Prof Silvia Fiore

Prof Marie-Odile Simonnot

Andreina Laera

Anna Gienlik

Ramita Khanongnuch

Suchanya Wongrod

Ephane Simon
email: stephane.simon@unilim.fr

Gilles Guibaud
email: gilles.guibaud@unilim.fr

Piet N L Lens

Yoan Pechaud

David Huguenot

Eric D Van Hullebusch

Lead sorption by biochar produced from digestates: Consequences of chemical modification and washing

Keywords:

Sewage sludge digestate (SSD) and the organic fraction of municipal solid waste digestate (OFMSWD) are currently considered as alternative feedstocks for biochar production due to the high amount of the organic solid waste remaining at the end of the treatment. The pyrolysis of solid digestate is known as an alternative to promote the recycling of organic wastes and generate added-value bio-products (e.g. biochar). Generally, the digestate biochar has a much lower sorption capacity for metal(loid)s compared to activated carbons. Therefore, chemical treatment is considered as a potential option to improve the biochar surface properties and thus inducing a better sorption ability for metal(loid)s on the biochar surface.

In this present work, the SSD and OFMSWD derived biochars were treated with 2 M KOH or 10% H2O2 followed by batch washing or batch and subsequent column washings with ultrapure water. The physicochemical properties including the pH of point of zero charge (pHPZC), the Brunauer-Emmett-Teller surface area (SBET) and cation exchange capacity (CEC) were determined for all the biochars in order to link their improved surface properties to the enhanced sorption ability for metal(loid)s. All the biochars were then used to study the influence of chemical treatment and biochar washing procedure on the sorption behavior of Pb(II), Cd(II) and As(III, V) through the batch sorption kinetics and isotherms. Moreover, the As redox state distribution (i.e. As(III) and As(V)) during the As(III) sorption onto the biochar surface and in liquid solution was determined by using solid-liquid extraction followed by liquid chromatographic analysis. X organische verbindingen uit deze biochar die kunnen reageren met Pb 2+ en zo Pb-ligand complexen in de oplossing kunnen vormen. Bovendien vertoonde de As redox distributie een significante oxidatie (70%) van As(III) tot As(V) in KOH-gemodificeerde SSD-biochar met batchwassing, terwijl As(III) gedeeltelijk werd geoxideerd (7%) in de KOH-gemodificeerde SSD-biochar met batch en daaropvolgend kolomwassingen. Dit benadrukt de belangrijke rol van de wasprocedure voor de sorptie van metalen, in het bijzonder voor Pb(II) en As(V).

De As-extractie gevolgd door vloeistofchromatografische analyse werd met succes vastgesteld om de As(III)-oxidatie met behulp van ascorbinezuur kwantitatief te herstellen en te behouden. Tijdens de sorptiekinetiek kan As(III) stabiel of gedeeltelijk geoxideerd zijn, afhankelijk van de biocharbehandeling. Bovendien werd de oxidatie van As(III) sterk geïnduceerd door het biochar materiaal en in mindere mate door het vrijkomen van opgeloste verbindingen uit de biochar.

Samenvattend kan digestaat met chemische behandeling, gevolgd door een goede biocharwasprocedure met, succes worden gebruikt als potentieel sorbens om de Pb(II), Cd(II) en As(III, V) sorptiecapaciteit te verbeteren. Bovendien kan de As redox-verdeling tussen de biochars en in oplossing tijdens de sorptie worden gemeten door As-extractie en chromatografische analyse, waardoor een beter inzicht wordt verkregen in de transformatie tussen As(III) en As(V) in het sorptiesysteem voor biocharvloeistofsorptie.

Results showed increases of the pHPZC, SBET and CEC after chemical treatment of the biochar, in accordance with the enhanced sorption ability for Pb(II), Cd(II) and As(V). For instance, the maximum sorption capacity (Qm) was increased from 1.6 µmol g -1 (As(V)) and 15.4 µmol g -1 (Cd(II)) on the raw SSD biochar to 8.1 µmol g -1 (As(V)) and 306.1 µmol g -1 (Cd(II)) after the H2O2 and KOH treatment, respectively (at initial pH 5.0).

Similarly, the Qm of Pb(II) was also increased from 31.4 µmol g⁻ 1 (raw SSD biochar) to 121.9 µmol g⁻ 1 on the H2O2 modified SSD biochar. However, the sorption capacity for Pb(II) was not determined after KOH treatment due to the failing of the Langmuir isotherm model to fit the experimental data. This indicates that insufficient washing of the KOH-modified SSD biochar can hinder the Pb(II) sorption due to the release dissolved organic compounds from this biochar that may interact with Pb 2+ and thereby forming Pb-ligand complexes in the solution. In addition, the As redox distribution II showed a large oxidation (70%) of As(III) to As(V) in KOH-modified SSD biochar with batch washing, while As(III) was partially oxidized (7%) in the KOH-modified SSD biochar with batch and subsequent column washings. This highlights an important role of washing procedure for sorption of metal(loid)s, particularly for Pb(II) and As(V).

The As extraction followed by liquid chromatographic analysis was successfully established to quantitatively recover and preserve As(III) oxidation with the use of ascorbic acid. During the sorption kinetics, As(III) may be stable or partially oxidized depending on the biochar treatment. In addition, the oxidation of As(III) was strongly induced by the biochar material and to a lesser extent by the release of dissolved compounds from the biochar.

In summary, digestate biochars with the chemical treatment followed by a proper biochar washing procedure can be successfully used as potential sorbents to enhance the Pb(II), Cd(II) and As(III, V) sorption capacity. Moreover, the determination of As redox distribution on the biochars and in liquid phase during the sorption process can be achieved through the As extraction and chromatographic analysis, providing a better understanding of the transformation between As(III) and As(V) in the biochar-liquid sorption system.

III

Tiivistelmä

Jätevesilietteen (SSD) ja kiinteän jätteen lietteen orgaaninen osuuden (OFMSWD) katsotaan tällä hetkellä olevan vaihtoehtoisia raaka-aineita biocharin tuotantoon käsittelyn jälkeisen korkean jäljelle jääneen kiinteän orgaanisen jätteen määränsä ansiosta. Kiinteän lietteen pyrolyysi tunnetaan vaihtoehtoisena menetelmänä, jolla edistetään orgaanisten jätteiden kierrätystä ja tuotetaan lisäarvoa tuottavia biotuotteita (esim. biochar). Yleisesti biochar on paljon alhaisempi sorptiokapasiteetti metalloid teihin verrattuna aktiivihiiliin. Siksi kemiallisen käsittelyn katsotaan olevan vaihtoehto biocharin pinnan ominaisuuksien parantamiseen ja täten paremman metall(oid)ien sorptiokyvyn indusointiin biocharin pinnalla.

Tässä työssä, SSD ja OFMSWD pohjaiset biochar it käsiteltiin 2 M KOH:lla tai 10% H2O2:lla jonka jälkeen ne eräpestiin tai eräpestiin ja kolonnipestiin ultrapuhtaalla vedell'. Fysikokemialliset ominaisuudet mukaanlukien isoelektrisen pisteen pH:n (pHPZC), Brunauer-Emmet-Tellerin pinta-alan (SBET) ja kationinvaihtokapasiteetin (CEC) määriteltiin kaikille biochareille, tavoitteena liittää niiden paremmat pintaominaisuudet metall(oid)ien lisääntyneeseen sorptiokykyyn. Kaikkia biochareja käytettiin sen jälkeen kemiallisen käsittelyn ja biochar pesun vaikutuksen tutkimiseen Pb(II):n, Cd(II):n ja As(III, V):n sorptiokäyttäytymiseen eräsoprptiokinetiikan ja isotermian avulla. Lisäksi, As redox-tila jakauma (As(III) ja As(V)) As(III):n sorption aikana biochar pintaan ja neste yhdisteeseen määriteltiin käyttämällä kiinteä-nesteuuttoa ja sen jälkeistä nesteen kromatograafista analyysia.

Tulokset osoittivat pHPZC:n, SBET:n ja CEC:n lisääntymisen biocharin kemiallisen käsittelyn jälkeen Pb(II):n, Cd(II):n ja As(V):n tehostetun sorptiokyvyn mukaisesti.

Esimerkiksi maksimaalinen sorptiokapasiteetti (Qm) kasvoi 1,6 umol g⁻ 1 :stä (As(V)) ja 15,4 umol g⁻ 1 :stä (Cd(II)) raakaa SSD-biocharista arvoon 8,1 umol g⁻ 1 (As( V)) ja 306,1 µmol g⁻ 1 (Cd(II)) H2O2:n ja KOH-käsittelyn jälkeen (alussa pH 5,0). Samoin Pb(II):n Qm:ää lisättiin 31,4 µmol g⁻ 1 :stä (raakaa SSD-biocharia) 121,9 µmol g⁻ 1 :een H2O2modifioidulla SSD-biocharilla. Pb(II):n sorptiokapasiteettia ei kuitenkaan määritetty KOH-käsittelyn jälkeen, koska Langmuir-isotermimallia ei saatu sopimaan kokeellisiin tuloksiin. Tämä osoittaa, että KOH-modifioidun SSD-biocharin riittämätön pesu voi haitata Pb(II)-sorptiota, joka johtuu liuenneista orgaanisista yhdisteistä, jotka voivat olla vuorovaikutuksessa Pb 2+ :n kanssa ja siten muodostaa Pb-ligandikomplekseja liuoksessa. Lisäksi As redox -jakauma osoitti suurta hapetusta (70%) As(III):sta IV As(V):hen KOH-modifioidussa SSD-biocharissa eräpesulla, kun taas As(III) hapetettiin osittain (7%) KOH-modifioitu SSD-biochar, jossa on erä-ja myöhemmät kolonnipestiin. Tämä korostaa pesumenettelyn tärkeää merkitystä metall(oid)in sorptiolle, erityisesti Pb(II):lle ja As(V):lle.

As-uutto ja sen jälkeinen nestekromatografinen analyysi suoritettiin onnistuneesti As(III):n hapettumisen kvantitatiivisen palautumisen ja säilyttämisen saavuttamiseksi askorbiinihapon avulla. Sorptiokinetiikan aikana As(III) voi olla stabiili tai osittain hapettunut biochar käsittelystä riippuen. Lisäksi biochar materiaali indusoi voimakkaan As(III):n hapettumisen ja vähäisemmän hapettumisen liuenneiden yhdisteiden vapautumisella biocharista.

Yhteenvetona voidaan todeta, että liete biocharit, joilla on kemiallinen käsittely ja oikeanlainen biochar pesumenettely, voidaan käyttää onnistuneesti sorbentteina Pb(II), Cd(II) ja As(III, V) sorptiokyvyn parantamiseksi. Lisäksi As redox-jakauma biocharilla ja nestemäisissä liuoksissa sorption aikana voidaan saavuttaa As-uutolla ja kromatografisella analyysillä, mikä antaa paremman käsityksen As(III):n ja As(V):n välisestä transformaatiosta biochar-neste-sorptiossa systeemissä.

V Sommario I digestati dai fanghi di depurazione (SSD) e dalla frazione organica dei rifiuti solidi urbani (OFMSWD) sono attualmente considerati materie prime alternative per la produzione di biochar in seguito all'elevata quantità di rifiuti solidi organici che rimangono alla fine del trattamento. La pirolisi del digestato solido è nota come alternativa per promuovere il riciclaggio di rifiuti organici e genera bio-prodotti a valore aggiunto (ad esempio biochar). In generale, il biochar da digestato ha una capacità di adsorbimento molto inferiore per metal(loid)i rispetto ai carboni attivati. Pertanto, il trattamento chimico è considerato come opzione per migliorare le proprietà superficiali del biochar e quindi indurre una migliore capacità di adsorbimento per metal(loid)i sulla superficie del biochar.

Nel presente studio, biochar derivati da SSD e OFMSWD sono trattati con una soluzione 2 M KOH o 10% H2O2, e in seguito lavati in batch o lavati in batch e successivamente in colonne con acqua ultra pura. Le proprietà fisico-chimiche del biochar tra cui il punto di carica zero (pHPZC), l'area superficiale di Brunauer-Emmett-Teller (SBET) e la capacità di scambio cationico (CEC) sono determinati al fine di associare le loro proprietà superficiali migliorate con la capacità di adsorbimento per i metal(loid)i. In seguito, il biochar è usato per studiare le conseguenze del trattamento chimico e delle procedure di lavaggio sul comportamento di adsorbimento di Pb(II), Cd(II) e As(III, V) mediante le cinetiche di adsorbimento e isoterme in batch. Inoltre, la distribuzione dello stato di ossidazione di As (cioè As(III) e As(V)) durante l'esperimento di assorbimento di As(III) sulla superficie del biochar e nella fase liquida è stata determinata utilizzando l'estrazione solido-liquido seguita da un' analisi cromatografica.

I risultati hanno mostrato un aumento di pHPZC, SBET e CEC dopo il trattamento chimico del biochar, in accordo con l'aumentata capacità di adsorbimento per Pb(II), Cd(II) e As(V). Ad esempio, la capacità massima di adsorbimento (Qm) è aumentata da 1,6 μmol g -1 (As(V)) e 15,4 μmol g -1 (Cd(II)) sul biochar SSD grezzo a 8,1 μmol g -1 (As(V)) e 306.1 μmol g -1 (Cd(II)) dopo il trattamento con H2O2 e KOH, rispettivamente (a pH iniziale 5.0). Allo stesso modo, il Qm di Pb(II) è aumentato da 31,4 µmol g -1 (biochar grezzo SSD) a 121,9 µmol g -1 sul biochar SSD modificato con H2O2. Tuttavia, la capacità di adsorbimento di Pb(II) in seguito al trattamento con KOH non è stata determinata a causa del fallito modello di isoterma Langmuir per spiegare i dati sperimentali.

VI

Ciò indica che un lavaggio insufficiente del biochar SSD modificato con KOH può ostacolare l'adsorbimento di Pb(II) dovuto ai composti organici disciolti da questo biochar che possono interagire con Pb 2+ e quindi formare complessi di Pb-legante nella soluzione. Inoltre, la distribuzione dello stato redox di As ha mostrato un'ampia ossidazione (70%) di As(III) a As(V) nel biochar SSD modificato con KOH e con lavaggio in batch, mentre As(III) è stato parzialmente ossidato (7%) nel biochar SSD modificato con KOH e con lavaggio in batch e successivo lavaggio in colonna. Ciò evidenzia un ruolo importante della procedura di lavaggio per l'adsorbimento di metal(loid)i, in particolare per Pb(II) e As(V).

L'estrazione di As, seguita da un'analisi cromatografica liquida, è stata considerata di successo per recuperare quantitativamente e preservare l'ossidazione di As(III) con l'uso di acido ascorbico. Durante la cinetica di adsorbimento, As(III) può essere stabile o parzialmente ossidato a seconda del trattamento del biochar. Inoltre, l'ossidazione di As(III) è stata fortemente indotta dal materiale da cui deriva il biochar e in misura minore dal rilascio di composti disciolti dal biochar.

In sintesi, il biochar prodotto dal digestato, trattato chimicamente e lavato con una corretta procedura può essere utilizzato con successo come potenziale assorbente di Pb(II), Cd(II) e As(III, V). Inoltre, la distribuzione dello stato di ossidazione di As sul biochar e le fasi liquide durante esperimenti di adsorbimento può essere verificata mediante l'estrazione di As e l'analisi cromatografica, che forniscono evidenza delle trasformazioni tra As(III) e As(V) nel sistema di adsorbimento biochar-fase liquida.

VII

Résumé

Les digestats des boues d'épuration (SSD) et les digestats de la fraction organique des déchets ménagers (OFMSWD) ont été récemment considérés comme des sources potentielles pour la production de biochars en raison des quantités grandissantes de digestats solides restant à la fin de la digestion anaérobie. La pyrolyse des digestats solides est connue comme une technique pour promouvoir le recyclage des déchets organiques et générer des bio-produits à valeur ajoutée (i.e. biochar). En outre, en raison d'une capacité de sorption des métal(loïde)s des biochars moins bonnes par rapport aux charbon actifs traditionnels, la modification chimique des biochars bruts est considérée comme une alternative pour améliorer les propriétés de surface des biochars et induire ainsi une meilleure capacité de sorption des métal(loïde)s.

Dans ce travail, les biochars SSD et OFMSWD ont été traités avec 2 M de KOH ou 10% de H2O2, suivis d'un lavage en batch seul ou batch combiné avec un lavage en colonne à l'aide d'eau ultrapure. Les analyses des propriétés de biochar, le pH du point de charge nulle (pHPZC), la surface spécifique de Brunauer-Emmett-Teller (SBET) et la capacité d'échange cationique (CEC) ont été effectuées sur les biochars bruts et modifiés afin de relier leurs propriétés de surface au comportement de sorption vis-àvis des métal(loïde)s. Tous les biochars ont ensuite été utilisés pour étudier l'influence du traitement chimique et de la procédure de lavage des biochars sur le comportement de sorption du Pb(II), Cd(II) et As(III, V) à travers l'étude de la cinétique et des isothermes de sorption. De plus, l'évolution de l'état redox As (i.e. As(III) et As(V)) pendant la sorption de l'As(III) sur la surface du biochar et en solution liquide a été déterminée par extraction solide-liquide suivie d'une analyse en chromatographie liquide.

Les résultats ont montré des augmentations de pHPZC, SBET et CEC après traitement chimique du biochar, concomitant avec l'augmentation de la capacité de sorption pour le Pb(II), le Cd(II) et l'As(V). Par exemple, la capacité de sorption maximale (Qm) a été augmentée de 1,6 µmol g -1 (As(V)) à 15,4 µmol g -1 (Cd(II)) sur le biochar de SSD brut à 8,1 µmol g -1 (As(V)) et 306,1 µmol g -1 (Cd(II)) après le traitements au H2O2 et KOH, respectivement (au pH initial de 5,0). De même, la valeur de Qm du Pb(II) a augmenté de 31,4 µmol g -1 (biochar de SSD) à 121,9 µmol g -1 sur le biochar modifié par H2O2. Néanmoins, la capacité de sorption du biochar SSD modifié par KOH n'a pas été déterminée en raison de l'impossibilité de modéliser les données expérimentales avec le modèle de l'isotherme de Langmuir. Cela indique qu'un lavage insuffisant du biochar VIII SSD modifié par KOH peut inhiber la sorption de Pb(II) en raison de la libération de composés organiques dissous de ce biochar pouvant interagir avec Pb 2+ et ainsi former des complexes Pb-ligand dans la solution. En outre, l'étude de la distribution de l'état redox de l'arsenic a montré une oxydation importante (70%) de As (III) en As (V) dans le biochar SSD traité au KOH avec lavage par batch, tandis que l'As(III) a été partiellement oxydé (7%) dans le biochar SSD traité au KOH avec un lavage en colonne. Ceci met en évidence le rôle important de la procédure de lavage sur l'efficacité de la sorption des métal(loïde)s, en particulier pour le Pb (II) et l'As (V).

