Accepted Manuscript Incorporation of electrochemical advanced oxidation processes in a multistage treatment system for sanitary landfill leachate Francisca C. Moreira, J. Soler, Amélia Fonseca, Isabel Saraiva, Rui A.R. Boaventura, Enric Brillas, Vítor J.P. Vilar PII:

S0043-1354(15)30019-1

DOI:

10.1016/j.watres.2015.05.036

Reference:

WR 11311

To appear in:

Water Research

Received Date: 23 March 2015 Revised Date:

8 May 2015

Accepted Date: 19 May 2015

Please cite this article as: Moreira, F.C., Soler, J., Fonseca, A., Saraiva, I., Boaventura, R.A.R., Brillas, E., Vilar, V.J.P., Incorporation of electrochemical advanced oxidation processes in a multistage treatment system for sanitary landfill leachate, Water Research (2015), doi: 10.1016/ j.watres.2015.05.036. This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

ACCEPTED MANUSCRIPT

SANITARY LANDFILL LEACHATE

BIODEGRADABLE ORGANIC FRACTION DEGRADATION AMMONIUM OXIDATION (NITRIFICATION)

RECALCITRANT ORGANIC FRACTION DEGRADATION

HUMIC ACIDS REMOVAL SUSPENDED SOLIDS REMOVAL

BIODEGRADABILITY ENHANCEMENT

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ALKALINITY REMOVAL

2nd BIOLOGICAL PROCESS

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EAOP

COAGULATION

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1st BIOLOGICAL PROCESS

EAOP BIODEGRADABLE ORGANIC FRACTION DEGRADATION

NITRATES REMOVAL (DENITRIFICATION)

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Incorporation of electrochemical advanced oxidation processes in a

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multistage treatment system for sanitary landfill leachate

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Francisca C. Moreira a,*, J. Soler a, Amélia Fonseca b, Isabel Saraiva b, Rui A.R. Boaventura a, Enric

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Brillas c, Vítor J.P. Vilar a,*

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a

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Departamento de Engenharia Química, Faculdade de Engenharia, Universidade do Porto, Rua Dr.

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Roberto Frias, 4200-465 Porto (Portugal)

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b

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Guardeiras, Apartado 3003, 4471-907 Moreira da Maia (Portugal)

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LSRE - Laboratory of Separation and Reaction Engineering - Associate Laboratory LSRE/LCM,

Efacec Engenharia e Sistemas, S.A. (Unidade de Negócio Ambiente), Rua Eng. Frederico Ulrich -

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c

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Facultat de Química, Universitat de Barcelona, Martí i Franquès 1-11, 08028 Barcelona (Spain)

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*

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Tel.: +351 918257824; Fax: +351 225081674; E-mail address: [email protected] (Vítor J.P. Vilar)

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Tel.: +351 914332022; Fax: +351 225081674; E-mail address: [email protected]

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(Francisca C. Moreira)

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Laboratori d’Electroquímica dels Materials i del Medi Ambient, Departament de Química Física,

Corresponding authors:

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ACCEPTED MANUSCRIPT Abstract

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The current study has proved the technical feasibility of including electrochemical advanced

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oxidation processes (EAOPs) in a multistage strategy for the remediation of a sanitary landfill

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leachate that embraced: (i) first biological treatment to remove the biodegradable organic fraction,

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oxidize ammonium and reduce alkalinity, (ii) coagulation of the bio-treated leachate to precipitate

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humic acids and particles, followed by separation of the clarified effluent, and (iii) oxidation of the

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resulting effluent by an EAOP to degrade the recalcitrant organic matter and increase its

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biodegradability so that a second biological process for removal of biodegradable organics and

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nitrogen content could be applied. The influence of current density on an UVA photoelectro-Fenton

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(PEF) process was firstly assessed. The oxidation ability of various EAOPs such as electro-Fenton

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(EF) with two distinct initial total dissolved iron concentrations ([TDI]0), PEF and solar PEF

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(SPEF) was further evaluated and these processes were compared with their analogous chemical

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ones. A detailed assessment of the two first treatment stages was made and the biodegradability

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enhancement during the SPEF process was determined by a Zahn-Wellens test to define the ideal

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organics oxidation state to stop the EAOP and apply the second biological treatment. The best

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current density was 200 mA cm-2 for a PEF process using a BDD anode, [TDI]0 of 60 mg L-1, pH

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2.8 and 20 ºC. The relative oxidation ability of EAOPs increased in the order EF with 12 mg [TDI]0

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L-1 < EF with 60 mg [TDI]0 L-1 < PEF with 60 mg [TDI]0 L-1 ≤ SPEF with 60 mg [TDI]0 L-1, using

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the abovementioned conditions. While EF process was much superior to the Fenton one, the

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superiority of PEF over photo-Fenton was less evident and SPEF attained similar degradation to

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solar photo-Fenton. To provide a final dissolved organic carbon (DOC) of 163 mg L-1 to fulfill the

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discharge limits into the environment after a second biological process, 6.2 kJ L-1 UV energy and 36

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kWh m-3 electrical energy were consumed using SPEF with a BDD anode at 200 mA cm-2, 60 mg

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[TDI]0 L-1, pH 2.8 and 20 ºC.

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ACCEPTED MANUSCRIPT Keywords: Sanitary landfill leachate; EAOPs; Biological treatment; Coagulation; Biodegradability

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enhancement

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1. Introduction Municipal solid waste (MSW) is predominantly disposed of into sanitary landfills, where the

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percolation of rainfall in combination with the decomposition of the solid wastes by simultaneous

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and interrelated biological, chemical and physical changes leads to the generation of a highly

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contaminated liquid called “leachate” (Tchobanoglous and Kreith, 2002). The landfill leachate can

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reach the adjacent surface and groundwater, thus causing potentially serious hazards on the

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surrounding environment and public health (Fatta et al., 1999). The characteristics of the landfill

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leachate depend on diverse factors such as the amount, composition and moisture of the MSW, age

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of the landfill, hydrogeology and climate of the site and seasonal weather variations (Qasim and

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Chiang, 1994). Four main groups of pollutants are present in this effluent: (i) dissolved organic

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matter, including volatile fatty acids and more recalcitrant compounds such as humic and fulvic

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acids, (ii) inorganic ions like ammonium, chloride, potassium and sodium, (iii) heavy metals like

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cadmium, chromium, copper, lead, nickel and zinc and (iv) xenobiotic organic compounds

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originated from household or industrial chemicals (aromatic hydrocarbons, phenols, chlorinated

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aliphatics, pesticides and plastizers), available in low contents below 1 mg L-1 (Kjeldsen et al.,

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2002).

