher environmental impacts

than SPF for GWP (13%), probably due to CO2 emissions of endoge-nous respiration. AEP and POP contributions of both scenarios werecomparable. These results agree with previous LCA works at labora-tory scale [7]. Nevertheless, it has to be taken into account that the

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A. Serra et al. / Chemical Engin

onsideration of infrastructure and equipments could change theonclusions of the LCA study. In this sense, the final choice of pol-utant removal strategy should be balanced considering that, whilehe integrated chemical–biological treatment has slightly lowernvironmental impacts, the SPF process takes place in a shorterime. At pilot or industrial scale, a shorter reaction time mightnvolve the need for less reaction volume and less infrastructure.

. Conclusions

Respirometry (following the same procedure proposed byuisasola et al. [25]) is presented as alternative test for quantifica-

ion of both biodegradability and toxic/inhibitory effect of pollutedater, and can be used as suitable technique for the determination

f the appropriate time of coupling in a chemical–biological waterreatment scheme.

Laboratory scale SPF process coupled to activate sludge SBR bio-ogical treatment is an effective strategy for treatment of aqueousolutions polluted with 500 mg L−1 of biorecalcitrant �-MPG. SPFrocess and integrated SPF-biological treatment are presented asnvironmentally friendly strategies to treat biorecalcitrant organicatter, at least at laboratory bench-scale. Future studies should be

arried out to evaluate environmental impacts of both strategies atilot plant or industrial scale.

Comparison of the LCA of the two scenarios considered showshat integrated chemical–biological treatment is slightly betterhan the SPF process, the later presenting higher environmentalmpacts for the majority of impact categories, although it has to beaken into account that consideration of infrastructure and equip-

ents in an more complete LCA could change the conclusions ofhe study.

cknowledgements

The authors want to thank the Ministerio de Ciencia e Inno-ación of Spain for financial support through the CTQ2008-00178esearch project and for Spanish government fellowship FPU AP005-26 granted to Anna Serra Clusellas. We also want to thank Dr.ulián Carrera of Chemical Engineering Department of UAB for hisuide in the respirometric assays.

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