L'extraction de l'arsenic fixé par les biochars suivie d'une analyse par chromatographie en phase liquide a été établie avec succès pour récupérer quantitativement et préserver l'oxydation de l'As(III) à l'aide d'acide ascorbique. Au cours de la cinétique de sorption, As (III) peut être stable ou partiellement oxydé en fonction du traitement chimique subit par les biochars. Il a été montré que l'oxydation de As(III) était fortement induite par le biochar et, dans une moindre mesure, par des composés dissous libérés par les biochars.

En résumé, les biochars de digestat modifiés par traitement chimique suivi d'une procédure de lavage appropriée du biochar peuvent être utilisés avec succès comme sorbants de Pb(II), Cd(II) et As(III, V). En outre, l'évolution de la distribution redox de l'arsenic dans les biochars et les solutions liquides à l'aide de l'extraction de liquide solide et de l'analyse chromatographique a été déterminée. Cela permet de mieux comprendre la transformation entre As(III) et As(V) lors la sorption de l'arsenic sur les biochars.

IX

Samenvatting

Zuiveringsslib digestaat (SSD) en de organische fractie van stedelijk vast afval digestaat (OFMSWD) worden momenteel beschouwd als alternatieve grondstoffen voor de productie van biochar vanwege de hoge hoeveelheid organisch vast afval die overblijft aan het einde van de behandeling. De pyrolyse van vast digestaat staat bekend als een alternatief om de recycling van organisch afval te bevorderen en bioproducten met toegevoegde waarde (bv. biochar) te genereren. Over het algemeen heeft de biochar van digestaat een veel lagere sorptiecapaciteit voor metalen in vergelijking met actieve kool. Daarom wordt chemische behandeling beschouwd als een mogelijke optie om de eigenschappen van het biocharoppervlak te verbeteren en zo een betere sorptiecapaciteit voor metalen op het biocharoppervlak te bewerkstelligen.

In dit werk werden de SSD en OFMSWD afgeleide biochars behandeld met 2 M KOH of 10% H2O2, gevolgd door batchwassing of batchwassing gevolgd door kolomwassing met ultrapuur water. De fysisch-chemische eigenschappen, waaronder de pH van het punt van nul lading (pHPZC), de Brunauer-Emmett-Teller oppervlakte (SBET) en kation exchange capacity (CEC) werden bepaald voor alle biochars om hun verbeterde oppervlakte-eigenschappen te koppelen aan de verbeterde sorptiecapaciteit voor metalen. Alle biochars werden vervolgens gebruikt om de invloed van chemische behandeling en de wasprocedure op het sorptiegedrag van Pb(II), Cd(II) en As(III, V) te bestuderen door de batch sorptiekinetiek en isothermen. Bovendien werd de As redoxstatusverdeling (d.w.z. As(III) en As(V)) tijdens de As(III) sorptie op het oppervlak van de biochar en in vloeibare oplossing bepaald met behulp van vaste-vloeistofextractie gevolgd door vloeistofchromatografische analyse.

De resultaten toonden stijgingen van de pHPZC, SBET en CEC na chemische behandeling van de biochar, in overeenstemming met de verbeterde sorptiecapaciteit voor Pb(II), Cd(II) en As(V). Zo werd bijvoorbeeld de maximale sorptiecapaciteit (Qm) verhoogd van 1,6 µmol g -1 (As(V)) en 15,4 µmol g -1 (Cd(II)) op de ruwe SSD-biochar tot 8,1 µmol g -1 (As(V)) en 306,1 µmol g -1 (Cd(II)) na, respectievelijk, de H2O2-en KOH-behandeling (bij initiële pH 5,0). Ook de Qm van Pb(II) werd verhoogd van 31,4 µmol g -1 (ruwe SSDbiochar) tot 121,4 µmol g -1 op de H2O2 gemodificeerde SSD-biochar. De sorptiecapaciteit voor Pb(II) werd echter niet bepaald na de KOH-behandeling als gevolg van het falen van het Langmuir isotherm model om de experimentele gegevens te fitten. Dit wijst erop dat onvoldoende wassen van de KOH-gemodificeerde SSDbiochar de Pb(II)-sorptie kan belemmeren als gevolg van het vrijkomen van opgeloste

List of Symbols and Abbreviations

BET

Brunauer-Emmett-Teller BMSW bat Biochar derived from organic fraction of municipal solid waste digestate with batch washing BMSW-H2O2 bat Hydrogen peroxide modified biochar from organic fraction of municipal solid waste digestate with batch washing BMSW-KOH bat Potassium hydroxide modified biochar from organic fraction of municipal solid waste digestate with batch washing BSS bat Biochar derived from sewage sludge digestate with batch washing BSS Biochar derived from sewage sludge digestate with batch and subsequent continuous column washing BSS-H2O2 bat Hydrogen peroxide modified biochar from sewage sludge digestate with batch washing BSS-H2O2

Hydrogen peroxide modified biochar from sewage sludge digestate with batch and subsequent continuous column washing BSS-KOH bat Potassium

Introduction

Metal(loid)s like arsenic (As), lead (Pb) and cadmium (Cd) are toxic and carcinogenic elements that mainly originate from anthropogenic activities such as discharges of wastewater from industries, agricultural use of pesticides and mining. Urban rivers are found to be increasingly contaminated with several metal(loid)s, particularly As, Cd and Pb [START_REF] Islam | Heavy metal pollution in surface water and sediment: A preliminary assessment of an urban river in a developing country[END_REF]. The metal(loid) pollutants often generate concerns over their potential effects on human health and the environment. Singh et al. (2015) reported that about 150 million people are affected by As poisoning in drinking water worldwide. To date, the use of certain metal(loid)s such as As-containing compounds has been banned due to a severe impact to health hazards. However, since the As, Cd and Pb are non-biodegradable and persistent, they tend to remain for a long time in the environmental cycle. Hence, the reduction of the As, Cd and Pb toxicity in water bodies becomes necessary.

To decrease the toxicity of As, Cd, and Pb in polluted water, the understanding of chemistry of each element is very important. The toxicity of the As, Cd, and Pb depends significantly on the chemical species of each specific element. Arsenic, as a metalloid, mainly exists in four oxidation states: arsenate (As(V)), arsenite (As(III)), arsenic (As(0)), and arsine (As(-III)). Among these species, arsenite (As(III)) and arsenate (As(V)) are known as the most toxic forms of As (Hughes et al., 2011), whereas metals like Cd and Pb mostly occur in divalent ion forms (i.e. Cd(II) and Pb(II)) in the environment.

Because of the acute toxicities of As(III, V), Cd(II) and Pb(II), many treatment approaches such as chemical precipitation, oxidation, coagulation and flocculation, ion exchange, phytoremediation, membrane separation and adsorption have been used to reduce the concentrations of these pollutants in water (Inyang et al., 2016;Jadhav et al., 2015;Jain and Singh, 2012;Singh et al., 2015). Sorption is considered as a cost efficient and simple treatment technique and biochar has recently been used as a potential sorbing material for metal(loid)s due to a great abundance of various biowaste feedstocks used for the production of biochar (Ding et al., 2016;Ofomaja et al., 2014;Peng et al., 2018;Wang et al., 2016;Zhou et al., 2017).

Currently, by-product organic waste digestates generated after the organic waste and wastewater treatment plants (e.g. sewage sludge digestate and the organic fraction of municipal solid waste digestate) are considered as an opportunity to produce low-cost and effective biochars for metal(loid)s removal from water/wastewater. Pyrolysis (300-650 °C) is the most favorable thermal process that helps to improve the digestate properties as well as adding more value to the biochar products. The obtained biochars usually contain increased surface functional groups and suitable physical and chemical properties, and thereby attracting more metal(loid)s to sorb onto these sorbents [START_REF] Kambo | A comparative review of biochar and hydrochar in terms of production, physico-chemical properties and applications[END_REF]. Nevertheless, certain raw biochars, particularly from the sewage sludge digestate, exhibit low adsorption capacities toward metal(loid)s, the biochar modification is required to enhance the sorption ability of such inorganic pollutants.

Biochar modification techniques such as physical and chemical treatments are considered as alternatives for improvement of the biochar properties. Physical treatment via steam activation provides new porosities and increases surface area (SBET) on the biochar. A significant increase of the SBET was observed on a willow biochar (i.e. from 11.4 to 840 m 2 g -1 ) with steam activation [START_REF] Kołtowski | Effect of biochar activation by different methods on toxicity of soil contaminated by industrial activity[END_REF]. However, the steam activation has a high operation cost due to its operation at high temperature (>800 °C) (Wang and Liu, 2018). Chemical modification of biochar is alternatively considered as a low-cost approach since no heat is used during the treatment. Biochar exposed to acidic, oxidative or alkaline solutions (e.g. H2O2 and KOH) induces the oxidation of biochar surface and thus more surface functional groups, particularly carboxyl and hydroxyl groups, are potentially generated onto the biochar surface (Sizmur et al., 2017;Xue et al., 2012). Such improved properties from chemical treatments could facilitate more sorption for metal(loid)s onto the biochar (Jin et al., 2014;Petrovic et al., 2016;Regmi et al., 2012). However, certain chemicals such as KOH have the ability to dissolve organic compounds from biochar (Lin et al., 2012;Liou & Wu, 2009;Liu et al., 2012). These dissolved organic compounds (DOC) could interfere with the metal sorption onto biochar by forming soluble metal-ligand complexes [START_REF] Mancinelli | Dissolved organic carbon content and leachability of biomass waste biochar for trace metal (Cd, Cu and Pb) speciation modelling[END_REF]. Therefore, the implementation of a biochar washing procedure after chemical treatment is required.

Generally, after the biochar chemical modification, several batch washings with ultrapure water until the pH becomes stable are performed on the biochar (Huang et al., 2017;Regmi et al., 2012;Wu et al., 2017) without any concern on the releasing compounds (e.g. DOC, PO4 3-, CO3 2-) from the modified biochar. A continuous washing of biochar by using a column reactor is considered as another option to eliminate all the releasable compounds from the biochar. It is recommended that the proper washing procedures after biochar treatment should be implemented in order to obtain a cleanedbiochar prior to its use as a potential sorbent for metal(loid)s.

This thesis focuses on the biochar production from organic waste digestates (i.e.

sewage sludge digestate and the organic fraction of municipal solid waste digestate) and its application as a sorbing material for As(III, V), Cd(II) and Pb(II) removal from aqueous solutions. The chemical modification by using H2O2 and KOH and washing procedures were applied on the biochars in order to obtain the chemically-modified biochars with improved surface properties. This study is also the first to report proper biochar washing procedures after the chemical treatment. The physical, chemical and structural properties of raw and chemically-modified biochars are also analyzed in detail.

The sorption kinetics and isotherms for As(III, V), Cd(II) and Pb(II) were performed in batch systems to investigate the sorption mechanisms and sorption behaviors during sorption process. Furthermore, the assessment of As redox species evolution during the As(III) sorption by biochars was performed. This study is also the first to highlight the role of biochar towards the As redox modifications.

2 Theoretical background

Metal(loid)s pollution in environment

Metals and metalloids water contamination are of high concern due to their persistence and toxicity towards aquatic organisms and human beings even at low concentrations. Metals like lead (Pb) and cadmium (Cd) are considered as toxic and carcinogenic elements in Europe [START_REF] Tóth | Heavy metals in agricultural soils of the European Union with implications for food safety[END_REF], which dominantly prevail in divalent ions, i.e. Pb(II) and Cd(II). The pollution sources of these metal elements mostly originate from smelters, mining, agricultural activities and wastewater discharges from industries. As a consequence, the contamination of Cd(II) and Pb(II) could pose severe threats to the aquatic environment. The Pb(II) poisoning of children has been heavily found in Haina (Dominican Republic), as a result from a car battery recycling factory (Kaul et al., 1999). [START_REF] Islam | Heavy metal pollution in surface water and sediment: A preliminary assessment of an urban river in a developing country[END_REF] reported a Cd(II) and Pb(II) contamination in an urban river in Bangladesh regarding the exceeded limits of such elements in this river, while the sediments were moderately to heavily contaminated with Cd(II), Pb(II) and arsenic (As).

Arsenic (As) is known as a metalloid which features both metal and nonmetal properties. Arsenic occurs naturally in the environment and its pollution was mostly caused from the mobilization under natural conditions. Nevertheless, anthropogenic activities such as mining, ore smelting and use of arsenic in industry and agriculture are other possible sources of As contamination in water. In the past century, arsenic has been used in pesticides, paint pigment, wood preservatives and constituents for products (Hughes et al., 2011). However, the use of arsenic-containing compounds has been restricted today due to their significant impact towards health hazards. Singh et al. (2015) reported that more than 150 million peoples are globally threatened by As contamination in drinking water and about 70% was in Asia, particularly in Bangladesh, China and India. Bangladesh was found to be exposed the largest As poisoning in groundwater in the world history [START_REF] Argos | Arsenic exposure from drinking water, and all-cause and chronic-disease mortalities in Bangladesh (HEALS): A prospective cohort study[END_REF]. Based on the World Health Organization (WHO) guidelines, the provisional concentration limit of As in drinking water is 10 µg L -1 . Nevertheless, the As concentrations in surface water and drinking water vary from 1-5000 µg L -1 depending on the sources of arsenic pollution (Singh et al., 2015). [START_REF] Kim | Distribution of inorganic arsenic species in mine tailings of abandoned mines from Korea[END_REF] reported heavy As contamination at the levels of 12-828 µmol g -1 in mine tailings in Korea which acts as a main source of As contamination in groundwater.

To reduce the concentration of As in polluted water, understandings of As chemistry and toxicity are of great importance. Arsenic toxicity is directly linked to its chemical speciation (particularly the redox state), which mainly exists in four oxidation states: arsenate (As(V)), arsenite (As(III)), arsenic (As(0)), and arsine (As(-III)) and the pH is an important parameter to control the solubility of arsenic in water. Among four species of As, the As(V) is the most stable form in aerobic water, while As(III) is predominant in reduced redox condition (Singh et al., 2015). Arsenic generally presents in various chemical forms which classifies into two types: organic and inorganic forms. The organic species of As such as monomethylarsonic acid (MMA) and dimethylarsinic acid (DMA) show intermediate toxicity, while the inorganic forms of As (i.e. As(III) and As(V)) are the most toxic ones that humans are exposed in the environment (Hughes et al., 2011).

Due to the acute toxicity of As(III, V), Cd(II) and Pb(II), numerous treatments such as chemical precipitation, oxidation, coagulation and flocculation, ion exchange, phytoremediation, membrane separation and adsorption have been dedicated to decrease their concentrations in contaminated water and effluent (Inyang et al., 2016;Jadhav et al., 2015;Jain and Singh, 2012;Singh et al., 2015). Among these techniques, sorption is known as a cost effective and simple treatment (Ding et al., 2016;Ofomaja et al., 2014;Peng et al., 2018;Wang et al., 2016;Zhou et al., 2017). To date, biochar has been increasingly used as a potential sorbing material in sorption treatment. Sorption of As(III, V), Cd(II) and Pb(II) by biochars produced from agricultural residues (e.g. pine wood, rice husk and switchgrass) has been intensively reported (Liu and Zhang, 2009;Regmi et al., 2012;Wang et al., 2015b). Nevertheless, by-product organic digestates obtained from organic wastes and/or wastewater treatment plants are currently considered as promising and abundant materials to produce low-cost and effective biochars for treating metal(loid)s in polluted water.

Biochar production from organic waste digestate

Over the last decade, biogas production by anaerobic digestion (AD) has been considered as a main source of renewable energy which has been essentially developed all over Europe with the commissioning of about 13,000 biogas facilities [START_REF] Plana | A review of the current digestate distribution models: Storage and transport[END_REF]. During the AD process, biogas is produced as a main product, and the remaining organic waste (i.e. digestate) as a by-product. Due to a significant growth of the biogas sector in Europe, the digestate quantity has substantially faced an important increase. The production capacities of digestate can be approximately 20 m 3 per year or about 10 tons per year at an averagely biogas plant capacity of 500 kW [START_REF] Plana | A review of the current digestate distribution models: Storage and transport[END_REF]. Consequently, the European Union (EU) has made progress on the development of digestate management in many countries; however, further essential implementation is still going on.

Currently, the agricultural utilization of digestate is one of the most applicable options in the EU. Nevertheless, the reuse of digestate in agriculture has faced technical problems due to insufficient management to have a well valorization system for digestate. In addition, a much higher loading capacity of digestate than its application on land (i.e. 1-2 times per year) hinders the digestate utilization in the fields [START_REF] Alvarenga | Sewage sludge, compost and other representative organic wastes as agricultural soil amendments: Benefits versus limiting factors[END_REF][START_REF] Fytili | Utilization of sewage sludge in EU application of old and new methods: A review[END_REF]. Landfilling is another alternative for the digestate treatment which accounts for 35-45% in Europe [START_REF] Plana | A review of the current digestate distribution models: Storage and transport[END_REF]. However, this technique will be phased out by the EU legislation due to high metal contents in certain types of digestates, particularly from pig manure. The digested pig manure usually contains high Cu and Zn concentrations that may significantly cause an environmental risk to soils [START_REF] Nkoa | Agricultural benefits and environmental risks of soil fertilization with anaerobic digestates: A review[END_REF][START_REF] Zhang | Nutrient status and contamination risks from digested pig slurry applied on a vegetable crops field[END_REF].

Today, pyrolysis of organic waste digestates (e.g. the organic fraction of municipal solid waste and sewage sludge digestates) has created a substantial interest by scientific community, particularly in Europe. This technology is considered as an environmentally sustainable approach due to its abilities to recycle the organic waste digestate, reduce the digestate quantity and generate added value bio-products such as biochar. The digestate-derived biochar can be further used as a potential sorbent towards metal(loid)s in polluted water.

The organic fraction of municipal solid waste digestate (OFMSWD) is the by-product sludge remaining after the organic fraction of municipal solid waste anaerobic treatment.