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In recent years, several technologies have been applied to landfill leachate remediation like: (i)

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biological treatments (aerobic and anaerobic) (Wiszniowski et al., 2006; Renou et al., 2008; Abbas

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et al., 2009), (ii) physical/chemical methods (flotation, coagulation/flocculation, activated carbon

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adsorption, air stripping, ion exchange and membrane separation processes such as microfiltration,

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nanofiltration, ultrafiltration and reverse osmosis) (Kurniawan et al., 2006; Wiszniowski et al.,

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2006; Renou et al., 2008; Abbas et al., 2009), (iii) advanced oxidation processes (AOPs) such as

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ozone and ozone based processes (O3, O3/UV, O3/UV/H2O2), TiO2/UV photocatalysis, Fenton

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(Fe2+/H2O2), photo-Fenton (PF) (Fe2+/H2O2/UV) and solar photo-Fenton (SPF) (Fe2+/H2O2/UV-Vis)

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(Wiszniowski et al., 2006; Renou et al., 2008; Abbas et al., 2009) and (iv) electrochemical AOPs

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ACCEPTED MANUSCRIPT (EAOPs) (Deng and Englehardt, 2007). Among EAOPs, the focus of the scientific community has

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been anodic oxidation (AO) using anode materials like boron-doped diamond (BDD) (Zhao et al.,

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2010; Anglada et al., 2011), PbO2 (Lei et al., 2007; Fernandes et al., 2014), RuO2 (Quan et al.,

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2013; Xiao et al., 2013) and carbon-based electrodes (Nageswara Rao et al., 2009; Zhang et al.,

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2010), which embraces indirect oxidation of organics by hydroxyl radicals (•OH) formed at the

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anode surface via O2 evolution from water oxidation by Eq. (1), denoted as M(•OH), and by

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electrogenerated active chlorine species (HClO, ClO− and Cl2).

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M + H2 O → M(• OH) + H+ + e

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(1)

To the best of our knowledge, the degradation of sanitary landfill leachates by the potent

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EAOPs based on H2O2 electrogeneration defined by Brillas et al. (2009) such as electro-Fenton

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(EF), UVA photoelectro-Fenton (PEF) and solar PEF (SPEF) has not been reported yet. In all these

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EAOPs and when a one-compartment cell is used, H2O2 is directly electrogenerated at the cathode

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from the two-electron reduction of injected O2 via Eq. (2) simultaneously with the generation of

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M(•OH) adsorbed at the anode surface by Eq. (1) (Brillas et al., 2009; Sirés et al., 2014).

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O2 (g) + 2 H+ + 2 e

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→ H2 O2

(2)

BDD anodes are widely employed since they produce very high amounts of BDD(•OH) from

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Eq. (1) (Panizza and Cerisola, 2005). The Fe2+ is externally added and reacts with electrogenerated

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H2O2, yielding Fe3+ along with •OH in the bulk from Fenton’s reaction Eq. (3) to oxidize organics

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(Brillas et al., 2009). Moreover, the cathodic reduction of Fe3+ occurs according to Eq. (4).

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Fe2+ + H2 O2 → Fe3+ + • OH + OH

(3)

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Fe3+ + e → Fe2+

(4)

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The PEF and SPEF methods can count on the irradiation of the contaminated solution by

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artificial light and natural sunlight, respectively, which improves the pollutants degradation by

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virtue of: (i) the photoreduction of photoactive Fe(III)-hydroxy complexes like FeOH2+, producing

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more Fe2+ and •OH according to Eq. (5) (Sun and Pignatello, 1993), and (ii) the direct photolysis of

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ACCEPTED MANUSCRIPT Fe3+ complexes with some organic intermediates, especially those containing the carboxylate group

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as shown in the general Eq. (6), which also promotes the Fe3+ ion regeneration (Horváth and

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Stevenson, 1992; Zuo and Hoigne, 1992; Faust and Zepp, 1993).

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Fe(OH)2+ + hv → Fe2+ + • OH

(5)

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Fe(OOCR)2+ + hv → Fe2+ + CO2 + R•

(6)

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Due to the complexity and recalcitrant nature of the landfill leachate matrix and in order to

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reduce the total cost of the treatment, strategies based on integrated biological-physical-chemical

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techniques have been under focus (Guo et al., 2010; Anfruns et al., 2013; Silva et al., 2013b; Silva

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et al., 2013c; De Torres-Socías et al., 2015). Recently, Vilar and co-workers in cooperation with

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Efacec Engenharia e Sistemas, S.A. company have published an European Patent (EP 2 784 031)

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(Saraiva et al., 2014) concerning an innovative integrated treatment strategy for raw sanitary landfill

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leachate including the following treatment steps: (i) initial biological process for removal of

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biodegradable organic compounds, nitrification/denitrification reactions and alkalinity reduction,

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(ii) coagulation with subsequent separation of the formed sludge for the removal of humic acids,

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suspended solids and other species filtering the radiation, (iii) PF and/or SPF processes for

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recalcitrant compounds degradation and biodegradability enhancement and (iv) final biological

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polishing step to reduce organic matter and nitrogen to values in agreement with the discharge

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limits into the environment. One of the main drawbacks of the PF/SPF processes is related with the

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high H2O2 consumption, turning these technologies too costly and originating H2O2 delivery

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constraints in isolated areas (Brillas et al., 2009). In this context, the alternative use of quite similar

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processes based on H2O2 electrogeneration can be a very interesting solution.

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The present paper is focused on the integration of EAOPs with H2O2 electrogeneration in the

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third step of the multistage strategy based on the aforementioned patented technology for the

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remediation of raw sanitary landfill leachates. The influence of current density (j) on the PEF

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process was firstly assessed and at the best attained conditions the comparative efficiency of EF

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ACCEPTED MANUSCRIPT with two distinct initial total dissolved iron concentrations ([TDI]0), PEF and SPEF treatments was

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evaluated. Further comparison of EAOPs with their analogous chemical processes was examined.

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Moreover, the biological oxidation and coagulation treatment steps preceding the EAOP were

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assessed in detail and the biodegradability enhancement along the SPEF process was determined by

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means of a Zahn-Wellens test to define the ideal organics oxidation state for further application of

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the second biological treatment.

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2. Experimental

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2.1. Chemicals

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Anhydrous sodium carbonate, used to provide alkalinity along the biological treatment, was of

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analytical grade purchased from Merck. Iron(III) chloride 40% (w/v), applied as coagulant, was of

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commercial grade from Quimitécnica.com. Concentrated sulfuric acid, hydrochloric acid and

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sodium hydroxide, used for pH adjustment, were of analytical grade supplied by Pronalab, Sigma-

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Aldrich and Merck, respectively. Iron(II) sulfate heptahydrate, used as catalyst, was of analytical

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grade purchased from Panreac. Hydrogen peroxide 30% (w/v), used as oxidant in AOPs, was of

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analytical grade purchased from Fisher Scientific. All the other chemicals were either of HPLC

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grade or analytical grade supplied by VWR-Prolabo, Sigma-Aldrich, Panreac, Merck, Fisher

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Scientific and Pronalab. Ultrapure and pure water used for analyses were obtained by a Millipore®

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Direct-Q system (18.2 MΩ cm resistivity at 25 ºC) and a reverse osmosis system (Panice),

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respectively.