The organic fraction of municipal solid waste (OFMSW) mainly composes of organic portion which accounts for 30-40% in Europe [START_REF] Cesaro | The evolution of compost stability and maturity during the full-scale treatment of the organic fraction of municipal solid waste[END_REF]. The OFMSW is usually qualified by high moisture content and biodegradability regarding a large content of organic waste, food waste and leftovers from residences, restaurants, factories and markets [START_REF] Peng | Sustainable management of digestate from the organic fraction of municipal solid waste and food waste under the concepts of back to earth alternatives and circular economy[END_REF]. After the AD of the OFMSW, biogas and digestate are generated. The by-product digestate varies in compositions depending on the source, collection time and system [START_REF] Alibardi | Composition variability of the organic fraction of municipal solid waste and effects on hydrogen and methane production potentials[END_REF]. Based on the characteristics of the OFMSWD, it may cause adverse impacts and pose risks in the environment due to odors, volatile organic compounds (VOCs) and groundwater contamination by leachate [START_REF] Cesaro | The evolution of compost stability and maturity during the full-scale treatment of the organic fraction of municipal solid waste[END_REF].

The OFMSWD is usually enriched in Cu and Zn (Pituello et al., 2014), which is likely to concentrate during the AD process [START_REF] Tampio | Agronomic characteristics of five different urban waste digestates[END_REF]. A direct application of this organic waste in the field may be a concern due to the presence of metals as well as organic micropollutants at relatively high concentration in this waste digestate. Therefore, management options for the digestate are proposed to stabilize this waste digestate via several technologies such as thermal processes (e.g. pyrolysis and gasification) [START_REF] Garlapalli | Pyrolysis of hydrochar from digestate: Effect of hydrothermal carbonization and pyrolysis temperatures on pyrochar formation[END_REF]. Pyrolysis is considered as an alternative approach to eliminate the VOCs and recover material and/or energy from digestates including the OFMSWD as well as the one from anaerobic wastewater treatment plant (i.e. sewage sludge digestate).

Sewage sludge digestate (SSD) is an organic by-product generated after the anaerobic wastewater treatment. The SSD is rich in Cu, Cr, Pb and Zn (Pituello et al., 2014) as well as Fe and P (Yuan et al., 2015) due to the addition of FeCl3 for phosphate precipitation during the wastewater treatment. Moreover, the SSD is also rich in mineral fractions, depending on the diversity and complexity of each sewage sludge (Zielińska et al., 2015). High ash contents were found in sewage sludge (59%) due to the dominant mineral fractions such as SiO2, CaSO4•2H2O and CaCO3 contained in this waste (Zielińska et al., 2015). Like the OFMSW digestate, the SSD contains relatively high metal contents, thus thermal treatments remain as attractive alternatives to retain metals in the added-value solid biochar product.

Thermal treatment technologies (e.g. pyrolysis, hydrothermal carbonization and gasification) are known to improve the quality of digestate by adding value to the bioenergy products (i.e. syngas, bio-oil and biochar). Thermal treatments are normally operated at temperature ranging from 300 to 1000 °C (Inyang et al., 2016;Novotny et al., 2015). Figure 2.1 shows the overall biochar production via several technologies including slow and fast pyrolysis, gasification, and hydrothermal processes. Under slow pyrolysis, biochar is produced under low heating rates (10-30 °C min -1 ) and long residence times (5 min-12 h) at temperatures ranging from 300 to 650 °C (Figure 2.1).

In general, higher temperatures result in lower biochar yields due to partial degradation of lignin and cellulose [START_REF] Kambo | A comparative review of biochar and hydrochar in terms of production, physico-chemical properties and applications[END_REF]. The biochar production via slow pyrolysis provides high biochar yields (25-35%) and oxygen-containing surface functional groups such as hydroxyl and carboxyl [START_REF] Kambo | A comparative review of biochar and hydrochar in terms of production, physico-chemical properties and applications[END_REF]. Such functions may favor the sorption capacity towards metal(loid)s. Fast pyrolysis has been used mainly to produce biofuels (bio-oil) as main products, with biochar and syngas as by-products. Under fast pyrolysis, the process performs at 400-700 °C with a very high heating rate (~1000 °C s -1 ) and short residence times (< min) (Figure 2.1). High bio-oil yields (75%) and low biochar yields (10-15%) are often observed in fast pyrolysis (Mohan et al., 2014). Due to lower biochar yields and higher operation costs from fast pyrolysis, slow pyrolysis is recognized as a more favorable process for biochar production. The slow pyrolysis usually converts biomass (e.g. digestate) into a C-rich biochar with suitable physicochemical properties on biochar surface. Pre-treatments of digestate as a feedstock, e.g. sieving to small particle size (<2 mm) and drying (65-80 °C) to reduce the moisture to less than 10% are required to obtain homogenous biochar (Pituello et al., 2014). Similar to biomass, the digestate is mainly comprised of cellulose, hemicellulose, lignin, protein and fatty acids (Pituello et al., 2014). The transformation of digestate takes place along increasing temperature. Thermal decomposition of digestate occurred at temperature between 250 and 350 °C (Pituello et al., 2014). Hemicellulose started to decompose at temperatures from 200 to 260 °C, whereas cellulose and lignin degradation occurred at temperatures of 240-350 °C and 280-500 °C, respectively (Novotny et al., 2015). The biochar physicochemical properties are substantially affected by the production temperature. Zielińska et al. (2015) found that the pH values of sewage sludge biochars continuously increased along higher temperature in which the biochar becomes more alkaline at temperature beyond 500 °C. During pyrolysis, the organic matter from digestate also undergoes structural modifications such as the conversion of aliphatic forms to aromatic compounds (Zama et al., 2017), provide more stability (e.g. aromaticity) of biochar structures.

Table 2.1 shows the comparison of biochar sorption capacities for Pb(II) produced at different pyrolysis temperatures. From Table 2.1, the peanut shell biochar has the highest maximum sorption capacity (Qm) (254.8 µmol g -1 ) for Pb(II), while the lowest Qm was found on biochar from rice husk (8.6 µmol g -1 ) (at initial pH 5.0). The biochar from pine wood also showed a low sorption for Pb(II) (18.8 µmol g -1 ), while much higher Pb(II) sorption abilities were found on biochars from dairy waste digestate, sugar beet digestate and medicine residue (196.9-248.0 µmol g -1 ) (Table 2.1). Due to low sorption capacities of certain biochars, they are not completely suitable as sorbing materials. Therefore, pre-or post-treatments of biochars such as physical and chemical activations (Sizmur et al., 2017) to improve the sorption efficiency for metal(loid)s become necessary. Physicochemical properties such as C, H, O, N, mineral content, specific surface area and structural properties are also developed on the biochar surface (Pituello et al., 2014;Yuan et al., 2015;Zielińska et al., 2015). Due to such improved properties of biochar, it has abilities to retain nutrients, reduce carbon dioxide emission in soil and remove several contaminants both in water and soil (Ahmad et al., 2014;Inyang et al., 2016;Novotny et al., 2015). Biochar generally contains both negative and positive surface charges, depending on the pH.

The pH of point of zero charge (pHPZC) is the pH at which the biochar is net neutral or contains equal numbers of positive and negative charges on its surface. At the pH condition below the pHPZC, the biochar exhibits net positive charges, while it is net negatively-charged at the pH condition beyond the pHPZC. Changing of the pH condition will alter the surface charges of biochar and thus it will strongly affect the possibility of electrostatic interaction. Therefore, the pH is required to be optimized in order to improve the sorption of metal(loid)s onto the biochar. At circumneutral pH, where negatively-charged biochars are predominant, they have been used for sorption of metals like lead (Pb), zinc (Zn) and cadmium (Cd) from polluted water (Ding et al., 2016;Ho et al., 2017;Park et al., 2017).

Electrical conductivity (EC) and cation exchange capacity (CEC) are also known as important parameters for the biochar. The EC is a measure of dissolved ionic species present in the solutions, which is often measured in term of the capacity to transmit electrical current. Since the electrical current is transported by the ions in the solutions, the EC increases as the solutions contain more dissolved ionic species. Therefore, the EC is considered as a key parameter for the biochar application towards metal(loid) sorption from the aqueous solutions. The CEC is a measure of the biochar ability to exchange cationic elements (e.g. Ca 2+ , Mg 2+ , K + and Na + ) on its surface with metal ions in the solutions. Therefore, the CEC parameter can be an indicator to show the possibility of positively-charged metal ions (e.g. Pb 2+ and Cd 2+ ) to sorb onto the biochar surface via cation exchange mechanism. 2.2, the highest sorption ability for As(V) was found on a carbonaceous nanofiber with the Qm of 670 µmol g -1 , in a comparable range with a magnetic carbonaceous tea waste (507 µmol g -1 ) (both at initial pH 5.0). In addition, biochar from paper mill sludge was able to efficiently sorb both As(V) and Cd(II) with the Qm of 303 and 369 µmol g -1 , respectively (Table 2.2). However, biochar produced from rice straw showed relatively low Qm for both As(III) and As(V), i.e. 5.9 µmol g -1 and 7.3 µmol g -1 , respectively (Table 2.2). Due to low sorption abilities of certain biochars, biochar modification is considered as an effective approach to increase the sorption efficiency for metal(loid)s in water.

Modification and washing of biochar

Biochar modification technologies such as physical and chemical activation have recently been used to improve properties of biochar. Physical activation generally provides high-temperature steam to clean biochar pore sites and create new porosities, which consequently increases surface area (SBET) on the biochar. [START_REF] Lima | Physicochemical and adsorptive properties of fast-pyrolysis bio-chars and their steam activated counterparts[END_REF] found an increase of the SBET of biochar produced from broiler litter, from 4.6 to 136 m 2 g -1 after steam activation. A significantly higher SBET on a steam activated willow-derived biochar, from 11.4 to 840 m 2 g -1 , was also reported [START_REF] Kołtowski | Effect of biochar activation by different methods on toxicity of soil contaminated by industrial activity[END_REF]. Nevertheless, the steam activation is often operated at high temperature (>800 °C), resulting in a high operation cost, particularly for a large scale operation (Wang and Liu, 2018).

Chemical treatment of biochar is the most common and low-cost technique as no heat is required during the operation. Acidic or alkaline treatments induce an oxidation of biochar, which consequently creates more oxygen-containing functional groups on the biochar surface (Sizmur et al., 2017;Xue et al., 2012). The SBET also increases after exposing the biochar to chemical solutions (e.g. H2O2 and KOH) (Jin et al., 2014;Wang and Liu, 2018). Acidic treatment (e.g. H2O2 as an oxidant) helps to remove mineral fractions from biochar, thus inducing a more hydrophilic property of the biochar [START_REF] Shen | Surface chemical functional groups modification of porous carbon[END_REF]. An increase of surface functional groups, particularly carboxyl, was also observed on the biochar treated with H2O2 (Xue et al., 2012). Similarly, alkaline treatment also produces more surface functional groups (e.g. hydroxyl and carboxyl), but increases basicity in the biochar material (Huang et al., 2017;Jin et al., 2014;Petrovic et al., 2016). Such improved properties from both H2O2 and KOH treatments induce more sorption for metal(loid)s like As(III), As(V), Cu(II), Cd(II) and Pb(II) onto the modified biochars (Jin et al., 2014;Petrovic et al., 2016;Regmi et al., 2012).

Due to a chemical property of KOH to dissolve ash and condense organic matter (e.g. lignin and cellulose) in the modified biochar (Lin et al., 2012;Liou & Wu, 2009;Liu et al., 2012), a release of dissolved organic compounds (DOC) and mineral ash from the modified biochar can be observed. The release of DOC from biochar could interact with metal ions such as Cd(II), Cu(II) and Pb(II), and thereby affecting the mobility of such elements through the formation of soluble metal-ligand complexes [START_REF] Mancinelli | Dissolved organic carbon content and leachability of biomass waste biochar for trace metal (Cd, Cu and Pb) speciation modelling[END_REF]. Since this DOC interferes the sorption of metals onto the biochar, the elimination of such released compounds from biochar becomes necessary. To date, a concern to implement a proper biochar washing procedure after chemical treatment is scarce and requires more attention by scientists.

In general, after the chemical treatment of biochar, several batch washings of biochar with ultrapure water until the pH becomes stable are usually performed (Huang et al., 2017;Regmi et al., 2012;Wu et al., 2017). Nevertheless, these batch washings are often conducted without a concern on the release of organic or inorganic compounds (e.g. DOC, PO4 3-, CO3 2-, Ca 2+ and Mg 2+ ) from the biochar. Currently, a continuous column washing of biochar is proposed as an alternative to control the elimination of the releasable compounds and thereby obtain a complete-washed biochar. The selection of biochar washing procedure should be studied with a caution, especially considering the test on biochar ability to release organic and inorganic compounds, particularly after chemical treatment. The properly washed biochar can be further used as a potential sorbent to remove metal(loid)s from water, without any interference of dissolved compounds during sorption treatment.

Application of biochar for metal(loid)s removal from water

Recently, the application of biochar for metal(loid)s removal from polluted water has received attention by many researchers. Biochar is recognized as an effective sorbing material to immobilize both metals and oxyanions (such as As) due to a great abundance of feedstocks as well as suitable physicochemical properties regarding the biochar's sorption ability. Studies on the sorption of metal(loid)s by biochars are important to tackle water pollutants and understand the sorption mechanisms in order to determine the sorption behaviors of every inorganic element.

To assess the sorption of metal(loid)s by biochar in aqueous solutions, a set of batch experiments on adsorption kinetics and isotherms is often conducted. The adsorption kinetics are performed to obtain the time required for each metal(loid) to reach equilibrium state. Generally, metal sorption by the biochar reaches the equilibrium within 24 h, while a longer time may be required for other pollutants (such as organic pollutants). After obtaining the equilibrium time, the sorption isotherms are further conducted to predict the sorption mechanisms and the maximum adsorption capacity (Qm) of biochar obtained from the Langmuir isotherm equation (see section 4.3.2 in chapter 4).

Figure 2.2 shows the main sorption mechanisms involved in metal(loid)s sorption by the biochars including cation exchange, surface complexation, surface precipitation and physical adsorption. Among these mechanisms, cation exchange of metals with Ca 2+ and Mg 2+ is considered as one of the main contributor for the sorption of positivelycharged metals (e.g. Pb 2+ and Cd 2+ ) by biochars (Figure 2.2), which accounts for 40-52% (Li et al., 2017). Cation exchange can be also occurred with K + and Na + to a lesser extent (<8.5%) (Li et al., 2017), which is also reported by Lu et al. (2012). Furthermore, the surface complexation between metals and surface functional groups of biochars (such as carboxyl and hydroxyl groups) also plays an important role, contributing for about 40% of metal removal (Li et al., 2017). In addition, surface precipitation may also occur as the sludge-based biochars usually contain a high amount of phosphate (PO4 3-) and carbonate (CO3 2-) on the biochar surface (Figure 2.2). A high specific surface area of biochar could favor a physical sorption of metal(loid)s onto the biochars (Agrafioti et al., 2014). The relative importance of these sorption mechanisms depends on the origin of biochar feedstocks. From the literature, sorption of Pb(II) by activated carbons (Cechinel et al., 2014), agricultural waste-derived biochars (e.g. pine wood or rice husk) (Liu & Zhang, 2009), natural zeolite and kaolinite clay (Andrejkovicova et al., 2016;Jiang et al., 2009) have also been intensively reported. The higher metal(loid)s sorption capacities are probably due to favorable physicochemical properties (e.g. surface charges) of such biochars (Liu & Zhang, 2009;Zielińska et al., 2015). As previously discussed, the sorption abilities for Cd(II) by carbonaceous materials are also high, compared to the digestatebased biochar (Table 2.2). However, even if low sorption abilities for As(V) and Cd(II) by the SSD digestate biochar were observed (Table 2.2), the modification of this biochar induces increases of the sorption toward these pollutants. For instance, the Cd(II) sorption capacity (at initial pH 5.0) of the SSD biochar was enhanced from 15.4 µmol g -1 (raw biochar) to 218.7 and 306.0 µmol g -1 , respectively, on the H2O2 and KOH modified biochars (Wongrod et al., 2018b). Jin et al. (2014) found an enhanced As(V) sorption (at initial pH 6.0) by municipal organic waste-derived biochar after alkaline treatment, i.e. from 325.6 to 412.4 µmol g -1 . The Qm for Pb(II) was also increased from 368.7 µmol g -1 (raw manure biochar) to 818.0 µmol g -1 on the H2O2-modified biochar (Wang and Liu, 2018). The findings from literature affirm that the chemically-modified biochars are efficient to physically or chemically sorb both metals and metalloids. Nevertheless, due to many existing chemical species of some oxyanions, particularly As in water (as described in section 2.1), the investigation of arsenic redox transformation during biochar-As sorption treatment should be considered. This is to tackle the arsenic species evolution in water bodies as well as in the solid phase biochar since the arsenic has different affinity toward different biochar types, and therefore the sorption behaviors of biochars can be altered through the interaction with different As species.

Arsenic contamination in surface water and groundwater has become a major concern to affect human health worldwide, particularly in Asia (Singh et al., 2015). Human exposure to As even at low concentrations poses a severe threat to human health [START_REF] Argos | Arsenic exposure from drinking water, and all-cause and chronic-disease mortalities in Bangladesh (HEALS): A prospective cohort study[END_REF]. Of all As species, arsenite (As(III)) and arsenate (As(V)) are noted as the most toxic forms in the environment. The sorption is known as a simple and potential treatment technique to immobilize As from water. Therefore, the assessment of As redox species distribution during the As sorption by biochar becomes interesting since there is no information on the role of biochar toward the As redox modification.

The transformation between As(III) and As(V) is naturally occurs both in soil and water, depending on the surrounding environmental conditions (e.g. pH and redox potential). Since As(III) is more mobile and weakly bound to solid material, the As(III) is far more toxic than As(V) (Manning et al., 2002). Therefore, the biochar can be used to induce the oxidation of As(III) to As(V), and as a result reducing the As toxicity in polluted water. Chemical redox speciation of As is of great importance to quantify and tackle As(III) and As(V) species both on solid phase of biochar and in liquid exposition solutions during the sorption between As species and the biochar.

Generally, arsenic speciation in solid-phase biochar can be accessed via X-ray absorption near edge structure (XANES) spectroscopy (Niazi et al., 2018a(Niazi et al., , 2018b)). Nevertheless, due to a relatively high operation cost and a less accessibility to synchrotron facilities, the use of this equipment is often limited to few researchers. Hence, there is a need to implement a technology for the analysis of As redox species on solid-phase samples that can be easily accessed by the scientific community. Solidliquid extraction is considered as a conventional technique that can be efficiently used to recover As from the solid material. Solid-liquid extraction followed by separation techniques (e.g. liquid chromatography (LC)) coupled to spectrometric detection techniques (e.g. atomic fluorescence spectroscopy (AFS) and inductively coupled plasma mass spectrometry (ICP-MS)) is currently considered as an effective technique for determination of the As speciation on the solid-phase biochar after the sorption process. The use of these coupling techniques can provide a useful information on the As redox distribution in both liquid solutions and on the biochar surface.