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2.2. Sanitary landfill leachate

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The leachate was collected in March and April 2014 from a MSW sanitary landfill located

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nearby Porto, Portugal, in operation since 1999 and enlarged in 2010. Before collection, the landfill

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leachate undergone a pre-treatment in an aerated lagoon situated onsite through an activated sludge

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biological process under both aerobic and anoxic conditions, promoting partial elimination of 5-day

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biochemical oxygen demand (BOD5), chemical oxygen demand (COD) and dissolved organic

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carbon (DOC) together with partial ammonium oxidation and nitrogen removal. For simplicity

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purposes, this effluent will be nominated as raw landfill leachate along the current paper. Table 1

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displays its main physicochemical characteristics.

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2.3. Biological process The biological system was composed of a thermostatically controlled 12 L capacity reactor

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vessel with a conical bottom. This reactor was equipped with a mechanical stirrer (CAT Scientific,

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model R50D) and a 4000 L h-1 air pump (Aqua Medic, model Mistral 4000) connected to three

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small air stone diffusers located at the bottom. Raw landfill leachate (8 L) and previously

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centrifuged activated sludge from the biological reactor of the MSW sanitary landfill located nearby

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Porto (1.5 L) were added to the system. The pH was maintained between 6.5 and 9.0 through the

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addition of Na2CO3, which also provided alkalinity to be consumed in the nitrification process.

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Dissolved oxygen was maintained between 2 and 4 mg O2 L−1 and temperature at 27 ºC. Samples

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were taken throughout the biological treatment to assess total dissolved carbon (TDC), dissolved

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inorganic carbon (DIC), DOC and inorganic ions. The biological oxidation was stopped when a

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residual content of ammonium was reached. After 3 h of sedimentation, the supernatant was

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carefully transferred to another container. Various biological oxidation batches were performed to

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obtain enough volume for all the trials.

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2.4. Coagulation/aeration process

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Mixtures of ca. 30 L of the previous bio-treated landfill leachate were subjected to coagulation.

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To do this, FeCl3 was added to the landfill leachate free of sludge as described by Saraiva et al.

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(2014), followed by acidification to pH 3.3 (using H2SO4 up to a concentration of around 2.0 g

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SO42- L-1 and thereon HCl) and mechanical stirring for 15 min at 100 rpm. After sedimentation for

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48 h, the clarified effluent was carefully transferred to a container where it was aerated with a 4000

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L h-1 air pump (Aqua Medic, model Mistral 4000) for 3 h. After 24 h of sedimentation, the

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supernatant, i.e. the final pre-treated landfill leachate, was moved to another container to be used in

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EAOPs. Note that long sedimentation times were implemented to ensure complete settling and

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shorter times of 12 h or less can be used with the same efficiency (Saraiva et al., 2014).

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2.5. EAOPs system at laboratory scale The lab-scale flow plant was of 2.2 L of capacity and its main components were: (i) a

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thermostatically controlled 1.5 L cylindrical glass vessel under vigorous magnetic stirring, (ii) a

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photoreactor composed of a borosilicate tube allocated in the focus of two stainless steel reflectors,

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i.e. double compound parabolic collector (CPC), with an irradiated volume of 694 mL, and (iii) an

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electrochemical filter-press MicroFlowCell reactor from ElectroCell (Tarm, Denmark) with a 10

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cm2 BDD anode and a 10 cm2 carbon-PTFE air-diffusion cathode. A detailed description of these

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components has been reported by Moreira et al. (2014). The EAOPs experimental procedure for this

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system is given in Supplementary Material. In PEF experiments, the irradiation was supplied by a

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Philips TL 6W/08 fluorescent blacklight blue lamp, which emits UVA light in the wavelength

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region between 350 and 410 nm with λmax at 360 nm, allocated in the middle of the borosilicate

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tube and protected by a concentric inner quartz tube. In SPEF trials, the photoreactor was tilted 41º

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(local latitude), the top reflector was removed and the solar radiation was measured by a global UV

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radiometer (Kipp & Zonen B.V., model CUV5) placed at the same angle. The accumulated UV

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energy (QUV, in kJ L-1) for both configuration was calculated according to the procedure described

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in Supplementary Material.

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2.6. EAOPs system at pilot scale

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The pilot flow plant, with 35 L of capacity, was installed at the roof of the Chemical

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Engineering Department, Faculty of Engineering, University of Porto, Portugal (latitude: 41º 10’

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41.33’’; longitude: 8º 35’ 49.28’’). A sketch of this novel plant is displayed in Fig. 1a. It had two

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distinct configurations, either using the electrochemical components to apply an EAOP or

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bypassing the electrochemical devices to execute an AOP. The main constituents of the EAOPs

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system were: (i) a 20 L recirculation conical tank, (ii) a 5 L thermostated acrylic cylindrical reactor

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ACCEPTED MANUSCRIPT equipped with an outer jacket and an internal glass coil, both connected to a temperature controller

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(Surcis S.L., model TCH5), (iii) a structure composed of CPCs and (iv) an electrochemical cell,

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schematized in Fig. 1b.

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A comprehensive description of the characteristics of these components and the EAOPs

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experimental procedure for this system are given in Supplementary Material.

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2.7. Analytical determinations

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TDC and DIC were measured in a TOC-VCSN analyzer (Shimadzu). DOC was given by the

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difference between both data. Total dissolved nitrogen was determined in the same analyzer

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coupled with a TNM-1 unit. The specific energy consumption per unit DOC mass (ECDOC, in kWh

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(kg DOC)-1) and per unit volume (EC, in kWh m-3) were calculated via Eqs. (7) and (8),

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respectively (Flox et al., 2007):

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ECDOC =

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EC =

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where 1000 is a conversion factor (in mg g-1), Ecell is the average applied cell voltage (in V), I is the

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applied current (in A), t is the electrolysis time (in h), Vs is the solution volume (in L) and

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∆(DOC)exp is the experimental DOC concentration decay (in mg L-1).

1000 Ecell I t Vs ∆(DOC)exp

Ecell I t

(7) (8)

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The solution pH and temperature were measured by a WTW inoLab 730 laboratory meter.