Aims of the present work

Due to a great availability of digestates, the valorization of these waste streams is considered as an alternative option to eliminate the toxic volatile organic compounds (VOCs) from the digestates and obtain added-value biochar product. The biochar can be further used as a medium in sorption process, particularly for metal(loid) removal.

The aim of this research work was to develop biochars derived from by-product organic waste digestates as sorbents for metal(loid)s removal from water. To improve their sorption capacity, chemical treatments and biochar washing procedure were investigated. The specific objectives of this study are summarized as follows:

-Determine the changes of physicochemical properties of biochars produced from sewage sludge digestate (SSD) and the organic fraction of municipal solid waste digestate (OFMSWD) before and after chemical treatments (Papers I, II and III).

-Investigate the effect of chemical treatments and washing procedures of biochars on the enhancement of Pb(II), Cd(II) and As(III and V) sorption capacity (Papers I, II and III).

-Observe the As(III) and As(V) redox state distribution in the solid biochars and in solutions during arsenic sorption by using conventional chromatographic analysis instead of solid-phase analysis (Papers II and III).

Material and methods

Biochar preparation

Two organic waste digestates were used as feedstocks for biochar production and then studied as sorbents for metal(loid)s removal from water (Papers I, II and III). Dewatered sewage sludge digestate (SSD) was collected in the storage tank after dewatering and drying processes from a domestic wastewater treatment plant (WWTP) located in Limoges, France. The SSD was further dried and then submitted to slow pyrolysis at industrial scale (Biogreen® technology) at 350 °C for 15 min (Table 4.1). These production conditions were selected based on previous studies by Pituello et al. (2014) and Yuan et al. (2015). The organic fraction of municipal solid waste digestate (OFMSWD) was obtained from a solid waste treatment plant located in France. As illustrated by Figure 4.1, the OFMSWD was dried at 65 °C for 24 h to reduce the initial moisture content to less than 10% and sieved into a particle size of 2 mm to remove impurities such as plastic bags, needles and glasses from the samples before pyrolysis (Paper I). The biochar was then produced under slow pyrolysis at lab scale using porcelain crucibles with lidcovers (Haldenwanger 79 MF, Germany) in a muffle furnace. To find the optimal conditions for biochar production, the OFMSWD was pyrolyzed at different temperatures: 200, 300, 350, 400, 450 and 500 °C, at a residence time of 1 h and a heating rate of 15 °C min⁻ 1 (Figure 4.2). Results showed a partial conversion of the OFMSWD to the biochar at temperatures of 200-350 °C due to a remaining of organic matter (brown color) in the crucible. In addition, the presence of ash content at temperatures beyond 400 °C was observed, according to remained-oxygen in the muffle that may induce the combustion reaction (Figure 4.2). Therefore, pyrolysis at 400°C for 1 h was at the optimal conditions regarding a complete conversion of organic matter into biochar. As discussed in chapter 2, post-treatments after biochar production are required to enhance the biochar sorption capacity for metal(loid)s. Both the SSD and OFMSWD biochars were thus chemically-treated with either a 10% H2O2 solution (i.e. 100 g biochar per 1 L of a 10% H2O2 solution) (modified from Xue et al., 2012) or a 2 M KOH solution (100 g biochar per 2.5 L of a 2 M KOH solution) (modified from Jin et al., 2014) with continuous stirring at 25 °C for 2 h. To eliminate releasable ions (e.g. PO4 3 ⁻, HCO3⁻/CO3 2 ⁻, Ca 2 ⁺ and Mg 2 ⁺) and organic compounds from the biochar after the chemical treatments, the biochars were then washed with ultrapure water. Two different approaches were tested: (1) a batch washing and ( 2) the same batch washing followed by a continuous column washing. The overview of two different washing procedure after biochar treatments is illustrated in For the batch washing, 2 g of biochar were washed with 200 mL of ultrapure water at 20 (± 2) °C. The batch washing was performed in triplicate in a row until the pH became stable (Jin et al., 2014;Wu et al., 2017;Xue et al., 2012). For the column washing, a part of the batch-washed biochar was submitted to a subsequent continuous washing using ultrapure water in a glass column (2.8 cm in diameter and 42.5 cm in height). Glass beads (2.0 mm in diameter) were filled at the bottom of the glass column to maintain the flow distribution. Peristaltic pumps (Ismatec Reglo Analog, Model No. ISM827, Ismatec SA Company, Switzerland) were used to control the up-flow velocity (0.77 ± 0.01 cm min⁻ 1 ) with a hydraulic retention time of 6 h. The pH, electrical conductivity (EC), the concentrations of PO4 3 ⁻, HCO3⁻/CO3 2 ⁻ and dissolved organic carbon (DOC) were measured to follow the progress of washing (Paper I). The column washing was considered completed once these parameters reached threshold values, i.e. stable pH, similar EC between inflowing and outflowing rinsing water solutions, and low concentrations (<1 mg L⁻ 1 ) of PO4 3 ⁻, HCO3⁻/CO3 2 ⁻ and DOC at the outflowing rinsing water solutions. Once the biochar washing procedure were accomplished, the biochars were dried in an oven at 50 °C for 24 h and kept in a desiccator for further use.

The batch and column-washed biochars were consecutively studied (Figure 4.3) to investigate the influence of the washing procedure on sorption capacity for Pb(II), Cd(II) and As(III, V) as well as its influence on the conversion of redox-sensitive elements (in solution and on the solid phase) during As(III) sorption (Table 4.2). The raw OFMSWD biochar and its H2O2 and KOH modified biochars with batch washing are labeled as BMSW bat , BMSW-H2O2 bat and BMSW-KOH bat , respectively. The SSD biochars and its H2O2 and KOH modified biochars with batch washing are denoted as BSS bat , BSS-H2O2 bat and BSS-KOH bat , respectively. The H2O2 and KOH modified SSD biochars with batch and subsequent column washings are labeled as BSS-H2O2 and BSS-KOH, respectively. All information concerning chemical treatments, washing procedure and sample labelling are reported in Table 4.2.

Biochar characterization

Chemical, physical and structural analyses were performed on feedstocks, raw and modified biochars to determine the changes of properties occurring at each modification step, i.e. pyrolysis and chemical treatments. The pH and EC of material suspension in water as well as the pH of point of zero charge (pHPZC), the total metal(loid) contents and the cation exchange capacity (CEC) of raw and modified biochars were determined. The Brunauer-Emmett-Teller (BET) N2 sorption, X-ray diffraction (XRD) and Fourier transform infrared spectroscopy (FTIR) were applied to, respectively, determine specific surface area, crystalline structure and surface functional groups of the raw and modified biochars (Papers I and II). In addition, arsenic redox distribution in liquid solution and solid phase of biochars were also determined using liquid chromatography coupled to atomic fluorescence spectrometry (LC-AFS).

All analytical methods and instruments used in all experiments are shown in Table 4.3 and detailed in the following subchapters. To measure the pH of the raw feedstocks, the raw and chemically-modified biochars, a suspension of solid sample in deionized water (1 g: 20 mL) was stirred continuously for 5 min and let to settle for 15 min, then the sample suspension was measured using a pH-meter (LPH 330T, Tacussel, France). The electrical conductivity (EC) of all biochars were measured using a conductivity-meter (CDM 210, Radiometer, Denmark).

The pH of point of zero charge (pHPZC) of raw and modified biochars was determined from zetametry by determining the zeta potential at different pH ranges (Mahmood et al., 2011). Firstly, 40 mL of background electrolyte solution (0.01 M NaNO3) containing 0.2 g of biochar were equilibrated with continuous magnetic stirring at 20 (± 2) °C for 40 min. The pH of the mixture was adjusted in the range of 2.0-10.0 by using HNO3 (0.1, 0.5 and 1 M) and NaOH (0.1, 0.5 and 1 M). After 40 min of stirring, the pH was measured and corrected until each solution reached a stable targeted pH. The suspension was then filtered with a 0.45 µm cellulose acetate (CA) syringe filter before determining the zeta potential (Zetasizier Nanoseries (Nano-ZS), Malvern, UK). This filtration was performed to prevent the settlement of biochar particles in the measurement cell of the zetameter and to avoid the clogging of the device by big particles. We hypothesize that all biochar particles display the same behavior, neglecting a particle size effect. The pHPZC values of biochars were determined from the different zeta potential along the pH ranges (2.0-10.0).

For the determination of the biochar cation exchange capacity (CEC), 2 g of biochar sample (dry weight) were mixed with 40 mL of 0.05 N cobalt hexamine trichloride ([Co(NH3)6]Cl3) solution (99% w/w, Sigma-Aldrich) in polypropylene tubes (Aran et al., 2008). The suspensions were shaken for 1 h to allow the sorption of cobalt (Co) on solid phase of biochars using an orbital shaker (KS 501 digital, IKA TM , USA) at 60 rpm, then centrifuged at 7000 g for 10 min (Rotina 420, Hettich, Germany). The supernatant was filtered through Whatman polyethersulfone (PES) syringe filters (0.2 µm). The pH (Präzisions-pH-meter E510, Metrohm AG, Switzerland) and absorbency at 472 nm (UV-vis spectrophotometer, Lambda 365, PerkinElmer, USA) of each sample were measured immediately. All experiments were performed in triplicate and the results are reported as mean values. The CEC (meq 100 g⁻ 1 or cmol⁺ kg⁻ 1 ) can be calculated from differences of absorbencies between initial [Co(NH3)6]Cl3 solution and remaining Co in residual solution using Eq. ( 1) (Aran et al., 2008):

CEC = [ Abs Co -Abs sample Abs Co ] × 50 × V m × 100 (Eq. 1),
where AbsCo and Abssample refer to the absorbency (at 472 nm) of 0.05 N [Co(NH3)6]Cl3 and of sample (supernatant after filtration), respectively. V and m correspond to solution volume in liter and biochar dry mass in gram, respectively.

Total metal contents including Al, Ca, Cd, Cr, Cu, K, Ni, Pb, Mg, Mn, Na, Fe and Zn in raw feedstocks and biochars were determined by inductively coupled plasma optical emission spectroscopy (ICP-OES) (Optima 8300, PerkinElmer, France) after acid digestion performed according to the Standard Method 3050B (EPA, 1996) by using H2O2 (30% w/w, Sigma-Aldrich), concentrated HCl (37% w/w, Merck) and HNO3 (65% w/w, Merck). The digestion was performed using a heating block at 95 (± 5) °C. The digested solutions were filtered using a Whatman grade 1 qualitative filter paper and metal(loid) elements.

During the continuous column washing of the raw and chemically-modified biochars, liquid samples were collected at the outlet of the column along the time to measure pH, EC, DOC, PO4 3 ⁻ and HCO3⁻/CO3 2 ⁻. DOC was measured following the AFNOR French standard method (NF EN 1484Juillet 1997;AFNOR, 1997a). A colorimetric method (NF EN 1189Janvier 1997;AFNOR, 1997b) was applied to measure PO4 3 ⁻ with a spectrophotometer at λ 700 nm using AFNOR standard method. For HCO3⁻/CO3 2 ⁻ analysis, a titrimetric method from AFNOR (NF EN ISO 9963-1 Février 1996; AFNOR, 1996) was performed using HCl (10⁻ 3 M) as a titrant and phenolphthalein and mixed bromocresol green-methyl red as indicators. The total releasable amounts of Ca 2+ and Mg 2+ were measured using microwave plasma atomic emission spectroscopy (MP-AES), respectively, at λ 422.673 nm and λ 285.213 nm.

Brunauer-Emmett-Teller specific surface area (SBET) of all biochar samples were measured using the N2 sorption method at 77 K. Before measurement, the biochars were dried at 105 °C for 5 h (Memmert, France).

X-ray diffraction (XRD) was employed to identify the crystalline structures of the raw feedstocks, the raw and modified biochars. The samples were grounded to less than 100 µm particle size and were characterized using a powder diffractometer (AXS D8, Bruker, Germany) with Cu Kα radiation at 1.54 Å wavelength over the 2θ range from 10° to 32°, for 10 seconds per step, at 40 kV voltage and 40 mA current with a Solx (Si-Li) detector. The crystalline compounds present in biochar were identified using the X'Pert HighScore software for data analysis with the reference from the International Center for Diffraction Data (ICDD).

Fourier transform infrared spectroscopy (FTIR) was used to identify the functional groups present on biochar surfaces. Each sample was mixed with KBr in a ratio of 1 mg sample (particle size 100 µm) per 200 mg KBr and the pellet was prepared by compression under vacuum (Jouraiphy et al., 2005). The analysis was performed using a Shimadzu IRAffinity-1 Spectrometer with a deuterated-triglycine sulfate detector. The scanning was performed from 4000 to 800 cm⁻ 1 wavenumber at a resolution of 16 cm⁻ 1 .

Batch sorption experiments

Sorption kinetics and isotherms were performed to, respectively, determine the time required to reach sorption equilibrium and the maximum sorption capacity of biochars. Three sorption kinetic models (i.e. pseudo-first-order (PFO), pseudo-second-order (PSO) and intraparticle diffusion) and two sorption isotherm models (i.e. Langmuir and Freundlich) were used to describe the experimental data as well as the sorption processes (Zhao et al., 2015a).

All batch sorption experiments for As(III), As(V), Cd(II) and Pb(II) were separately performed at an initial pH value of 5.0 (± 0.2), adjusted using HNO3 or NaOH (Papers I, II and III). The stock solutions of As(III), As(V), Cd(II) and Pb(II) were prepared separately from AsNaO2, Na2HAsO4.7H2O, Cd(NO3)2 and Pb(NO3)2. The solid/liquid ratio was set to 4 g of biochar per 1 L of metal(loid) solution and the resulting suspensions were shaken at 20 (± 2) °C using an orbital shaker (KS 501 digital, IKATM, USA) at 180 rpm. The equilibrium time for sorption tests, determined from sorption kinetic studies, was 24 h. All sorption experiments were conducted in duplicate (Papers I and III) and triplicate (Paper III), and the mean values are reported followed by standard deviation. Additional analyses were conducted if a variation over 10% was observed.

After sorption equilibrium time was reached, the solid phase of biochar was separated from the solution by filtration through Whatman polyethersulfone (PES) syringe filters (0.2 µm). The remaining Cd and Pb concentrations in solution were determined by microwave plasma atomic emission spectroscopy (MP-AES) at λ 228.8 nm and 405.781 nm, respectively (Paper I). The As concentration in the solution was measured using graphite furnace atomic absorption spectrometry (GF-AAS) at λ 193.7 nm.

The study of sorption kinetics for metal(loid)s removal from aqueous solutions is important, providing the information about the reaction pathways and the mechanism of the sorption processes. Sorption kinetics are normally used to describe the uptake rate of pollutants (e.g. metal(loid)s) sorbed by adsorbents (e.g. biochar) at different contact times. Therefore, it is significant to know the kinetic rate for metal(loid)s removal from aqueous solutions until the equilibrium state is reached. In general, three types of kinetic models: the pseudo-first-order (PFO), pseudo-second-order (PSO) and intraparticle diffusion are usually used to explain the experimental data.

The PFO kinetic model was first introduced around the 19 th century by Lagergren (1898). However, the kinetic studies became more popular when [START_REF] Ho | Pseudo-second order model for sorption processes[END_REF] have rearranged the PFO equation into a linearized form and proposed the PSO kinetic model to better describe the correlation of the experimental data. The intraparticle diffusion kinetic model was used to determine the sorption mechanisms such as the degree of boundary layer like film diffusion and mass transfer between liquid-solid interface (Nethaji et al., 2013;Z. Zhao et al., 2015). The following equations are illustrated as followed:

Pseudo-first-order:

log(Q e -Q t ) = logQ e - k 1 t 2.303 (Eq. 2), Pseudo-second-order: t Q t = 1 k 2 Q e 2 + t Q e (Eq. 3),
Intraparticle diffusion:

Q t = k i t 1 2 ⁄ + C (Eq. 4),
where Qt and Qe are sorption capacity (µmol g⁻ 1 ) at time t (h) and at equilibrium, k1

(h⁻ 1 ), k2 (g(µmol h)⁻ 1 ) and ki (µmol (g h 1/2 )⁻ 1 ) are the rate constants for the PFO, PSO and intraparticle diffusion kinetic models, respectively (Liu and Zhang, 2009;Z. Zhao et al., 2015). C is the intraparticle diffusion constant related to the boundary layer diffusion. All sorption kinetic parameters and corresponding uncertainties were obtained by non-linear regression using Statistica software (v6.1, StatSoft).

Sorption isotherms are generally used to determine the adsorption capacity of metal(loid)s onto biochars at different metal(loid) concentrations in bulk solutions under the equilibrium condition. Several isotherm models have been used to describe the experimental data such as the Langmuir, Freundlich, Temkin, and Dubinin-Radushkevich models. Langmuir and Freundlich, the two most common isotherm models, were used to study the sorption behaviors and predict the removal efficiency for metal(loid)s. The following equations are expressed as follows:

Langmuir model:

C e Q e = 1 K L Q m + C e Q m (Eq. 5),
Freundlich model:

lnQ e = lnK F + 1 n lnC e (Eq. 6),
where Ce is the equilibrium solution concentration (µM), Qe and Qm are the adsorption capacity at equilibrium and the maximum adsorption capacity (µmol g⁻ 1 ), respectively.

KL and KF are the sorption constants for the Langmuir and Freundlich isotherm models, respectively, and n is the Freundlich constant which indicates the favorability of sorption.

Arsenic analysis during sorption tests

During sorption for As, the redox reaction between As(III) and As(V) can be possibly occurred. Therefore, the study on arsenic redox state evolution during sorption for As(III) by biochars was further performed in aqueous solutions under batch experiments (Paper III). Total As and As species (i.e. As(III) and As(V)) were individually quantified in both liquid solutions and on the solid-phase of biochars during and/or after the sorption process. The detailed information for As analysis is described in the following subchapters.

Total As was analyzed using GF-AAS (240Z, Agilent Technologies, USA) at λ 193.7 nm. The redox distribution of As species was determined by LC-AFS with hydride generation (HG) (PS Analytical Millennium Excalibur, PS Analytical, UK) with conditions detailed by Wan et al. (2014). Arsenic redox evolution in solutions obtained from solid-liquid extraction was analyzed within 5 h after recovering the samples in order to avoid the oxidation of As(III) to As(V).

After sorption experiments, the amount of total As sorbed onto the biochar (Qsol., in µmol) was calculated from the difference between the initial and final concentrations of As (Ci and Cf, in µmol L⁻ 1 ) and multiple with the total volume of the solution (V, in L) according to Eq. ( 7):

Q sol. = (C i -C f ) × V (Eq. 7),
The sorption capacity of biochar (in µmol g -1 ) for As can be further obtained by dividing the Qsol. with the mass of biochar (on dry weight).