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Conductivity and redox potential were determined by a HANNA Instruments HI 9828

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Multiparameter analyzer. Alkalinity, turbidity, COD, BOD5, total suspended solids (TSS), volatile

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suspended solids (VSS), total nitrogen, total phosphorous, sulfite, sulfide and free chlorine were

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measured according to the Standard Methods for the Examination of Water and Wastewater

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(Clesceri et al., 2005). H2O2 concentration was determined by the colorimetric metavanadate

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method (Nogueira et al., 2005). TDI was obtained from the colorimetric 1,10-phenantroline

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standardized procedure (ISO6332:1998, 1998). The specific ultraviolet absorbance at 254 nm

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(SUVA254, in L mg-1 m-1) was obtained by dividing the UV absorbance at 254 nm (in m-1) of

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samples by their DOC concentration (in mg L-1). UV-Vis measurements were carried out using a

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VWR UV-6300PC or a Merck Spectroquant® Pharo 100 spectrophotometer. The biodegradability of samples was determined by a 28-day Zahn-Wellens test from the Test

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Guideline no. 302 B (OECD, 1992). A mixture composed of (i) 240 mL of sample at neutral pH,

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(ii) activated sludge from an urban WWTP of Northern Portugal previously centrifuged and (iii)

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mineral nutrients (KH2PO4, K2HPO4, Na2HPO4, NH4Cl, CaCl2, MgSO4 and FeCl3) was added to an

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open glass vessel magnetically stirred and kept in the dark at 25 ºC. Control and blank experiments

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were prepared using the highly biodegradable glucose and pure water, respectively, instead of

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sample. The percentage of biodegradation (Dt) was calculated through Eq. (9) (OECD, 1992):

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Dt = 1 − C T A

CBA

×100

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(9)

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where CT and CB are the sample and blank DOC concentrations (in mg L-1) determined at the

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sampling time t, respectively, and CA and CBA are the corresponding sample and blank DOC

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concentrations measured 3 h after beginning the test.

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The analytical procedures for inorganic ions and low-molecular-weight carboxylic acids

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(LMCA) determination are reported in Supplementary Material. Before TDC, DIC, DOC, total dissolved nitrogen, TDI, free chlorine, UV absorbance at 254,

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inorganic ions and LMCA analysis, the samples were filtered by 0.45 µm Nylon filters from

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Whatman. Before biological tests, the excess of H2O2 was removed using a small volume of 0.1 g L-

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3. Results and discussion

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3.1. Characterization of the raw landfill leachate

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catalase solution (2500 U mg-1 bovine liver) after adjusting the sample pH to 6.5-7.5.

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As can be seen in Table 1, the raw landfill leachate displayed the following main

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characteristics: (i) very dark brown color that can be associated to the presence of humic acids

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(Nieder and Benbi, 2008), (ii) very strong odor, (iii) slight alkaline pH, (iv) high DOC and COD

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contents with a low biodegradable organic fraction (BOD5/COD ratio of 0.04-0.1), (v) high nitrogen

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ACCEPTED MANUSCRIPT content mainly due to N-NH4+ species (66-74%), (vi) high alkalinity, (vii) high conductivity of

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21.0-23.3 mS cm-1 corresponding to a high calculated ionic strength of 0.206 M, (viii) absence of

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phosphate, (ix) low total phosphorous concentration and (x) very low [TDI] of 2.2-3.0 mg L-1.

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While N-NH4+, total nitrogen, alkalinity, conductivity and total phosphorous values are typically

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found for landfill leachates from sites within 3-6 years, BOD5 and COD values are commonly

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attributed to older landfill leachates (Kjeldsen et al., 2002; Tchobanoglous and Kreith, 2002; Ziyang

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et al., 2009), but in the present study they may be related to the partial removal of biodegradable

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organic compounds in the aerated lagoon pre-treatment. In fact, the landfill leachate before

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biological lagooning from the same MSW sanitary landfill was previously characterized by greater

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BOD5 of 1325 mg O2 L-1 and COD of 7426 mg O2 L-1 (Silva et al., 2013c).

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3.2. Biological process

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The main goals of the aerobic biological process were to remove biodegradable organic matter,

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oxidize ammonium ions and consume alkalinity. The biological system was under well-operation as

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pointed out by: (i) an average sludge volume index (SVI) for the various batches of 55 mL g-1,

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indicative of excellent sludge settling and compaction characteristics, typically between 50 and 100

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mL g-1 (Cheremisinoff, 1994) and (ii) an average food to microorganism (F/M) ratio based on

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BOD5 and VSS of 0.035 g substrate per g biomass per day, which corresponds to an extended

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aeration process (Spellman, 2004). The totality of data on the characteristics of the landfill leachate

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throughout the various biological treatment batches used in the further discussion is not displayed.

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Nevertheless, Fig. 2 exemplifies one biological treatment batch in terms of DOC and nitrogen

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content and Table 1 shows the characteristics of the landfill leachate at the end of the biological

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treatment.

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The biodegradable organic carbon fraction was almost totally removed after 42-47 h of

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biological treatment, reaching 90-95% of BOD5, 13-33% of DOC and 9-31% of COD abatements.

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The low DOC and COD removals suggest that the major fraction of the raw landfill leachate,

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bioresistant compounds, probably humic substances (Silva et al., 2013c). In general, the treatment

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of landfill leachates by applying activated sludge reactors attains higher COD removals from 29%

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to 69% (Baumgarten and Seyfried, 1996; Bae et al., 1999; Silva et al., 2013c). The lower organic

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removal here achieved can be attributed to the partial removal of biodegradable organic compounds

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in the upstream biological lagooning, since the applied landfill leachate showed a lower BOD5

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content of 180-300 mg O2 L-1 in contrast with BOD5 contents greater than 1000 mg O2 L-1 for the

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cited studies. In all cases, the biological process was inefficient for the removal of the refractory

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compounds.

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Typically, the biological nitrification process comprises the oxidation of ammonium to nitrite

M AN U

276

RI PT

267

277

by Eq. (10) and the subsequent oxidation of nitrite to nitrate via Eq. (11) (Metcalf & Eddy, 2004).

278

NH+4 + O2 → NO2 + 2H+ + H2 O

279

NO + 2 O2 → NO3

3

-

2

-

(11)

TE D

1

(10)

Gerardi (2002) and Metcalf & Eddy (2004) have stated the following considerations for

281

nitrification: (i) optimal temperature range from 28 to 32 ºC, (ii) optimal dissolved oxygen around

282

3.0 mg L-1, (iii) optimal pH from 6.7 to 8.0 and (iv) alkalinity used as a carbon source by nitrifying

283

bacteria in a ratio of 7.14 mg CaCO3 per mg N-NH4+ oxidized. Consequently, the conditions

284

employed in the current activated sludge reactor, i.e. temperature of 27 ºC, dissolved oxygen of 2-4

285

mg L-1, pH of 6.5-9.0 and permanent availability of alkalinity, might maximize the nitrification

286

process.