Since there is no soft extraction procedure for As from biochar available in the literature, an extraction procedure was designed (Paper III) based on previous studies reported by Thomas et al. (1997), Montperrus et al. (2002) and [START_REF] Zhang | Efficiency and mechanisms of Cd removal from aqueous solution by biochar derived from water hyacinth (Eichornia crassipes)[END_REF].

The extraction was performed by 0.3 M phosphoric acid (H3PO4, 85% w/w, Carlo ERBA). Ascorbic acid (0.5 M) (99.5% w/w, Fluka) was added to avoid the As oxidation in the solution during the extraction steps (Xu et al., 2015).

After sorption tests, biochars were quickly rinsed with ultrapure water in order to eliminate the excessive As solution from the biochar. The biochar (0.15 g) was then transferred into a Teflon tube with 25 mL of the extracting solution and heated at 80 °C for 20 min using a microwave device (Multiwave GO, Anton Paar, France). Then, the extracted solutions were recovered by centrifugation (Multifuge X3 FR, Thermo Fisher Scientific) at 3400 rpm for 20 min. The supernatant was collected and filtrated through a 0.2 µm PES syringe filters, then transferred to a 50 mL gauged flask completed with ultrapure water before analysis by LC-AFS. All extractions were performed in triplicate.

After sorption experiments, the biochar (about 0.15 g) was rinsed 3 times with ultrapure water and transferred into a digestion tube. 6 mL of concentrated nitric acid (HNO3, 69.5% w/w, Panreac ITW) were added then 6 mL of hydrogen peroxide (H2O2, 30% w/w, Carlo ERBA) were slowly added into the tube to avoid fast reaction due to the strong oxidizing property of this chemical. The digestion tube was left for 12 h to avoid over-pressure during microwave-assisted digestion regarding the high initial release of CO2 from samples. 3 mL of concentrated hydrochloric acid (HCl, 37% w/w, VWR) were then added to the tube and the microwave-assisted digestion was performed at 180 °C for 4 h in 4 cycles (1 h per 1 cycle) to ensure a complete digestion of biochar material. The digested solutions were recovered and diluted with ultrapure water to a constant volume of 50 mL. The filtration of samples was performed using 0.2 µm PES syringe filters prior to total As analysis by GF-AAS at λ 193.7 nm.

Statistical analysis

The physicochemical properties of biochar and sorption experiments for metal(loid)s were performed in triplicate and the values are reported as means followed by standard deviation unless otherwise stated. Constants for sorption kinetic and isotherm parameters were stimulated by the non-linear regression using Statistica software (v6.1, StatSoft). Statistical analysis of the experimental data was performed by the t-test with two-tailed distribution at a statistical significance level of P≤0.01 for data comparison.

Result and discussion

Characterization of digestates and biochars

Table 5.1 shows the properties of the organic fraction of municipal solid waste digestate (OFMSWD) and its derived biochar (BMSW), and of the sewage sludge digestate (SSD) and the SSD-derived biochar (BSS) obtained after pyrolysis. Comparison of two digestates, a higher pH value was found on the OFMSWD (8.0 ± 0.1) compared to the SSD (6.0 ± 0.1). This is in accordance to higher alkali metal elements (e.g. Ca, Mg, Na and K) contained in the OFMSWD than in the SSD (Table 5.1). However, metals such as Cu, Mn and Zn were mainly found in both solid digestates.

Pyrolysis also induced changes of pH values between digestates (raw feedstocks) and biochars. For both the OFMSWD and SSD, higher pH values were observed for biochars, i.e. from 8.0 (± 0.1) for OFMSWD to 8.9 (± 0.1) for BMSW and from 6.0 (± 0.1) for SSD to 6.5 (± 0.1) for BSS. The increases of pH can be due to the concentration of inorganic elements/compounds in the biochars according to the separation of alkali metal salts from the organic matrix [START_REF] Shinogi | Pyrolysis of plant, animal and human waste: Physical and chemical characterization of the pyrolytic products[END_REF]Zielińska et al., 2015).

During pyrolysis, metals were retained in the biochars, while As (metalloid) was partly volatilized at the temperature below 400 °C. The As transformation during thermal treatment was also observed by Duan et al. (2017). Consequently, higher metal contents were found in both the BMSW and BSS than in the digestates, whereas a lower As content was found in the BSS (0.63 ± 0.02 µmol g⁻ 1 ) compared to the SSD (1.05 ± 0.04 µmol g⁻ 1 ) (Table 5.1). These findings were in correlation to the study from Pituello et al. (2014). Due to a strong uncertainty of As value for the BSS and no As information for both the OFMSWD and BMSW, more caution should be concerned when using these samples. It was also noticed that a 4-fold higher Ca content was found in the BMSW compared to the BSS (Table 5.1). This could induce the cation exchange between Ca 2+ and metal ions and thus more metal sorption can be expected on the BMSW. Nevertheless, both the BMSW and BSS display the presences of Al, Fe and Mn (Table 5.1) that could be partially in metal oxide forms such as Al2O3, Fe2O3 and MnO. Therefore, this could promote the metal ions to sorb onto the biochars (Agrafioti et al., 2014;Wang et al., 2015a). show the change of XRD spectra between feedstocks and biochars after pyrolysis of the OFMSWD and SSD, respectively. XRD analysis showed that OFMSWD contained the crystalline structures of quartz (SiO2) and calcite (CaCO3) (Figure 5.1(a)). Similar trends were also found in its derived biochar with slight shifts on the XRD spectra. On the other hand, only SiO2 was detected in SSD and BSS (Figure 5.1(b)). The results showed no significant differences of XRD spectra between the digestates and the biochars. This indicates that the pyrolysis only induced the insignificant modification of surface properties of biochars. Nevertheless, the presence of CaCO3 on the biochar, in this case on the BMSW, may contribute to more sorption of metal elements (Al-Degs et al., 2006;Jiang et al., 2009). ). Similar surface functional groups were also detected on their derived biochars. Nevertheless, a strong intensity of -CH2 scissoring (1427 cm⁻ 1 ) was only found in the OFMSWD and its biochar (BMSW), whereas C=C stretching vibrations (1600 cm⁻ 1 ) was only detected in the SSD and BSS. Such surface groups presented on the biochars can take part in the sorption interaction for metal(loid)s (Chen et al., 2011;Pituello et al., 2014). Results from FTIR spectra indicated that the pyrolysis only induced small variations between the digestates and biochars, particularly at temperature below 400 °C. The results also demonstrated the differences in surface functional groups between BMSW and BSS corresponding to different feedstock type and pyrolysis temperature. To understand the effect of chemical treatment on biochar properties, the sewage sludge digestate-derived biochar (BSS) was selected to study its physicochemical properties before and after chemical treatment. Table 5.2 shows the corresponding properties of the raw BSS and its chemically-modified biochars.

Chemical treatment had an effect on the pH changes of the biochars. Similar pH values were found for both the BSS (6.4 ± 0.1) and its H2O2-modified biochar (6.5 ± 0.1), whereas the value was strongly increased for KOH-modified biochar with batch and Note: all values reported are mean of triplicate followed by standard deviation, except pHPZC and zeta potential.

subsequently continuous column washings (8.4 ± 0.1). The increase of pH is likely due to the property of KOH to increase the hydroxyl (-OH) groups as well as the basicity on the modified biochar (Fan et al., 2016;Li et al., 2014).

The electrical conductivities (EC) of the raw and modified biochar suspensions remained in similar ranges, except for KOH-modified biochar with batch washing (BSS-KOH bat ). A much higher EC of the BSS-KOH bat among other biochars can be due to the ability of KOH to modify the biochar surface by dissolving inorganic elements/compounds from the biochar (Lin et al., 2012;Liou and Wu, 2009;Liu et al., 2012). Therefore, a release of inorganic ions (e.g. phosphate and carbonate) from this KOH-modified biochar contributes to its high EC value (Paper I).

Figure 5.3 shows the evolution of the zeta potential along the pH ranges for raw and modified biochars from SSD. The pH of point of zero charge (pHPZC), determining the pH at which the net charge on the biochar surface is zero, was obtained from this curve. The corresponding values of the zeta potential at pH 5 and pHPZC are presented in Table 5.2. Results showed similar pHPZC values for both BSS (~2.7) and BSS-H2O2 (~2.9), whereas a slightly higher pHPZC was observed for BSS-KOH (~3.4). Nevertheless, due to small variations of pHPZC (2.7-3.4) among all biochars, they are not prone to strongly modify the pH on the sorption for metal(loid)s, except if the experiments are performed at pH below 5. From the literature, the pHPZC of biochar has been rarely reported and the information on the variability of biochar with different chemical treatments is limited. However, our pHPZC results are in consistent with biochars from a straw residue (~1.9) (Qiu et al., 2009) and paper mill sludges (4.0-6.0) (Guo et al., 2018). In addition, the negative values of zeta potential at pH 5 (Table 5.2) indicated that the raw and modified biochars were net negatively charged, for instance -16.5 mV and -17.1 mV for the BSS-H2O2 and BSS-KOH bat , respectively. The cation exchange capacity (CEC) of all biochars are given in Table 5.2. Results showed a slight increase of CEC after H2O2 treatment of biochar, i.e. from 1.9 (± 0.1) cmol⁺ kg⁻ 1 (BSS) to 2.9 (± 0.1) cmol⁺ kg⁻ 1 (BSS-H2O2), whereas a 6-fold increase of CEC was observed on the BSS-KOH. A significantly higher CEC in the KOH-modified biochar may result from the chemical property of KOH to modify the surface of biochar, as previously discussed. Therefore, more possibility of exchangeable cations (e.g. Na + , K + , Mg 2+ and Ca 2+ ) associated on the biochar to exchange with metal elements like Pb(II) and Cd(II). However, the CEC values found (Table 5.2) were still far lower than those reported for wood biochar (45.7 cmol⁺ kg⁻ 1 ) and straw-derived biochar (483.4 cmol⁺ kg⁻ 1 ) (Ding et al., 2016;Jiang et al., 2014).

The SBET of all biochars were also improved (7-20 times) after the chemical treatments, e.g. from 0.4 (± 0.1) m 2 g⁻ 1 for raw biochar to 5.7 (± 0.1) m 2 g⁻ 1 for BSS-H2O2 and 7.9 (± 0.1) m 2 g⁻ 1 for BSS-KOH (Table 5.2). The increase of SBET implies that H2O2 and KOH induced higher porosities on the modified biochars. Our results are consistent with those reported for sewage sludge biochars (4.0-14.3 m 2 g⁻ 1 ) (Agrafioti et al., 2013;Yuan et al., 2015) and pineapple-peel-derived biochars (0.7-2.1 m 2 g⁻ 1 ) (Wang et al., 2016). However, our SBET values are still far lower than those reported for activated carbons (1215.0-1316.0 m 2 g⁻ 1 ) (Gong et al., 2015).

The total As content in raw and chemically-modified SSD biochars was analyzed and the values are reported in Table 5.2. Significant decreases of the As content were found after treating the raw biochar with KOH or H2O2, for example from 0.63 (± 0.02) µmol g⁻ 1 for BSS to 0.39 (± 0.02) µmol g⁻ 1 for BSS-H2O2 and 0.17 (± 0.02) µmol g⁻ 1 for BSS-KOH (Paper III). The reduction of As in the modified biochars can be due to the ability of KOH and H2O2 to modify the solid phase of the raw biochar. As a result, As could be released from biochar matrix during the chemical treatments.

In summary, the increase of CEC and SBET values after biochar treatments highlights that the chemically-modified biochars, particularly with KOH provided higher exchangeable cations and active pore sites on their surfaces. Therefore, the sorption . For example, the high concentrations of PO4 3 ⁻ (55 mg L⁻ 1 ), CO3 2 ⁻ (99 mg L⁻ 1 ) and HCO3⁻ (1418 mg L⁻ 1 ) released from KOH-modified biochar were related to the high conductivity (391 µS cm⁻ 1 ) at the initial washing period (Figure 5.4(a, c)). The released compounds were continuously declined along the lower conductivities, indicating that these two profiles were correlated (Figure 5.4(a, c)). The similar trends were also observed for the BSS-H2O2 bat ; however, with much lower concentrations of the PO4 3 ⁻, CO3 2 ⁻, HCO3⁻ and DOC (Figure 5.4(b,d)).

Comparison between KOH and H2O2 treatments, the biochar washings demonstrated that the KOH-modified biochar released much higher DOC, PO4 3 ⁻ and HCO3⁻/CO3 2 ⁻ than the H2O2 biochar. This is according to different abilities of KOH and H2O2 to modify the solid phase of biochar (Paper I). In addition, the column washing was considered complete when stable pH values, similar EC of inlet/outlet solutions (i.e. 3-6 µS cm⁻ 1 ) and low concentrations (<1 mg L⁻ 1 ) of PO4 3 ⁻, HCO3⁻/CO3 2 ⁻ and DOC were observed, i.e. beyond 72 h for the KOH-modified SSD biochar (BSS-KOH bat ) and 22 h for the H2O2-modified biochar (BSS-H2O2 bat ).

From the previously published data (Huang et al., 2017;Regmi et al., 2012;Wu et al., 2017), pH stabilization was normally used as a key parameter for biochar batch washing. However, our findings highlight that it was not sufficient to eliminate the dissolved organic matter and releasable ions from biochar (Paper I). A direct control of the release of PO4 3 ⁻ and HCO3⁻/CO3 2 ⁻ from the biochar is of importance because these releasable ions can hinder the sorption for metal(loid)s onto the biochars. Therefore, the conductivity is proposed as a more accurate indicator rather than the pH (Paper I).

The continuous washing of biochars and the conductivity monitoring were further applied in Papers II and III. Due to the potential release of organic and inorganic elements/compounds from the KOH-modified biochars, they were further selected to study the influence of biochar washing procedures on the physicochemical properties of biochars (Paper II). The observations on the changes between the observed biochar properties in previous subchapters (5.1.1 and 5.1.2) and the ones found during biochar washing procedures were parallelly established.

The pH, EC, SBET and CEC, total As content of the KOH-modified biochars with batch and continuous washings are given in Table 5.2. The results showed that the uncomplete batch washing after biochar treatment influenced the change of biochar properties. The pH of KOH biochar was decreased after a subsequent column washing, i.e. from 10.0 (± 0.1) for BSS-KOH bat to 8.4 (± 0.1) for BSS-KOH. This is consistent with a dramatic decrease of EC value by 50-time in the BSS-KOH (6.2 ± 0.5 µS cm⁻ 1 ), compared to the BSS-KOH bat (324 ± 2 µS cm⁻ 1 ) (Table 5.2). These differences are likely due to the releasable ions (e.g. PO4 3 ⁻, HCO3⁻/CO3 2 ⁻, Ca 2 ⁺ and Mg 2 ⁺) from the BSS-KOH bat that significantly altered both the pH and EC values (Papers I and II). These findings are also supported by Yuan et al. (2015), as the authors reported a high EC value (7750 µS cm⁻ 1 ) in the sewage sludge biochar with no washing with ultrapure water after pyrolysis.

A decrease of CEC was observed for the BSS-KOH, compared to the BSS-KOH bat , i.e. from 20.8 (± 0.1) cmol + kg -1 (BSS-KOH bat ) to 13.4 (± 0.1) cmol + kg -1 (BSS-KOH).

Similarly, the total As were also reduced after the KOH biochar was subsequently washed in a column reactor. In this case, from 0.31 (± 0.03) µmol g⁻ 1 (BSS-KOH bat ) to 0.17 (± 0.02) µmol g⁻ 1 (BSS-KOH). However, a higher SBET was found on the BSS-KOH compared to BSS-KOH bat (Table 5.2). The reduction of CEC and total As content but increase in SBET could be explained by the elimination of releasable compounds (e.g. DOC, PO4 3 ⁻ and HCO3⁻/CO3 2 ⁻) from the biochars. Hence, lower risk of As toxicity and more available active pore sites, while lower exchangeable cations on the BSS-KOH.

These findings highlight that the pH, EC, SBET, CEC, and total As content were strongly influenced by the biochar washing procedure. Therefore, the careful washing procedure after biochar chemical treatment is recommended to avoid interferences during sorption for metal(loid)s.

Sorption kinetics for Pb(II), Cd(II) and As(III, V) by biochars

The study on sorption is highly important to provide insights regarding the sorption reactions and the sorption mechanisms. It also helps to predict the rate for metal(loid) removal from aqueous solutions and the sorption capacity of a sorbent (e.g. biochar) towards metal(loid) elements. In this study, the sorption experiments were performed to determine the influence of chemical treatments and biochar washing procedures on the sorption capacities of biochars for Pb(II), Cd(II) and As(III, V) (Papers I and II). The sorption kinetics were firstly conducted to obtain the time required for Pb(II), Cd(II) and As(III, V) to reach equilibrium state. After that, the sorption isotherms were performed to predict the maximum sorption capacity of biochar for each metal(loid) element. The link between the biochar properties and the sorption ability of biochar towards Pb(II), Cd(II) and As(III, V) are discussed in detail in the following subsections.

Figure 5.5 shows the influence of contact time on the Pb(II) sorption by raw biochars (i.e. BMSW bat and BSS bat ) from the organic fraction of municipal solid waste digestate (OFMSWD) and sewage sludge digestate (SSD) and their H2O2 or KOH modified biochars (BMSW-H2O2 bat , BMSW-KOH bat , BSS-H2O2 bat and BSS-KOH bat ) submitted to batch washing. The pseudo-first-order (PFO) and pseudo-second-order (PSO) kinetic models were used to describe the kinetic data for Pb(II) (Paper I). Table 5.3 shows the corresponding parameters obtained from fittings of the PFO and PSO models with the Pb(II) kinetic data by raw and chemically-modified biochars. The constants obtained (Table 5.3) show that the Pb(II) sorption kinetics were strongly fitted to the PSO model with high coefficients of correlation (R 2 >0.99) for all the raw and modified biochars. This suggests that simple chemical reactions between Pb(II) and biochars were literally taken place (Nethaji et al., 2013). Due to the best-fit of the Pb(II) kinetic data to the PSO model, the intraparticle diffusion kinetic model does not apply for Pb(II).

BMSW bat and BSS bat showed different Pb(II) sorption kinetics as the sorption equilibrium was achieved within 2 h for the BMSW bat , while it was beyond 5 h for the BSS bat (Figure 5.5). In addition, a rapid Pb(II) sorption at initial state followed by a slow reaction until reaching an equilibrium were observed on the BMSW bat , while a slower Pb(II) sorption reaction throughout the kinetic time was found on the BSS bat (Figure 5.5). These different behaviors were due to variations of biochar properties among different origins of digestate feedstocks. A stronger intensity of O-containing functional groups (1033 cm⁻ 1 ) in the OFMSWD than SSD was observed (Figure 5.2). Such property could promote a faster Pb(II) kinetic removal rate due to more possibility of Pb ions to bind onto the O-containing functional groups of biochar (Bogusz et al., 2017). Furthermore, a much higher cationic element content such as Ca, K, Mg and Na were found on the BMSW bat , compared to the BSS bat (Table 5.1). This could induce the cation exchange interaction with Pb(II) and thus enhancing the sorption ability onto the BMSW bat (Paper I). These results are in agreement with a previous study of Zhang et al. (2017) on the use of celery biochars with high alkaline mineral contents to remove metals from water.

Considering the effect of biochar chemical treatments with KOH or H2O2, the kinetic rates for Pb(II) sorption were different on the chemically-modified biochars (Paper I).

Results showed that the KOH induced faster Pb(II) sorption kinetic rates by both BSS-KOH bat and BMSW-KOH bat compared to their raw biochars (Figure 5.5). In this case, the equilibrium times for the KOH-modified biochars were achieved within 2 h. Similarly, a previous study from Regmi et al. (2012) showed an enhanced kinetic removal rate for metal by a switchgrass biochar with the KOH treatment. Nevertheless, the biochar treated with H2O2 may induce or hinder the Pb(II) sorption kinetics, depending on the biochar feedstock types. For instance, an equilibrium time for Pb(II) removal rate was slower on the BSS-H2O2 bat (beyond 7 h) than on the BSS bat (5 h). However, an improved equilibrium time for Pb(II) was observed on the BMSW-H2O2 bat (within 1 h) compared to its raw biochar (2 h) (Figure 5.5). In addition, the BMSW-H2O2 bat showed the highest PSO sorption kinetic rate (k2) among all biochars (Table 5.3). This suggests that Pb 2+ ions strongly interact with external surface sites of the biochar and consequently forming the outer-sphere complexes of Pb(II) onto its surface (Wang et al., 2015a). BMSW bat , BMSW-H2O2 bat and BMSW-KOH bat refer to raw biochar from the organic fraction of municipal solid waste digestate, and its H2O2 and KOH modified biochars with batch washing, respectively. BSS bat , BSS-H2O2 bat and BSS-KOH bat refer to raw biochar from sewage sludge digestate, and its H2O2 and KOH modified biochars with batch washing, respectively.

The improvement of equilibrium time on the chemically-modified biochars was directly linked to the alteration of biochar properties after the chemical treatment. Due to the ability of chemical reagents, particularly KOH, to modify the biochar properties such as increase/decrease pH on the biochar surface, dissolve inorganic compounds and condense organic matter from the biochar (Fan et al., 2016;Li et al., 2014). Consequently, this could promote the faster Pb(II) kinetic removal rates on the chemically-modified biochars. Nevertheless, the source of feedstocks should also be considered since this study found a slightly slower Pb(II) sorption rate on the H2O2modified biochar from SSD, compared to its raw SSD biochar.

Due to a potential release of organic/inorganic compounds from biochars observed in section 5.1.3, the batch and subsequent continuous washings were further applied to raw SSD biochar and its chemically-modified biochars. These manipulations were performed to eliminate the released compounds from the biochars in order to provide suitable biochars for Cd(II) and As(III, V) removal from aqueous solutions (Paper II).

The batch and subsequent column washings of the raw SSD biochar and its H2O2 and KOH modified biochars are noted as BSS, BSS-H2O2 and BSS-KOH, respectively.

Similar to Pb(II), the PFO and PSO kinetic models were used to stimulate the experimental data for Cd(II) and As(III, V) by all biochars (Paper II). The intraparticle diffusion kinetic model was additionally applied for the Cd(II) and As(III, V) in order to obtain more information on the sorption mechanisms (Zhao et al., 2015). Figure 5.6 shows the Cd(II) sorption by BSS, BSS-H2O2 and BSS-KOH as a function of contact time with the PSO kinetic model-fittings. All the corresponding kinetic constants of the PFO, PSO and intraparticle diffusion kinetic models are given in Table 5.4.

Results showed that the experimental data for Cd(II) sorption was best described by the PSO model (Paper II). Similar to Pb(II), faster PSO kinetic rates for the Cd(II) sorption were found on both the KOH and H2O2 modified biochars, compared to its raw biochar (BSS) (Figure 5.6). For instance, the BSS reached equilibrium after 10 h, while less than 5 h for the chemically-modified biochars (Figure 5.6). The enhancement of the Cd(II) sorption kinetic rates onto the chemically-modified biochars is likely due to the ability of chemicals (such as KOH or H2O2) to modify the biochar properties as previously discussed in a subparagraph for Pb(II) sorption kinetics. In addition, based on the intraparticle diffusion model-fitting parameters in Table 5.4, the C constants are not equal to zero, literally suggesting that there were some degrees of boundary layer and the rate limiting step was not only controlled by the intraparticle diffusion. Other mechanisms such as film diffusion and external mass transfer may take part in the sorption processes (Nethaji et al., 2013). In addition, the sorption kinetics for metals can be controlled by the particle sizes of biochars, and thus the kinetics may be limited by the diffusion of metals in the biochars [START_REF] Rees | Short-term effects of biochar on soil heavy metal mobility are controlled by intra-particle diffusion and soil pH[END_REF]. Apart from metals like Pb(II) and Cd(II), the effect of biochar chemical treatment on the sorption kinetics for metalloids such as arsenic (As) was further studied. This manipulation was performed to observe if the chemical treatment of biochar induces similar sorption behaviors for As, compared to those for Pb(II) and Cd(II). 5.5.

Comparison of the raw and modified biochars showed that the As(V) sorption kinetic rates (k2) of the BSS (0.630 ± 0.098) and BSS-H2O2 (0.619 ± 0.092) were similar (Table 5.5), reaching the equilibrium time after 10 h (Figure 5.7). However, unlike the Pb(II) and Cd(II), a much slower kinetic rate for As(V) sorption by the KOH modified biochar (BSS-KOH) than by the raw biochar was observed with the equilibrium time being reached beyond 24 h (Figure 5.7). Nevertheless, Wang et al. (2015a) also showed the slower As(V) sorption kinetic rate by biochar with chemical treatment, compared to its raw pine wood biochar. The slow rate for As(V) sorption by the KOH biochar could be due to the formation of inner layer complexes via surface complexation between the hydroxyl groups on the biochar surface and the arsenic species [START_REF] Velazquez-Peña | As(V) sorption by different natural zeolite frameworks modified with Fe, Zr and FeZr[END_REF]. This statement was also supported by previous studies [START_REF] Catalano | Simultaneous inner-and outer-sphere arsenate adsorption on corundum and hematite[END_REF]Wang et al., 2015b). In summary, the sorption kinetics for Pb(II), Cd(II) and As(V) onto all biochars exhibited the best-fit by the PSO model (Papers I and II), theoretically suggesting that the sorption between these metal(loid)s and biochars mainly occurred via simple chemical reactions (Nethaji et al., 2013). Moreover, the BSS-KOH showed the fastest kinetic rate for Pb(II) and Cd(II) sorption among other biochars, while the slowest rate was observed for As(V) sorption. This slower rate of the As(V) sorption onto the BSS-KOH could be described by its high ki rate constant (Table 5.5), suggesting that the intraparticle diffusion mainly involved in the sorption mechanism of As(V). These findings also indicate that the KOH biochars only accelerated the sorption kinetics for metals, while more time was required for As(V) to reach the equilibrium state by this biochar.

Due to the chemical treatment of biochar, particularly with KOH, that induced a strong release of organic and inorganic compounds from the biochar into solutions (see section 5.1.3), it was therefore decided to focus on the effect of biochar washing procedures after KOH treatment on the sorption for metal(loid) elements. To observe if the biochar washing procedures affect the metal(loid)s sorption, the KOH-modified biochar from SSD with batch washing was additionally submitted to a continuous column washing (see section 4.1.2). After these manipulations, the KOH-modified biochars with a batch washing (i.e. BSS-KOH bat ) versus batch and subsequent column washings (i.e. BSS-KOH) were further compared on the sorption for Cd(II) and As(III, V) (Paper II). 5.6. The PSO kinetic model showed the best-fit to explain the experimental data for Cd(II), As(III) and As(V) sorption by all the KOH modified biochars from SSD (R 2 >0.96) (Table 5.6).

From Figures 5.8(a) and 5.8(b), the sorption kinetic rates for Cd(II) and As(III) were insignificantly affected by the washing steps after the KOH treatment. In comparison with the batch washed BSS-KOH bat , the batch and additional column washed BSS-KOH showed a faster kinetic rate for Cd(II), with the PSO rate constant (k2) of 12.907 (± 0.479) g(µmol h)⁻ 1 , compared to 6.298 (± 0.338) g(µmol h)⁻ 1 for the BSS-KOH bat (Table 5.6, Paper II). A similar observation was found for As(III), but with a slight increase of the k2 values, i.e. from 0.205 (± 0.042) g(µmol h)⁻ 1 to 0.378 (± 0.067) g(µmol h)⁻ 1 with the addition of continuous column washing (Table 5.6, Paper II). The enhanced sorption kinetic rates for Cd(II) and As(III) can be linked to the increment of surface area of the KOH biochar with additional column washing. In this case, the surface area was increased from 3.0 (± 0.1) m 2 g⁻ 1 for the BSS-KOH bat to 7.9 (± 0.1) m 2 g⁻ 1 for the BSS-KOH (Table 5.2). The higher surface pore sites on the biochar can induce more sorption for Cd(II) and As(III) (Bogusz et al., 2017;Wu et al., 2017), and thus improving the kinetic reaction rates.

In contrast, a significant alteration of the sorption kinetics for As(V) was observed on the KOH modified biochars with different biochar washing steps (Figure 5.8(c)).

Compared to the BSS-KOH bat , the sorption kinetics for As(V) were much slower on the BSS-KOH with a 3-time decrease of the k2 value (Figure 5.8(c) and Table 5.6). This is probably due to the sorption conditions at initial pH 5.0 as As(III) remains in neutral form whereas As(V) is anionic. As a result, a significantly higher variation of As(V) sorption kinetics by the KOH biochars was found, compared to the As(III). This is likely due to the competition between the release of DOC and PO4 3 ⁻ from the BSS-KOH bat (section 5.1.3) and the As(V) oxyanion, while no interference of such releasable compounds from the BSS-KOH.

50

Furthermore, the different behaviors among two KOH biochars could be linked to the variations of their properties (e.g. zeta potential and surface area) before and after subsequent column washing (Table 5.2). Even if both the KOH biochars were net negatively charged on their surfaces at pH 5.0 (i.e. initial pH of sorption conditions); however, the BSS-KOH had an improved surface area and a 2-time higher zeta potential (i.e. lower negative charges), compared to the BSS-KOH bat (Table 5.2). Such improved properties of the BSS-KOH could induce more As(V) sorption onto this biochar and thus require more time to reach equilibrium state than the BSS-KOH bat . In addition, the kinetic data obtained from model-fittings can be used to explain the different sorption behaviors among the KOH biochars. From Table 5.6, the kinetic data showed that the As(V) sorption by the BSS-KOH was well described by both the PSO and intraparticle diffusion kinetic models (R 2 >0.99). This suggests that the intraparticle diffusion and/or chemical sorption could be the rate controlling step for the BSS-KOH. Consequently, the As(V) could transfer through diffusion and sorbed on the pore sites of the BSS-KOH and/or via chemical interaction between the functional groups of the BSS-KOH and the As(V) species. However, for the BSS-KOH bat , the PSO model was best fitted to describe the As(V) sorption (R 2 >0.96), indicating that chemical interaction between the As(V) species and the BSS-KOH bat was mainly occurred (Table 5.6).

These findings showed that the biochar washing procedures (batch versus batch and subsequent column washings) slightly affected the sorption kinetics for Cd(II) and As(III) by the KOH modified biochars. However, a much slower sorption of As(V) was observed on the KOH biochar with the additional column washing. This could be described by the changes of biochar properties, particularly surface area and zeta potential, before and after the additional column washing step. In addition, this slower rate of the As(V) by the BSS-KOH can be linked to the best-fit of the kinetic data to both the PSO and intraparticle diffusion model. This indicates that not only a simple chemical reaction but also the intraparticle diffusion model were occurred during the As(V) sorption by the BSS-KOH. Nevertheless, the chemical sorption were considered as the main mechanism for Cd(II) and As(III) sorption by the KOH biochars, in line with the PSO model assumption.

Sorption isotherms for Pb(II), Cd(II) and As(III, V) by biochars

The sorption isotherms for Pb(II), Cd(II) and As(III, V) by the different raw and chemically-modified biochars were studied to investigate the influence of biochar chemical treatments on the maximum sorption capacity (Qm) for these elements (Papers I and II). Due to various oxidation states of metalloids such as arsenic, the study on the evaluation of arsenic speciation during sorption experiments was also studied (Paper III).

The sorption isotherms for Pb(II) by raw biochars from SSD and OFMSWD (i.e. BSS bat and BMSW bat ) and their H2O2 or KOH modified biochars (with batch washing after chemical treatments) were carried out in batch experiments (Paper I). Due to the potential release of organic/inorganic compounds from biochars after chemical treatments as previously reported in section 5.1.3, all biochars were submitted to batch and subsequent continuous column washings to eliminate the releasable ions. After that, the raw and chemically-modified biochars from SSD (with batch and subsequent column washings) were selected to study the influence of biochar chemical treatments on the Cd(II) and As(V) sorption isotherms (Paper II). The sorption experiments for these metal(loid)s were performed at an initial pH of 5.0 and at equilibrium conditions (agitation at 180 rpm for 24 h). For all the sorption tests, two isotherm models, namely Langmuir and Freundlich, were chosen to stimulate the sorption behavior of each metal(loid) element.

Langmuir model showed the best fit to describe the sorption for Pb(II), Cd(II) and As(V) by all biochars (except the Pb(II) sorption by BSS-KOH bat due to non-fitting). This literally suggests the possibility of monolayer sorption of metal(loid)s on the homogeneous biochar surface (Zhao et al., 2015b). For this reason, the sorption capacity and sorption affinity of the biochar towards Pb(II), Cd(II) and As(V) were attained by the maximum sorption capacity (Qm) and Langmuir constant (KL) obtained from Langmuir isotherms. More information on the sorption isotherms for Pb(II), Cd(II) and As(V) were described in detail in the following subparagraphs. and KOH modified biochars with batch washing, respectively. BSS bat , BSS-H2O2 bat , and BSS-KOH bat refer to raw biochar from SSD, and its H2O2 and KOH modified biochars with batch washing, respectively.

Comparing the Qm between raw digestate biochars, an 11-time higher Qm of BMSW bat than that of BSS bat was reported (Table 5.7). However, the BSS bat had a larger KL value, compared to BMSW bat (Table 5.7), indicating a higher affinity of the BSS bat towards Pb(II) than the BMSW bat . These findings highlight that the Pb(II) sorption by biochar was significantly affected by the sources of biochar feedstocks. The differences of Pb(II) sorption capacities among the raw biochars can be linked to the biochar physicochemical properties. From Figure 5.2 and Table 5.1, a stronger intensity of Ocontaining functional groups and a 4-time higher Ca content in the OFMSWD than in the SSD were observed. As a result, more possibility of Pb ions to sorb onto the BMSW bat via the chemical sorption and/or cation exchange mechanisms may take place (Li et al., 2017). A previous study of [START_REF] Cataldo | Pb(II) adsorption by a novel activated carbonalginate composite material. A kinetic and equilibrium study[END_REF] also reported the cation exchange between Pb(II) and Ca-modified biochar. The constant values corresponding to Langmuir and Freundlich sorption models are given in Table 5.8.

From Table 5.7 and Table 5.8, the evolutions of the Langmuir parameters (Qm and KL values) of the biochars towards Pb(II), Cd(II) and As(V) were observed after biochar chemical treatments. The higher KL values for Pb(II) by the modified OFMSWD biochars than its raw biochar were observed, while the modified SSD biochars showed a lower KL value for Pb(II) after chemical treatment (except the BSS-KOH bat due to non-fitting) (Table 5.7, Paper I). Similar to Pb(II), the KL values for Cd(II) were also lower on the modified SSD biochars with batch and subsequent column washings (BSS-H2O2 and BSS-KOH), compared to the raw biochar (BSS) (Table 5.8). However, for As(V), the well-fitting of the Langmuir model for its sorption was found with no significant evolutions of KL values after chemical treatments (Table 5.8). These findings indicated that the chemical treatments induced a higher sorption affinity for Pb(II) on the modified biochars from OFMSWD (Paper I). However, the chemically-modified biochars from SSD showed a lower affinity towards Pb(II) and Cd(II), while the chemical treatments did not modify the affinity of the biochars for As(V) species (Papers I and II). The enhancement of the Qm values for Pb(II) sorption was observed on the biochars after the chemical treatments (Figure 5.9(a, b)). For the OFMSWD based biochars, it was increased from 357.1 µmol g -1 (raw biochar) to 434.7 µmolg g -1 and 500.0 µmol g -1 for BMSW-H2O2 bat and BMSW-KOH bat , respectively (Table 5.7). Similarly, for the SSD based biochars, the Qm values for Pb(II) were improved, from 31.4 µmol g -1 (BSS bat ) to 121.9 µmol g -1 (BSS-H2O2 bat ) (Table 5.7). Nevertheless, for the BSS-KOH bat , due to unexpected low Pb(II) sorption abilities at low equilibrium Pb(II) concentrations (Ce) (0-200 µM) (Figure 5.9(b)), Langmuir model was failed to explain the sorption behavior of this biochar. Therefore, no Qm value of the BSS-KOH bat was reported (Table 5.7). The alteration of Pb(II) sorption could be described by a potential release of DOC from the BSS-KOH bat (as previously discussed in section 5.1.3), which could possibly form soluble ligand complexes with Pb(II) and thus hinders the Pb(II) sorption onto the biochar, particularly at the low Pb(II) concentrations (<200 µM) (Paper I).