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287

The ammonium ion was almost completely oxidized after 42-47 h of biological treatment, with

288

an average specific nitrification rate of 13.6 mg N-NH4+ per h per g VSS, and afterwards it was

289

much more slowly converted because low amounts of Na2CO3 were provided for a minimal

290

alkalinity at the end of the treatment. This nitrification rate was very acceptable considering values

291

reported in literature of 4.9-12.6 mg N per h per g VSS at 20 ºC for a landfill leachate with 1199 mg

13

ACCEPTED MANUSCRIPT N-NH4+ L-1 (Spagni and Marsili-Libelli, 2009), 20.4 mg N-NH4+ per h per g VSS at 24-27 ºC for a

293

landfill leachate with 1452 mg N-NH4+ L-1 (Yusof et al., 2010) and 8.2 mg N-NH4+ per h per g VSS

294

at 26.9 ºC for a landfill leachate with 3864 mg N-NH4+ L-1 (Silva et al., 2013c). After 138-164 h,

295

ammonium removals of 98-100% were attained and high amounts of nitrite (1515-1862 mg N-NO2-

296

L-1) were formed but with almost null oxidation to nitrate (23-32 mg N-NO3- L-1). The negligible

297

nitrite oxidation might be due to the accumulation of free ammonia and free nitrous acid, which are

298

known inhibitors of nitrite-oxidizing bacteria (Villaverde et al., 1997; Canziani et al., 2006; Park et

299

al., 2015). The overall nitrification process consumed in average 9.3 g CaCO3 per L of raw leachate

300

or 8.5 mg CaCO3 per mg N-NH4+, which was only slightly superior to the stoichiometric ratio (7.14

301

mg CaCO3 per mg N-NH4+). At the end of the treatment, alkalinity residual values of 419-772 mg

302

CaCO3 L-1 and very low DIC values of 11-20 mg L-1 were achieved. The total nitrogen was

303

maintained unchanged, indicating null nitrogen gas stripping. Moreover, the biological treatment

304

led to (i) substantial odor minimization, (ii) 70-72% of sulfite oxidation and (iii) addition of 737-

305

1409 mg Na+ L-1 by Na2CO3 used to provide alkalinity.

306

3.3. Coagulation/aeration process

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The purpose of the coagulation process was to remove humic acids, suspended solids and

308

colloidal particles that are able to act as photons absorbers in the subsequent photo-assisted

309

electrochemical treatments and also to provoke the electrochemical cell clogging. In a first

310

approach, the bio-treated landfill leachate was subjected to coagulation using FeCl3 at pH 4.2,

311

according to Saraiva et al. (2014). However, the subsequent EAOP step performed at pH 2.8

312

revealed very fast pH decrease up to pH 2.0 and total oxidation of nitrite to nitrate before reaction

313

start under effluent recirculation in the EAOPs system, together with very low H2O2 accumulation

314

during reaction. The pH drop can be attributed to the oxidation of nitrous acid by oxygen via Eq.

315

(12) (Mudgal et al., 2007) since the use of the carbon-PTFE air-diffusion cathode introduces high

316

amounts of air in the EAOPs system. Eq. (12) occurs preferentially at pH 1.0-2.0, whereupon it is

AC C

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14

ACCEPTED MANUSCRIPT decelerated up to attain a very low rate at pH 4.5, where the equilibrium speciation dictates a higher

318

amount of nitrite ion to that of nitrous acid (Braida and Ong, 2000; Mudgal et al., 2007). On the

319

other hand, the oxidation of nitrite to nitrate can occur in the presence of H2O2 by the formation of

320

peroxynitrite (ONOO-) in acid medium via Eqs. (13) and (14) (Goldstein et al., 1996; Edwards and

321

Plumb, 2007), explaining the low H2O2 accumulation.

322

2HNO2 + O2 → 2HNO3

323

H2 O2 + HNO2 → ONOOH + H2 O

324

ONOOH → NO3 + H+

RI PT

317

(12)

(13)

SC

-

(14)

To avoid pH decay and extra consumption of H2O2 along the EAOP step, all nitrite ions were

326

previously oxidized to nitrate by performing the coagulation at pH 3.3 and subjecting the clarified

327

effluent to aeration for 3 h. This pH value was chosen since preliminary tests revealed an almost

328

null nitrite oxidation by aeration at higher pH values and, in addition, a final pH value of 2.2-2.9

329

was reached, allowing the EAOP to be applied with null/almost null pH correction. Note that an

330

alternative strategy for nitrite oxidation by aeration could be to perform coagulation at pH 4.2 and

331

afterwards decrease the pH to 3.3 to proceed with aeration. However, a previous study on the best

332

pH for coagulation has determined a slight enhancement on coagulation efficiency for pH values

333

below 4.2 (data not shown). Traditional PF processes applied to a landfill leachate with and without

334

nitrite revealed a reduction of 58% on the H2O2 consumption for the free-nitrite effluent.

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335

The coagulation of the bio-treated landfill leachate with subsequent clarification achieved: (i)

336

high organics removal of 63-65% for DOC and 44-51% for COD, (ii) TSS and VSS reductions of

337

57-67% and 40-54%, respectively, (iii) color changing from very dark brown to moderate yellowish

338

brown and (iv) SUVA254 decrease of 27-40% (see Fig. 2 and Table 1). The color changing can be

339

associated to the presence of fulvic acids and absence of humic acids (Nieder and Benbi, 2008) and

340

the SUVA254 decay can be strongly correlated to the precipitation of humic substances (Letterman et

341

al., 1999). Saraiva et al. (2014) have reported similar organic fraction removal and color

15

ACCEPTED MANUSCRIPT minimization and Silva et al. (2013c) have determined an abatement on humic substances of 37%

343

when a bio-treated landfill leachate was acidified from pH 8.4 to 3.0. Note that the acidification of

344

the current bio-treated effluent to pH around 3 only led to 32% DOC abatement along with color

345

changing to a lighter brown, suggesting the inability of this procedure to precipitate humic acids

346

with the same efficiency as the coagulation step. In addition, nitrogen compounds corresponding to

347

22-30% of total nitrogen and 26-37% of total dissolved nitrogen were precipitated during this step.

348

The precipitated dissolved nitrogen was composed of 19-23% of nitrite and 7-11% of organic

349

nitrogen compounds. Apart from nitrite precipitation (23-33% of total nitrite), some nitrite ions

350

were oxidized to nitrate ions in an extent of 31-33%.

SC

RI PT

342

The subsequent aeration of the coagulated landfill leachate led to total oxidation of nitrite to

352

nitrate, along with a pH decrease to 2.2-2.9 (see Fig. 2 and Table 1). Besides that, the total dissolved

353

organic nitrogen was reduced by 72-80% due to an extra clarification of the effluent.

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Moreover, the overall coagulation/aeration step induced to: (i) total/almost total inorganic

355

carbon fraction removal as well as alkalinity, (ii) almost total phosphorous removal likely due to the

356

precipitation of phosphorous compounds, (iii) the introduction of high amounts of sulfate ions from

357

44-83 to 1749-1917 mg L-1, (iv) an increase in chloride content from 2211-2906 to 3046-3822 mg

358

L-1 and (vi) the presence of a [TDI] of 11-21 mg L-1 from the addition of FeCl3 that can be applied

359

as catalyst in the subsequent EAOP step based on Fenton’s reaction chemistry. A volume of ca. 280

360

mL of sludge per L of effluent that require proper treatment for further storage in a specific site was

361

also produced.

362

3.3. EAOPs degradation

363

3.3.1. General

AC C

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354

364

The solution conductivity is a key parameter for the treatment of wastewaters by EAOPs as

365

long as high conductivity values allow to minimize the power consumption. The landfill leachate

366

showed a very high conductivity of 18.0-23.3 mS cm-1 along all treatment steps (see Table 1), being

16

ACCEPTED MANUSCRIPT 367

suitable for EAOPs application. This conductivity was even higher than that of electrolytes

368

commonly applied to treat synthetic solutions by EAOPs, e.g. the conductivity of 7.0 g Na2SO4 L-1

369

is 8.6 mS cm-1 (Moreira et al., 2014). Landfill leachate remediation by the classical PF process has been made by applying iron

371

contents in a wide range of 10-2000 mg L-1, with highlighting on 60-80 mg L-1 (de Morais and

372

Zamora, 2005; Primo et al., 2008; Silva et al., 2013a; Silva et al., 2013c; Saraiva et al., 2014).

373

Preliminary trials on [TDI]0 from 20 to 80 mg L-1 for the remediation of the current pre-treated

374

landfill leachate by PEF with a BDD anode operating at 200 mA cm-2, pH 2.8 and 20 ºC revealed a

375

best [TDI]0 of 60 mg L-1 (data not shown).

SC

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Fig. 3 shows the theoretical Fe3+ speciation diagram calculated by the chemical equilibrium

377

modelling system MINEQL+ (Schecher and McAvoy, 2007) based on landfill leachate ions average

378

concentration before the EAOP step for a total Fe3+ concentration of 60 mg L-1. From this diagram,

379

one can infer that the highest proportion of the most photoactive ferric iron-water complex, i.e.

380

FeOH2+ species, occurs at pH 2.8 (11% of Fe3+ molar fraction), together with no iron precipitation

381

as Fe(OH)3 (s). In fact, this pH value is often assumed as optimal for Fenton’s reaction because in

382

many solutions the dominant iron species in solution is FeOH2+ and iron precipitation does not take

383

place yet (Pignatello, 1992). Moreover, extra theoretical calculations (data not shown) advised both

384

negative and positive effects from the large sulfate content of the pre-treated landfill leachate

385

compared to the low sulfate content of the raw landfill leachate. Negative effects: lower FeOH2+

386

concentration (7.1 against 10 mg L-1, respectively) and higher HSO4- amount (72 against 3.3 mg L-1,

387

respectively). Positive effect: iron precipitation prevention (null iron precipitation against 33% of

388

iron precipitation as Fe(OH)3 (s), respectively). It is known that HSO4- scavenges •OH to form sulfate

389

radical (SO4•−) via Eq. (15) (Neta et al., 1988).

390

HSO4 + • OH → SO•4 + H2 O

AC C

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(15)

17

ACCEPTED MANUSCRIPT On the other hand, additional calculations on the chloride content (data not shown) revealed a

392

negligible influence of this species on FeOH2+ fraction and iron precipitation. Nevertheless, chloride

393

ion can be oxidized at the anode surface to form active chlorine species (HClO, ClO− and Cl2) that

394

contribute to organics removal (Chiang et al., 1995), although these species can react with organics

395

leading to harmful chlorinated organic by-products. While the high chloride concentration of the

396

landfill leachate favors the electrogeneration of active chlorine species, the H2O2 produced at the

397

cathode is a known dechlorination agent for free chlorine, avoiding their accumulation (Goldstein et

398

al., 2007). In fact, free chlorine was detected along reactions in low concentrations below 5.0 mg

399

Cl2 L-1.

400

3.3.2. Influence of current density on PEF process

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391

The pre-treated landfill leachate was subjected to a PEF process at [TDI]0 of 60 mg L-1, pH 2.8

402

and 20 ºC by applying j values from 25 to 300 mA cm-2. The high organic content of the effluent

403

justifies the high j used. Fig. 4a shows that the rise in j yielded higher DOC removal but this

404

increment was almost negligible from 200 to 300 mA cm-2. A pseudo-first-order kinetic model was

405

fitted to the experimental data according to the procedure described in Supplementary Material. The

406

corresponding kDOC values, displayed in Table 2, were 1.7 and 2.6/2.7 times higher for 100 and

407

200/300 mA cm-2, respectively, when compared to 25 mA cm-2. DOC dropped more sharply in the

408

first 30 min of reaction (2.1 kJ L-1 of accumulated radiation), in an extent of 14-33% for all j,

409

whereupon its decay turned slighter. In the same period, Fig. 4b shows small TDI decays near 9-

410

13%. These results suggest that in a first stage more easily degradable compounds were mineralized

411

and some complexes between Fe3+ and primary by-products were precipitated. Note that Fe3+ did

412

not precipitate in the initial matrix of the pre-treated landfill leachate, in agreement with Fig. 3

413

where the precipitation of Fe3+ as Fe(OH)3 (s) is not expected up to pH 2.9.

AC C

EP

TE D

401

414

The higher DOC decays for greater j values gave superior energy consumptions, as can be seen

415

in Fig. 4c. Despite the similar DOC decays at 200 and 300 mA cm-2, an average ECDOC value 1.7

18

ACCEPTED MANUSCRIPT times higher was reached for the higher j. Moreover, Fig. 4d shows quite similar and very high

417

H2O2 accumulations of 535-911 mg L-1 for times above 60 min for 200 and 300 mA cm-2,

418

suggesting the occurrence of parasitic reactions above 200 mA cm-2 that lead to the formation of

419

smaller relative amounts of BDD(•OH) and •OH in the bulk. These parasitic reactions can involve:

420

(i) the anodic oxidation of BDD(•OH) to O2 through Eq. (16), (ii) the dimerization of •OH to H2O2

421

via Eq. (17) and (iii) the destruction of •OH with H2O2 and Fe2+ as represented by Eqs. (18) and

422

(19), respectively (Sun and Pignatello, 1993; Marselli et al., 2003; Sirés et al., 2014). Hence, 200

423

mA cm-2 can be chosen as the best j value for the PEF treatment of the pre-treated landfill leachate

424

in the present study. Note that all tested j ensured the presence of H2O2 along the reaction time,

425

which in turn guaranteed the maximum production of •OH from Fenton’s reaction Eq. (3).