Similar to Pb(II), the sorption capacities for Cd(II) and As(V) were also enhanced on the modified SSD biochars (submitted to batch and additional column washings) (Figure 5.10(a, b)). From Table 5.8, the Qm values for Cd(II) and As(V) were improved, respectively, by 15-20 times and by 4-5 times after the H2O2 or KOH treatments. The enhancement of sorption abilities for Pb(II), Cd(II) and As(V) by the chemically-modified SSD biochars can be directly linked to the improved biochar physicochemical properties after chemical treatments (Table 5.2). The higher surface areas of the chemicallymodified SSD biochars, compared to its raw biochar (Table 5.2), are in agreement with their higher sorption capacities towards Pb(II), Cd(II) and As(III, V). These results are consistent with previous published data (Ding et al., 2016;Wu et al., 2017), who reported that the surface area plays an important role in the sorption of metal(loid)s by biochars. Moreover, increases of cation exchange capacities of the SSD biochars after the chemical treatments (Table 5.2) may induce more sorption of positively-charged metals like Pb(II) and Cd(II). Furthermore, a slight increase of pHpzc on the chemicallymodified SSD biochars, compared to the raw biochar, leads to the more positive charges on their surfaces (Table 5.2), and thus enhancing the sorption for anionic As(V) (at pH = 5.0).

Overall, these findings suggest that the biochar treatments with KOH or H2O2 could modify the biochar properties and thus induce more Pb(II) sorption on the modified biochars (Paper I). Nevertheless, a caution on the use of KOH for biochar treatment should be highly concerned to eliminate the releasable compounds from the biochar, and thereby preventing an alteration of the Pb(II) sorption on the isotherm curve.

In this section, the comparisons of the sorption isotherms for Pb(II) by the batch washed BSS-KOH bat and BSS-H2O2 bat versus batch and additional column washed BSS-KOH and BSS-H2O2 were performed (Paper I). This is to observe if the biochar washing steps after the KOH or H2O2 treatments affect the sorption behavior and sorption capacity towards Pb(II). After that, the KOH-modified biochars were selected to further study the influence of biochar washing steps on the sorption isotherms for Cd(II) and As(III, V) (Papers II). As previously mentioned in section 5.3.1, the Langmuir and Freundlich isotherm models were used to describe the sorption behavior for Pb(II), Cd(II) and As(III, V).

For Pb(II), Figure 5.11(a) represents its sorption isotherms on the KOH and H2O2 modified SSD biochars fitted with the Langmuir model. 5.11(d) and 5.11(e) show the comparisons of the Pb(II) sorption isotherms between the H2O2 and KOH modified biochars submitted to batch washing and batch and subsequent column washings, respectively. For Cd(II), As(III) and As(V), the experimental data fitted with the Langmuir model is shown in Figure 5.12(a), 5.12(b) and 5.12(c), respectively. All the constants of Langmuir and Freundlich isotherm models for these metal(loid)s are given in Tables 5.9 and 5.10. BSS-H2O2 bat and BSS-KOH bat refer to H2O2 and KOH modified biochars from sewage sludge digestate (SSD), respectively, with batch washing. BSS-H2O2, and BSS-KOH refer to H2O2 and KOH modified biochars from SSD, respectively, with batch and subsequent column washings.

From Table 5.9 and Table 5.10, the sorption data for Pb(II), Cd(II) and As(III, V) was well fitted with the Langmuir model (R 2 >0.96) for all the modified SSD biochars with both the batch washing and batch and subsequent column washings, except the Pb(II) sorption by the BSS-KOH bat . Nevertheless, the Langmuir model was able to describe the Pb(II) sorption on this KOH-modified biochar with the additional column washing (BSS-KOH) (Table 5.9). This is likely due to the elimination of DOC from the KOH biochar after the subsequent column washing (Figure 5.4(c), section 5.1.3). However, unlike the Pb(II), the release of the DOC from the BSS-KOH bat did not modify the Cd(II) sorption due to its well-fitting of the Langmuir model (Table 5.10) (section 5.1.3). This could be explained by a much higher KL value of the biochar towards Cd(II) (e.g. 0.032 L µmol⁻ 1 for BSS-KOH) than Pb(II)) (e.g. 0.007 L µmol⁻ 1 for BSS-KOH) (Tables 5.9 and 5.10). Consequently, a higher affinity of the biochar for Cd(II) than Pb(II) and/or a weaker reaction between Cd ions and the DOC in solutions could be occurred (Papers I and II). From the Langmuir parameters (Table 5.9), the higher KL values for Pb(II) were observed on the modified biochars with the additional column washing step, for instance, from 0.004 L µmol⁻ 1 for the BSS-H2O2 bat to 0.007 L µmol⁻ 1 for the BSS-H2O2.

Similarly, for Cd(II), the KL value of KOH-modified SSD biochars was also increased after the subsequent washing of biochar (Table 5.10). This suggests that the additional biochar washing step induced a higher sorption affinity of the chemically-modified towards Pb(II) and Cd(II). In contrast, for As(III, V), no significant differences of the KL values were found on the KOH biochars with different biochar washing steps (Table 5.10). This implies that no evolution in the affinity for As species towards the KOHmodified SSD biochars (Paper II). However, the lower KL value of As(V) than As(III) indicates that the As(V) species had less affinity for both the BSS-KOH and BSS-KOH bat , compared to the As(III) (Paper II). Considering the effect of biochar washings on the modified SSD biochars, no significant variations of the Pb(II) sorption were found on the H2O2-modified biochar with additional column washing at low Ce of Pb(II) (<200 µM, Figure 5.11(c)). This is likely due to a small release of DOC from the H2O2 biochars to alter the Pb(II) sorption (Figure 5.4(d), section 5.1.3). However, the Pb(II) sorption capacities were significantly enhanced on the KOH biochar with additional biochar washing (BSS-KOH) at low Ce of Pb(II) (<200 µM, Figure 5.11(b)). Nevertheless, no evolution of the Pb(II) sorption was observed among the KOH biochars with different washing steps at high Ce of Pb(II) (1000-5000 µM, Figure 5.11(a)). Compared to Pb(II), a slightly lower Cd(II) sorption ability was observed on the KOH biochar with the subsequent column washing (Figure 5.12(a)) as the Qm value was decreased from 372 ± 7 µmol g⁻ 1 (BSS-KOH bat ) to 306 ± 7 µmol g⁻ 1 for the BSS-KOH (Table 5.10). These findings suggest that the continuous column washing induced an improvement of the Pb(II) sorption capacity on the KOH-modified biochar at low equilibrium Pb(II) concentrations (<200 µM). However, the sorption for Pb(II) and Cd(II) was not significantly affected by the biochar washing steps, particularly at the high equilibrium concentrations (Papers I and II).

In contrast, a slight improvement of As(III) sorption capacity was observed on the KOH biochar with the subsequent column washing (BSS-KOH), while this biochar had a 10fold higher As(V) sorption, compared to the batch washed BSS-KOH bat (Figure 5.12(b, c)). The strong difference of the sorption capacity for As(V) can be due to a strong release of PO4 3 ⁻ from the BSS-KOH bat (Paper I) that may significantly compete with As(V) and thus hinder the As(V) sorption onto this biochar. It has been previously reported that the presence of PO4 3 ⁻ in solutions could facilitate the mobility of As(V) from the sorbents (Neupane et al., 2014). Due to a potential release of DOC and PO4 3 ⁻ from the BSS-KOH bat as previously observed in section 5.1.3, this could strongly affect the mobilization of As(V) rather than As(III) (Kim et al., 2018) since the DOC and PO4 3 ⁻ are also negatively-charged compounds. Han et al. (2011) and Manning et al. (2002) also reported that the presences of DOC and PO4 3 ⁻ in water could compete with oxyanions of As(V) under oxidizing and aerobic conditions, while As(III) is neutral at pH below 9.2. Furthermore, during the As(III) sorption, the arsenic redox distribution showed that about 70% of As(III) was oxidized to As(V) in the BSS-KOH bat , while a partial oxidation (7%) was found in the BSS-KOH (Paper II).

In summary, this section highlights that the proper biochar washing steps after the chemical treatments are required to eliminate biochar leachable compounds that may alter the sorption for metal(loid)s. From this study (Paper I), the complete washing by using a column reactor was necessary for the alkali modified biochars towards the Pb(II) and As(V) sorption. In addition, during this study, a variation of As species repartition, i.e. oxidation of As(III) into As(V), was observed with the KOH-modified SSD biochars. Therefore, an investigation on the redox distribution of arsenic between As(III) and As(V) during the sorption has been further investigated (Paper III).

The redox distribution of arsenic in the biochar-solution system

The KOH-modified SSD biochars were selected to further understand the influence of chemical modifications on the behavior of arsenic redox distribution in the biocharsolution system. Speciation analyses were performed during the As(III) sorption experiments by chromatography couple to atomic spectrometry (Paper III) which, in contrast to solid-phase techniques such as XANES, required to transfer arsenic from the biochar to a liquid phase. Thus, first of all, an extraction procedure for As(III) and As(V) in the solid biochars was tested and validated. After that, the influence of chemical treatments and washing procedures of biochars on the inorganic arsenic redox distribution during As(III) sorption was investigated (Paper III).

Before analysis, the validation of an extraction procedure for biochars is necessary to allow a quantitative recovery of total As and As species without any conversions between As(III) and As(V). In this study, the BSS-KOH bat was selected regarding its ability to oxidize As(III) (Papers II and III).

The addition of ascorbic acid in the extraction solution for As(III) stabilization was assessed by applying the extraction method to biochars with/without spiking with a known amount of As(III) just before starting the extraction procedure. Results showed that all spiked As(III) was recovered by this extraction procedure (Paper III). The results also suggest that when using only 0.3 M H3PO4, the spiked As(III) was significantly oxidized to As(V) (Paper III). Indeed, the addition of 0.167 µmol of As(III) to the biochar before performing the extraction step resulted in the increases of 0.117 (± 0.009) µmol and 0.047 (± 0.004) µmol for As(III) and As(V), respectively (Paper III). Therefore, about 29% conversion of spiked As(III) to As(V) was observed. Nevertheless, no As(III) oxidation was found during the extraction with H3PO4 and ascorbic acid. The differences of As(III) amounts between before and after spiking was 0.168 (± 0.009) µmol, matching with the As(III) added (0.167 µmol). Thus, the addition of ascorbic acid is essential to preserve As(III) stability during the extraction.

Moreover, the extraction yields of As sorption by the biochars were estimated after the sorption experiment for As(III) by recovering the biochars after the sorption and then submitting them to acid digestion and extraction (Paper III). Table 5.11 shows the comparison of total amount of sorbed arsenic determined after acid digestion or extraction procedure. Results showed that the total sorbed-As by the raw and H2O2 modified biochars were too low to be correctly quantified and thus no calculation of the extraction yields was performed (Table 5.11). For the BSS-KOH bat , the total sorbed-As was quantitatively recovered by the extraction, whereas for BSS-KOH about one fourth could not be extracted with these extraction conditions. The differences of the extraction efficiency may result from different sorption interactions of As from one biochar to the other. Nevertheless, the majority of As was extracted from the biochars and the identification of As species in the extracted solutions still gave an overview of the main As speciation and sorption by the biochars. During the As(III) sorption, the biochar samples were exposed in the solutions with an initial As(III) amount of 1.949 µmol. At the end of experiments, each biochar was recovered then separately submitted to acid digestion and to extraction for the determination of the total sorbed As and for the assessment of the redox evolution of sorbed As, respectively. The exposition solution was also analyzed to determine the remaining As amount and to assess the distribution between As(III) and As(V). The comparison of the As(III), As(V) and total As quantities in the exposition solutions and sorbed-As onto the biochars after the acid digestion and extraction procedure is reported in Table 5.12 and more details are described in the following subsections.

From Table 5.12, no significant differences of the total As between the initial and the final exposition solutions of BSS and BSS-H2O2 were observed, implying that no As was sorbed by these biochars. Nevertheless, the As concentrations from the final exposition solutions of both the BSS-KOH and BSS-KOH bat were 2-3 times decreased, compared to the initial As(III) concentration (Table 5.12). The deduced As concentrations showed that the arsenic was able to sorb onto both the BSS-KOH and BSS-KOH bat with the sorption efficiencies of 67% and 50%, respectively (from the calculations in liquid phase solutions).

Comparison of the arsenic redox distribution in the final exposition solutions showed that almost no As(III) was oxidized to As(V) for both BSS and BSS-H2O2 (Table 5.12).

For the KOH-modified SSD biochars, the assessment of redox distribution only concerned the unsorbed As. In the case of the BSS-KOH, no significant oxidation of As(III) was found, while about 28% of the remaining As was oxidized to As(V) for the BSS-KOH bat . These results are in agreement with the findings from Paper II with a large oxidation (70%) of As(III) in BSS-KOH bat and a partial oxidation (7%) in BSS-KOH during As(III) sorption.

To elucidate if the oxidation was induced by the released compounds from the biochar into the solution or by only the biochar, a control was conducted by replacing the biochar with the release of dissolved compound (RDC) solutions from BSS-KOH bat (Paper III). Figures 5.13(a) and 5.13(b) demonstrate a remaining of arsenic during the As(III) sorption, respectively, onto the BSS-KOH bat and with only the release of dissolved compounds (RDC) from the BSS-KOH bat (a control) as a function of time. During the sorption kinetics of As(III) on the BSS-KOH bat , results showed gradual decreases of As(III), according to reductions of total As over time (Figure 5.13(a)). However, from Figure 5.13(b), no complexation between arsenic and RDC was observed due to almost stable amounts of As(III), As(V) and total As along the time. Considering the As speciation demonstrated a slight oxidation of As(III) in the RDC solutions (9% of the final dissolved arsenic was As(V)) (Figure 5.13(b)), while up to 43% of the final dissolved arsenic was oxidized in the presence of the biochar (Figure 5.13(a)). These findings highlight that the biochar material mainly induced the oxidation of As(III) to As(V) and to a lesser extent by the RDC releasing from the biochar, while no transformation of As(V) to As(III) was found in solutions during the As(III) sorption (Paper III). Considering the As species from extraction of biochars after the sorption, similar distributions between As(III) and As(V) were observed for both the BSS-KOH and BSS-KOH bat (Table 5.12). The majority of arsenic was sorbed as As(III) (90-92%), while only 8-10% was sorbed as As(V) onto the KOH-modified SSD biochars (Table 5.12). The KOH-modified SSD biochars favored the sorption for As(III) rather than for As(V), in agreement with a previous finding (Paper II) on the comparison of the sorption between As(III) and As(V) by these biochars.

Among all the SSD based biochars, the BSS-KOH bat showed a higher oxidation of As(III) to As(V) than other biochars in the final solution (Table 5.12). This is probably due to the ability of this biochar to release the DOC into solution (section 5.1.3). Since the DOC could act as electron acceptors, more oxidation of As(III) to As(V) can be found during the As(III) sorption. A previous study (Dong et al., 2014) also reported that the presence of dissolved organic matter induced more As(III) oxidation.

For the BSS-KOH, the redox distribution of sorbed As was quite similar to the one observed in the final solution (Table 5.12). However, for the BSS-KOH bat , the As species distribution in the exposition solution and on the biochar were different: only 8% of the sorbed As was As(V) (from extraction), while As(V) species represent 28% of As in the final solution. This could be due to the As(V) reduction on the biochar during the sorption process, as also observed by Niazi et al. (2018aNiazi et al. ( , 2018b) ) who studied the As speciation on wood and leaf derived biochars by using the solid phase XANES technique. In the case for the H2O2 modified biochar, the low sorbed As by this biochar resulted in a low accuracy of the As redox distribution and thus should be taken with caution. The percentage of As(V) sorbed onto the BSS-H2O2 (24%) was higher than in the exposition solution (4%) (Table 5.12), suggesting that the As(III) could potentially oxidized to As(V) on the biochar during the sorption.

The results showed that the percentage of As(V) was not the same between the final solutions and the extracted biochars, except for the BSS-KOH. This is likely due to the low sorption ability of both the BSS and BSS-H2O2 for As, which lead to a low accuracy of the As distributions. Nevertheless, for the BSS-KOH, similar distributions of As(V) between the final solution and the extracted biochar can be linked to a complete washing of this biochar to eliminate all the releasable compounds that may disturb the sorption process (section 5.1.3).

The findings also suggest that the biochars could act as electron donors and/or electron acceptors. The reduction of As(V) to As(III) on biochars could be corresponding to the presence of biochar functions such as phenolic or alcoholic (-OH) and carboxyl (-COOH) groups on the biochar surface (Papers I and II). These functional groups could act as electron donors and thus inducing the reduction of As(V) (Choppala et al., 2016). Nevertheless, the oxidation of As(III) to As(V) on the biochar can also occur with the presence of redox active species (e.g. FeO(OH)) on the biochar (Niazi et al., 2018a(Niazi et al., , 2018b)). Based on the biochar properties (Table 5.1), the SSD biochar contained iron (Fe) (~1182 µmol g -1 ) and manganese (Mn) (~14 µmol g -1 ) that could be partially in metal oxide forms. As a result, these metal oxides could promote the redox transformation of As(III) to As(V) on the biochar [START_REF] Gude | As(III) oxidation by MnO2 during groundwater treatment[END_REF]Han et al., 2011;Manning et al., 2002;Vithanage et al., 2017).

Conclusions

This research work illustrates the capability of biochars produced from sewage sludge digestate (SSD) and the organic fraction of municipal solid waste digestate (OFMSWD) towards metals (i.e. Pb(II) and Cd(II)) and metalloids (i.e. As(III, V)) removal from water through the batch sorption kinetic and isotherm studies (Papers I and II). The chemical treatments (KOH or H2O2) followed by the biochar washing procedures (batch washing versus batch and subsequent column washings) were implemented for the first time in this study to improve the sorption performance for Pb(II), Cd(II) and As(III, V) on the chemically-modified digestate biochars (Papers I and II). In addition, the redox state distribution of As species (i.e. As(III) and As(V)) during the As(III) sorption on the solid phase of biochar and in liquid solutions was successfully monitored for the first time in this study (Paper III). The important findings of this thesis are critically summarized as follows:

Significant evolutions of the pH value, electrical conductivity, surface area, cation exchange capacity, and total As content were found after submitting the raw biochars to different chemical treatments and biochar washing procedures (batch washing versus batch and subsequent column washings) (Papers I and II). The additional continuous column washing was able to significantly eliminate the DOC and PO4 3-from modified the biochar, particularly with the KOH treatment.