426

2 BDD(• OH) → 2 BDD + O2 + 2 H+ + 2 e

(16)

427

2 • OH → H2 O2

(17)

428

H2 O2 + • OH → HO•2 + H2 O

429

Fe2+ + • OH → Fe3+ + OH

430

3.3.3. Comparative application of EF, PEF, SPEF, Fenton, PF and SPF processes

TE D

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SC

RI PT

416

(18) (19)

Various EAOPs such as EF with effluent iron content, i.e. 12 mg [TDI]0 L-1, and with 60 mg

432

[TDI]0 L-1, PEF with 60 mg [TDI]0 L-1 and SPEF with 60 mg [TDI]0 L-1 were applied to the

433

degradation of the pre-treated landfill leachate using 200 mA cm-2, pH 2.8 and 20 ºC. The

434

comparative mineralization ability for these EAOPs is illustrated in Fig. 5a1 and raised in the

435

sequence EF with 12 mg [TDI]0 L-1 < EF with 60 mg [TDI]0 L-1 < PEF ≤ SPEF, presenting DOC

436

removals of 34%, 42%, 72% and 78%, respectively, after 300 min of reaction. Table 2 shows kDOC

437

values 1.4, 2.6 and 3.6 times higher for EF with 60 mg [TDI]0 L-1, PEF and SPEF, respectively, in

438

comparison with EF with 12 mg [TDI]0 L-1. The DOC abatements attained for PEF and SPEF were

439

very similar either in terms of time (Fig. 5a1) or accumulated UV energy (Fig. 5a1.1). These results

440

allow assuming that: (i) the BDD(•OH) formed via Eq. (1) had a large participation on the

AC C

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431

19

ACCEPTED MANUSCRIPT mineralization process, but they were only able to partially reduce the DOC, (ii) the additional •OH

442

formed from Fenton´s reaction Eq. (3) in the EF process with 60 mg [TDI]0 L-1 only slightly

443

improved the effluent mineralization, (iii) the radiation provided in PEF and SPEF processes had a

444

crucial role on the effluent mineralization (30-36% higher mineralization for PEF and SPEF

445

compared to EF after 300 min of reaction) because of the additional •OH production from Eq. (5)

446

and the possible direct photolysis of Fe(III)-carboxylate complexes via Eq. (6) and (iv) persistent

447

organic intermediates were generated. Since artificial and solar radiations can be indistinctly

448

applied in terms of mineralization ability, the EAOP treatment step might include just one type of

449

radiation or integrate both PEF and SPEF processes. When deciding on EAOPs system

450

configuration, one must consider a decrease in electricity consumption when solar radiation is used.

451

In addition, Fig. 5a2 reveals the availability of H2O2 during all EAOPs treatments and a decreasing

452

H2O2 accumulation in the order EF with 12 mg [TDI]0 L-1 > EF with 60 mg [TDI]0 L-1 > PEF >

453

SPEF, in accordance with the increasing oxidation ability of these processes.

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441

On the other hand, chemical Fenton, PF and SPF processes were applied to the degradation of

455

the pre-treated landfill leachate for comparison with their analogous EF, PEF and SPEF processes,

456

respectively. In all the AOPs, the H2O2 was supplied in multiple small additions between 200 and

457

400 mg L-1 since similar approaches have improved the oxidation rate, avoiding the absence of

458

H2O2 and minimizing its consumption (Bacardit et al., 2007). [TDI]0 of 60 mg L-1, pH 2.8 and 20 ºC

459

were used. The DOC decay profiles for AOPs are displayed in Fig. 5a1 and, as expected, their

460

mineralization ability raised in the sequence Fenton < PF < SFP. The faster DOC decay attained in

461

SPF when compared to PF was slightly less pronounced in terms of accumulated UV energy (see

462

Fig. 5a1.2) and came along with a slightly higher H2O2 consumption (see Fig. 5a2). The Fenton

463

process induced a DOC removal of only 19% after 30 min of reaction that remained almost constant

464

up to the end of the process. This removal was 1.8 times lower than that of EF (see Table 2) because

465

in the latter also occurred: (i) the BDD(•OH) generation from Eq. (1), (ii) the cathodic Fe3+

AC C

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TE D

454

20

ACCEPTED MANUSCRIPT regeneration via Eq. (4) and (iii) even the direct oxidation of organic compounds on BDD surface.

467

The superiority of PEF and SPEF over PF and SPF, respectively, was less pronounced than that of

468

EF over Fenton, although the DOC decay was much less evident in the first 30 min of PF reaction

469

compared to PEF. These achievements suggest that the presence of radiation, mostly sunlight,

470

reduces the oxidative role of electrochemically generated oxidants because of the efficient action of

471

photolytic reactions (5) and (6).

RI PT

466

The UVA photolysis with addition of 700 mg L-1 of H2O2 at the reaction beginning yielded a

473

very slow DOC removal, only reaching 11% abatement in 300 min (see Fig. 5a1). This indicates a

474

very small participation of UVA radiation, H2O2 and PF reactions in the presence of 12 mg [TDI]0

475

L-1 of the effluent matrix on the degradation of the landfill leachate.

M AN U

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472

Moreira et al. (2014) have pointed out that the availability of H2O2 from the first instant of

477

reaction in the AOPs in contrast with the gradual accumulation in EAOPs can increase some

478

organics degradation during the first times of reaction. In this context, a PEF process with initial

479

700 mg L-1 H2O2 addition was performed to be compared with the regular PEF process. However,

480

only a very slight increment of 4% on DOC removal was accomplished after the first 15 min of

481

reaction, followed by a very similar DOC decay profile (see Fig. 5a3) with a very alike kDOC value

482

(see Table 2). These achievements suggest that the high applied j of 200 mA cm-2 provides a

483

sufficient H2O2 accumulation from the beginning of the process to maximize the Fenton’s reaction

484

Eq. (3) (see Fig. 5b2). Furthermore, the H2O2 initial concentration can influence differently the

485

degradation of distinct compounds.

486

3.4. Biodegradability enhancement during SPEF process

AC C

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476

487

The assessment of the biodegradability enhancement during the EAOP treatment step was

488

performed by means of a Zahn-Wellens test applied to various samples taken along a SPEF process

489

in the pilot plant using a Pt anode with 200 mA cm-2, [TDI]0 of 60 mg L-1, pH 2.8 and 20 ºC. Note

490

that the pilot and lab-scale systems presented distinct experimental conditions that affect the

21

ACCEPTED MANUSCRIPT mineralization process, namely electrode area/solution volume ratios of 4.1 and 8.7 cm2 L-1,

492

respectively, and irradiated volume/total volume ratios of 20% and 60%, respectively. Fig. 6

493

exhibits the Dt evolution for various samples at different treatment stages during the Zahn-Wellens

494

test and the characteristics of each sample in terms of DOC, COD and BOD5 at day 0 and day 28 are

495

collected in Table 3. As long as organics were degraded, more biodegradable samples were found,

496

with exception of samples S8 and S9, suggesting the presence of more recalcitrant intermediates

497

when DOC was reduced up to ca. 85 mg L-1. The treatment stage corresponding to sample S6 with

498

DOC of 163 mg L-1 and Dt of 61% can be selected as the best endpoint for the SPEF treatment since

499

at this oxidation degree COD could be reduced up to 102 mg O2 L-1 by means of a biological

500

process, thereby comprising with Portuguese (Decree-Law no. 236/98) and European (Directive no.

501

91/271/CEE) discharge limits for WWTPs effluents, i.e. COD values of 150 and 125 mg O2 L-1,

502

respectively. As a result, the best SPEF process checked in the present study should be performed

503

up to reach an accumulated UV energy of 6.2 kJ L-1 (ca. 147 min under an average solar UV

504

radiation intensity of 46 WUV m-2), consuming 36 kWh m-3 of electrical energy. The pre-treated

505

landfill leachate oxidized up to 163 mg L-1 of DOC and subjected to subsequent neutralization to

506

pH 7.5 and sludge removal by clarification for 3 h achieved COD, BOD5, total nitrogen and nitrate

507

values above the Portuguese and European regulations for WWTPs release into the environment

508

(see Table 1). These two latter steps led to a concentration decay of TDI from 50 mg L-1 to below its

509

detection limit (0.13 mg L-1) and the production of 53 mL of sludge per L of effluent that require

510

further adequate treatment. The Zahn-Wellens biological process was able to reduce COD and

511

BOD5 to values in agreement with regulations, but a subsequent biological denitrification step must

512

be performed to convert nitrates into nitrogen gas and hence comply with nitrate and total nitrogen

513

legislation limits (see Table 1).

AC C

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491

514

Ion-exclusion HPLC analysis on LMCA only revealed the formation of formic acid (tr = 18.5

515

min) in the SPEF treatment at the pilot plant, with a maximal concentration of 14 mg C L-1 and a

22

ACCEPTED MANUSCRIPT 516

contribution to DOC always below 4.5%. No significant changes in the content of inorganic ions of

517

the leachate matrix were determined by ion chromatography.

518

4. Conclusions The application of EAOPs to a multistage treatment strategy composed of a first biological

520

process, coagulation/aeration, an EAOP and a second biological process for a raw sanitary landfill

521

leachate remediation proved to be a successful approach concerning the elimination of organic

522

matter and nitrogen compounds.

RI PT

519

The raw sanitary landfill leachate exhibited very dark brown color, high amounts of recalcitrant

524

organic compounds and ammonium and high alkalinity. The first biological treatment led to the

525

removal of the biodegradable organic fraction corresponding to 13-33% of DOC, total ammonium

526

oxidation predominantly to nitrite and almost total alkalinity removal. The subsequent coagulation

527

process brought a large decolorization due to the removal of humic acids, TSS and VSS and came

528

along with 63-65% DOC abatement. The presence of nitrite ion proved to have a negative effect on

529

the EAOP step as long as it caused an extra H2O2 consumption and a high pH decrease and then a

530

subsequent aeration stage at pH 3.3 was applied to totally convert nitrite into nitrate.

TE D

M AN U

SC

523

The use of a current density of 200 mA cm-2 was chosen as the best value for a PEF process

532

using [TDI]0 of 60 mg L-1, pH 2.8 and 20 ºC. The EAOPs relative oxidative capability increased in

533

the order EF with 12 mg [TDI]0 L-1 < EF with 60 mg [TDI]0 L-1 < PEF with 60 mg [TDI]0 L-1 ≤

534

SPEF with 60 mg [TDI]0 L-1 when a BDD anode, 200 mA cm-2, pH 2.8 and 20 ºC were applied. The

535

comparison between EAOPs and their analogous AOPs revealed a highly superiority of EF over

536

Fenton, less pronounced superiority of PEF over PF and similarity between SPEF and SPF in terms

537

of oxidation ability.

AC C

EP

531

538

The assessment of the biodegradability enhancement along the SPEF process revealed the need

539

to perform the electrochemical treatment up to reach around 163 mg DOC L-1 to couple a further

540

biological process that allows to achieve a final wastewater quality in agreement with Portuguese

23

ACCEPTED MANUSCRIPT and European regulations for the discharge of effluents from a WWTP in terms of organic content.

542

Consumptions of 6.2 kJ L-1 of accumulated UV energy and 36 kWh m-3 of electrical energy were

543

attained when using a BDD anode at current density of 200 mA cm-2, [TDI]0 of 60 mg L-1, pH 2.8

544

and 20 ºC. The subsequent biological treatment should also include a denitrification step to convert

545

nitrates into nitrogen gas and thus comply with the limits imposed by regulations.

546

Acknowledgements

547

Financial support was partially provided by (i) AdvancedLFT project (reference FCOMP-01-0202-

548

FEDER-033960), financed by FEDER (Fundo Europeu de Desenvolvimento Regional) under

549

COMPETE program (Programa Operacional Fatores de Competitividade) of QREN (Quadro de

550

Referência Estratégico Nacional) within the I&DT system (Sistema de Incentivos à Investigação e

551

Desenvolvimento Tecnológico), (ii) UID/EQU/50020/2013 project, co-financed by FCT/MEC

552

(Fundação para a Ciência e a Tecnologia/Ministério da Educação e Ciência) and FEDER under

553

Program PT2020 and (iii) NORTE-07-0162-FEDER-000050 project, co-financed by FEDER,

554

QREN and ON2 program (Programa Operacional Regional do Norte). F.C. Moreira acknowledges

555

her Ph.D. fellowship SFRH/BD/80361/2011 supported by FCT. V.J.P. Vilar acknowledges the FCT

556

Investigator 2013 Programme (IF/01501/2013).

557

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Zuo, Y., Hoigne, J., 1992. Formation of hydrogen peroxide and depletion of oxalic acid in atmospheric water by photolysis of iron(III)-oxalato complexes. Environ. Sci. Technol. 26(5), 10141022.

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ACCEPTED MANUSCRIPT Table 1. Physicochemical characterization of the landfill leachate along the various stages of the treatment: raw, after biological process, after coagulation, after aeration, after SPEF/neutralization/clarification and after 28 days of Zahn-Wellens test. The discharge limits for WWTPs final effluents according

RI PT

to Portuguese legislation (Decree-Law no. 236/98) and European Directive no. 91/271/CEE are also displayed.

Color (diluted 1:20)

d. e

d. e

d. e

n.d. f

n.d. f

Odor

Very strong

Weak

Very weak

Very weak

n.d. f

SC

Moderate yellowish brown

Very weak

M AN U

Very dark brown

d.

-

d. e

Very dark brown

d.

Very light yellow

Moderate yellowish brown

Color

Odor (diluted 1:20)

Very light yellow

After SPEF, neutralization and clarification c

Raw

e

ELV d for DecreeLaw no. 236/98 or Directive no. 91/271/CEE

After aeration b

Parameter (units)

e

After 28 days of ZahnWellens

After coagulation b

After biological process a

n.d.

f

n.d.

f

n.d.

f

n.d. f

n.d. f (diluted 1:20) or n.d. f (diluted 1:20) or 6.0-9.0 or 3 ºC increase g or -

AC C

EP

TE D

8.2-9.0 7.0-8.0 2.7-3.4 2.2-2.9 6.2 7.5 pH 20 20 20 20 20 20 Temperature (ºC) 21.0-23.3 18.8-22.0 18.0-21.0 18.9-20.3 19.6 Conductivity (mS cm-1) h h h -1 7623-10593 419-772