All the sorption kinetic data for Pb(II), Cd(II) and As(V) on the raw and chemicallymodified biochars was best described by the pseudo-second-order (PSO) kinetic model, literally suggesting that the simple chemical sorption was considered as the rate controlling mechanism for these elements by the biochar (Papers I and II). The KOH or H2O2 treatments strongly induced faster kinetic rates for cationic metals, i.e. Pb(II) and Cd(II), on the modified biochars due to their higher PSO rate constants (k2), compared to the raw biochars both from SSD and OFMSWD (Papers I and II). The improved sorption kinetics for Pb(II) and Cd(II) were linked to the higher surface area and cation exchange capacity of the modified biochar after the chemical treatments. Nevertheless, a slower sorption rate for As(V) oxyanions was observed after treating the SSD biochar with KOH (i.e. BSS-KOH), while no variation was found between the raw and H2O2modified SSD biochars (i.e. BSS and BSS-H2O2, respectively) (Paper II). Different biochar washing procedures after the KOH treatment also induced the changes of the sorption kinetics towards Pb(II), Cd(II) and As(III, V). Compared to the KOH-modified SSD biochar with batch washing (BSS-KOH bat ), a batch and subsequent column washings of this biochar (BSS-KOH) induced faster sorption rates for both Cd(II) and As(III). However, the additional washing of the KOH biochar showed a slower sorption rate towards As(V) species (Paper II). This highlights that both the chemical treatments (KOH or H2O2) and the additional continuous column washing procedures induced faster sorption rates of the modified biochars towards Pb(II), Cd(II) and As(III). However, the KOH modification followed by the column washing procedures induced a slower sorption rate for As(V) on the modified biochar. This is likely due to the wellfittings of the As(V) sorption data with both the PSO and intraparticle diffusion kinetic models, highlighting that the chemisorption and/or the intraparticle diffusion were considered as the rate controlling mechanisms during the sorption for the As(V) species on this biochar (Paper II).

The chemical treatments of biochars significantly affected the sorption isotherms for Pb(II) (Paper I), As(III, V) and Cd(II) (Paper II). The results showed the well-fittings of the Langmuir isotherm model with the Pb(II), Cd(II) and As(III, V) sorption for all the raw and chemically-modified biochars, except the Pb(II) sorption by the batch washed BSS-KOH bat due to non-fitting (Papers I and II). Due to a potential release of the DOC from the BSS-KOH bat into solutions, these compounds could subsequently form solubleligand-complexes with Pb ions and thus altered the Pb(II) sorption behavior, particularly at low equilibrium Pb(II) concentrations (<200 µM). Nevertheless, the Pb(II) sorption was able to follow the Langmuir model on the KOH modified SSD biochar with a subsequent column washing (BSS-KOH) (Paper I). The sorption for As(V) was also significantly influenced by the improper washing of the KOH-modified biochar (BSS-KOH bat ). Due to the ability of this biochar to release PO4 3-into solutions, these releasable ions could compete with As(V) oxyanions and thus hinder the As(V) sorption on the biochar. Therefore, the additional continuous column washing after the biochar chemical treatment, particularly for the KOH, is highly recommended to prevent interferences during the sorption for these metal(loid)s.

The maximum sorption ability (Qm) for Pb(II), Cd(II) was significantly increased on the modified biochars after the chemical treatments (Papers I and II). The enhanced Qm of the chemically-modified biochars towards Pb(II), Cd(II) is probably due to the improved biochar properties, especially the cation exchange capacity, specific surface area and, to a less extent on the point of zero charge of biochar (Papers I and II). Nevertheless, the lower KL Langmuir constant values for Pb(II), Cd(II) were observed after the chemical treatments, indicating a lower sorption affinity for Pb(II) and Cd(II) on the modified SSD biochars. However, the chemical treatments did not modify the As(V) sorption affinity on the modified biochars. The additional biochar washing procedure induced a higher sorption affinity for Pb(II) and Cd(II) but a slightly lower affinity for As species both As(III) and As(V) on the modified biochars.

Arsenic as As(III) was found to be stable or partially oxidized to As(V) with time during the As(III) sorption experiments, depending on the biochar chemical treatments implemented (KOH or H2O2) (Paper III). In addition, a higher oxidation of As(III) to As(V) was observed in the KOH-modified biochar rinsed by a batch washing (Paper III). This oxidation was mainly induced by the biochar components and to a lesser extent by the release of DOC from the biochar (Paper III). The As speciation in the solid phase of different biochars was successfully accomplished by extraction with phosphoric and ascorbic acids followed by chromatographic analysis using liquid chromatography coupled to atomic fluorescence spectroscopy (LC-AFS) (Paper III). Moreover, ascorbic acid plays an important role in preserving As(III) oxidation during the biochar extraction.

The results on the solid-liquid extraction showed that As was mainly sorbed as As(III) (76-92%) by the raw and chemically-modified biochars (Paper III). However, the redox transformation between As(III) and As(V) in the solutions demonstrated that the reduction and/or oxidation may take place, according to different biochar treatments (Paper III). The occurrence of As redox transformation both in the liquid solution and on the biochar was successfully achieved in this study (Paper III). Thus, this allows a direct comparison of arsenic in solid and liquid phases and provide more insights on how to further assess the fate of As in aquatic environments.

Future perspectives

This thesis showed the improved sorption capacities of the chemically-modified biochars from solid digestates for Pb(II), Cd(II) and As(III, V) in single element sorption system. Therefore, further study on the competitive sorption of metal(loid)s by the biochar can be accomplished to stimulate the real sorption behaviors in the environments. In addition, the information about the sorption mechanisms (i.e. surface complexation, ion exchange and precipitation) involved during the sorption for metal(loid)s should be further conducted to highlight the sorption interactions between the biochar and metal(loid)s in the multi-element sorption system. Moreover, the competitive effect of each elements/compounds present in water (e.g. DOC, PO4 3-and CO3 2-) as well as the evolution of pH value can be the aim of future work.

The evaluation of arsenic speciation in biochar during arsenic sorption should be further conducted to provide more understanding of a complex system of arsenic in solutions and on the solid phase of biochars by using chromatographic analysis.

In this study, the chemical treatments (i.e. KOH or H2O2) followed by the continuous washing procedures of the biochar required a long time (~2-3 weeks) for the biochar preparations and a huge amount of water to eliminate the releasable compounds from the chemically-modified biochars. Therefore, alternative modification techniques such as coating the biochars with carbonaceous materials (e.g. graphene oxide) may provide more oxygen-containing functional groups on the biochar surface.

In this thesis, the sorption kinetic and isotherm studies of the Pb(II), Cd(II) and As(III, V) had been performed only in batch experiments for 24 h. Therefore, the sorption of these metal(loid)s in a continuous system could be further performed to test the stability of the biochar for metal(loid) elements in a long-term as well as to quantify the life-time of biochar before reaching a saturation stage.

Introduction

Metal pollution is of very high concerns for human health due to their persistence and toxicity in the environment even in low concentrations. Lead (Pb) has been recognized as one of the most toxic metals in Europe (T oth et al., 2016). Pb pollution often originates from smelters, mining, industrial discharges, car batteries and Pb-based piping for water supply. Discharge of untreated wastewater from the industry may cause an adverse effect to animals and humans. One study reported a case of severe Pb poisoning of children in Haina (Dominican Republic), attributed to a car battery recycling factory (Kaul et al., 1999). Many conventional treatment methods have been developed to decrease Pb levels in contaminated water, including chemical precipitation, coagulation, ion exchange and adsorption (Inyang et al., 2016).

Lead sorption by activated carbon (Cechinel et al., 2014), agricultural waste-derived biochars (e.g. pine wood or rice husk) (Liu and Zhang, 2009), natural zeolite and kaolinite clay (Andrejkovicova et al., 2016;Jiang et al., 2009) have been reported. Recent studies show the potential application of biochar in metalpolluted water treatment due to its high specific surface area and surface properties, e.g. surface charge and hydrophobicity (Liu and Zhang, 2009;[START_REF] Zieli Nska | Effect of sewage sludge properties on the biochar characteristic[END_REF]. Indeed, the solid organic byproduct generated by anaerobic digestion of sludge from wastewater treatment plants has been considered as an alternative source of raw material to manufacture adsorbents (i.e. biochar) for metal removal (Zhang et al., 2013). Biochar is a black solid char derived from the pyrolysis of organic waste materials in a limiting oxygen environment (Inyang et al., 2016). Through the pyrolysis technology, it promotes the recycling of organic waste and supports the environmental sustainability for the community. Biochar has been widely used for many purposes in the environment, such as soil conditioner [START_REF] Saifullah | Biochar application for the remediation of salt-affected soils: challenges and opportunities[END_REF] or filtration medium in wastewater treatment (Mohan et al., 2014). However, there are only few studies on the use of organic by-product sludge biochar to remove metals from water. Biochar produced from organic digested sludge has been used for As(V), Cd(II), Cr(III), Cu(II) and Ni(II) removal from water (Inyang et al., 2012;Jin et al., 2014). A few studies have been dedicated to the Pb sorption by sludge biochars, which can be found in the literature (Table S1).

The main mechanisms involved in Pb sorption onto the biochar are cation exchange, surface complexation, surface precipitation and physical adsorption (Ho et al., 2017;Li et al., 2017). Among these mechanisms, cation exchange of Pb with Ca 2þ and Mg 2þ and is the main contributor to Pb sorption by sludge-based biochar, accounting for 40e52%. Exchanges can also occur to a lesser extend with K þ and Na þ and (<8.5%) (Li et al., 2017), which is in a good agreement with the study of Lu et al. (2012). The surface complexation between Pb and surface functional groups of biochar (e.g. carboxyl and hydroxyl groups) also plays a major role, contributing for about 40% of Pb removal (Li et al., 2017). Surface precipitation can also occur since sludge biochars generally contain high amounts of phosphate (PO 4 3À ) and carbonate (CO 3

2À

) on their surface. Finally, the high surface area of biochar may favor physical adsorption of Pb onto biochar active pore sites (Agrafioti et al., 2014). The relative importance of these sorption mechanisms depends on the biochar feedstock.

Compared to activated carbon, the sorption capacity for Pb by biochar is low and thus numerous modification methods have been applied to improve it. The common treatment methods of biochar are physical activation with steam and chemical treatment with acids, oxidizing agents and alkali solutions (Sizmur et al., 2017). The steam activation of biochar is usually performed at high temperature (>800 C), thus increasing the adsorbent cost which it is not feasible for large scale operation (Wang and Liu, 2018). The chemical modification of biochar is considered as an inexpensive technique since no heat is required during the operation. Treatment of biochar with KOH increases the surface hydroxyl groups and the basicity on the biochar surface (Fan et al., 2016;Li et al., 2014), dissolves ash and condenses organic matter (e.g. lignin and celluloses) in the biochar (Lin et al., 2012;Liou and Wu, 2009;Liu et al., 2012). Modification of biochar with H 2 O 2 was found to increase Oecontaining functional groups, particularly carboxyl groups, on the biochar surfaces (Rajapaksha et al., 2016). Such chemical modifications could induce a leaching of organic matter and mineral ash from biochar pore sites. Thus, after the treatment, several batch washing of the biochar with ultrapure water are required until the pH becomes stable or neutral (Huang et al., 2017;Regmi et al., 2012;Wu et al., 2017). Usually, batch washing of biochar is performed without any concerns on the release of organic or inorganic compounds (e.g. PO 4 3À and CO 3

2À

) from biochar. Since this can influence the sorption of metals by biochar, the effective elimination of these released compounds should be considered. Unfortunately, there is currently a lack of information on the effect of biochar washing on the elimination of the released compounds from biochar after chemical treatment.

This work aims to study the consequences of chemical treatments of biochar and subsequent washing conditions on the Pb sorption capacity. Two chemical reagents (i.e. KOH and H 2 O 2 ) and two washing modes (semi-continuous batch washing as usually performed in the literature and continuous column washing) were applied to raw biochars from sewage sludge and the organic fraction of municipal solid waste. The consequences onto the Pb sorption were evaluated through adsorption kinetic and isotherm studies.

Materials and methods

Feedstocks and biochar preparation

Raw sewage sludge digestate (RSS) and raw organic fraction of municipal solid waste digestate (RMSW) were obtained separately from a wastewater treatment plant and from a solid waste treatment plant located in France. Biochar from RSS was industrially pyrolyzed at 350 C for 15 min using the Biogreen ® technology, while biochar from RMSW was produced at lab scale. Based on the study of Pituello et al. (2014), the RMSW was dried overnight at 65 C to reduce the initial moisture content to less than 10%. After that, it was crushed and sieved into a particle size of 2 mm to separate impurities such as plastic bags, needles and glasses. The RMSWderived biochar was produced at 400 C using porcelain crucibles with lid-cover (Haldenwanger 79 MF, Germany) in a muffle furnace (heating rate 15 C min À1 ) for 1 h. The obtained materials were left to cool down at room temperature with the lids cover.

As recommended by Jin et al. (2014), Wu et al. (2017) and Xue et al. (2012), biochars were washed by semi-continuous means with ultrapure water for 3e4 times (i.e. 2 g of biochar per 200 mL of ultrapure water per washing) until a stable pH was obtained. The RSS-derived biochar produced at 350 C and the RMSW-derived biochar produced at 400 C are named as SS sem and MSW sem , respectively.

Biochar chemical modification

In this study, KOH and H 2 O 2 were selected for the chemical modification of biochar due to the great enhancement of metal sorption (Rajapaksha et al., 2016). To prepare the modified biochar with H 2 O 2 , 2 g of biochar was placed into 20 mL of a 10% H 2 O 2 solution and shaken at 25 (±2) C for 2 h (modified from Xue et al., 2012). Biochar treated with KOH was prepared by mixing 2 g biochar with 100 mL of 2 M KOH solution and was shaken at 25 (±2) C for 2 h (modified from Jin et al., 2014). The total mass of biochar used was around 10 g per each chemical treatment. After chemical modification, biochars were all semi-continuously washed with ultrapure water. Some biochars were submitted to a subsequent continuously washing to study the influence of the washing conditions.

For the semi-continuous washing, the chemically modified biochar was washed in batch by stirring 2 g of biochar in 200 mL of ultrapure water at 20 (±2) C. The batch washing was repeated for 3e4 times until the pH became stable (Jin et al., 2014;Wu et al., 2017;Xue et al., 2012). For the continuous washing, a part of the semi-continuous washed biochar was subsequently washed by continuous circulation of ultrapure water in a glass column (2.8 cm in diameter and 42.5 cm in height). A peristaltic pump (Ismatec Reglo Analog, Model No. ISM827, Ismatec SA Company, Switzerland) was used to maintain the up-flow velocity of 0.77 (±0.01) cm min À1 . Glass beads (2 mm size) were used at the bottom of the column to generate the flow distribution. The column was flushed continuously for at least 70 h at 20 (±2) C with a hydraulic retention time of 6 h.

Once washed, the modified biochars were recovered using VWR filter papers, then dried in an oven at 50 C overnight and further kept in a desiccator prior to use.

Biochar characterization

The pH of the biochar was measured using a pH meter (LPH 330T, Tacussel, France). A suspension of 1 g of sample in 20 mL of deionized water was stirred continuously for 5 min and let to suspend for 15 min.

To quantify the total content of metal(loid)s (i.e. Al, As, Cd, Cr, Cu, Ni, Pb, Mn, Fe and Zn) and cationic macroelements (i.e. Ca, K, Mg and Na) in biochar, an acid digestion was performed using H 2 O 2 (30% w/w, Sigma-Aldrich) and concentrated HCl (37% w/w, Merck) and HNO 3 (65% w/w, Merck) according to EPA Method 3050B (EPA, 1996). The digestion was carried out in a heating block at 95 (±5) C. After filtration with a Whatman grade 1 qualitative filter paper, trace elements were analyzed using inductively coupled plasma optical emission spectroscopy (ICP-OES) (Optima 8300, Perki-nElmer, France).

An X-ray powder diffractometer (AXS D8, Bruker, Germany) was used to identify the crystalline structures of the biochars. Fourier transform infrared (FTIR) spectroscopy (IRAffinity-1, Shimadzu, Japan) was used to identify surface functional groups of biochar with a deuterated-triglycine sulfate (DTGS) detector. The details of the sample preparation for XRD and FTIR analysis are described in the Supplementary Information. Brunauer-Emmett-Teller (BET) specific surface area of chemically-modified biochar samples were determined using the N 2 adsorption method at 77 K (3Flex, Micromeritics, USA). Before BET analysis, the biochar samples were dried at 105 C for 5 h.

Characterization of solution after column washing of biochar

During continuous washing of SS-KOH sem and SS-H 2 O 2 sem , liquid samples were collected at the outlet of the column over time to measure pH (LPH 330T, Tacussel, France), conductivity (CDM 210 conductivity-meter, Radiometer, Copenhagen, Denmark), dissolved organic carbon (DOC), PO 4 3À , CO 3 2À and bicarbonate HCO 3 À . Dissolved organic carbon was determined using a TOC Analyzer (multi N/C ® 2100S, Analytikjena, Germany) according to the standard method from Afnor (Afnor, 1997a). A colorimetric method (Afnor, 1997b) was used to measure PO 4 3À with a spectrophotometer (DR1900, Hach, France) at l 700 nm. For CO 3 2À and HCO 3 À , they were analyzed by a titrimetric method [START_REF] Afnor | Qualit e de l'eau. Tome 2. D etermination de l'alcalinit e[END_REF] using 10 À3 M HCl as a titrant and phenolphthalein and mixed bromocresol green-methyl red as indicators. Microwave Plasma Atomic Emission Spectroscopy (MP-AES) (Agilent 4100, Agilent Technologies Inc., USA) was used to analyze the released concentration of Ca and Mg at l 422.673 nm and l 285.213 nm, respectively.

Sorption experiments

A stock solution of 1000 mg L À1 Pb(II) was prepared from Pb(NO 3 ) 2 . The solution was diluted in the range from 10 to 1000 mg L À1 and the initial solution pH was adjusted to 5.0 (±0.2) by using 0.1 M HNO 3 or 0.1 M NaOH. For all sorption experiments, 100 mg biochar was added to 25 mL of Pb solution and shaken at 20 (±2) C using an orbital shaker (KS 501 digital, IKA™, USA) at 180 rpm. All experiments were conducted in triplicate and the average values are reported. Additional analyses were performed when variations over 10% were observed. Samples were filtered through Whatman polyethersulfone syringe filters (0.2 mm) and analyzed by MP-AES (Agilent 4100, Agilent Technologies Inc., USA) at l 405.781 nm to quantify the remaining Pb concentration in solution.

Adsorption kinetic tests for Pb sorption onto raw and chemically-modified biochars from semi-continuous washing conditions (i.e. MSW sem , MSW-H 2 O 2 sem , MSW-KOH sem , SS sem , SS-H 2 O 2 sem and SS-KOH sem ) were performed to determine the time required to reach sorption equilibrium and the sorption kinetic constants using pseudo-first-order (PFO) and pseudo-second-order (PSO) models (kinetic equations are presented in Supplementary Information). These experiments were carried out with 100 mg L À1 Pb initial concentration, samples being collected after 10 min, 20 min, 30 min, 1 h, 2 h, 5 h, 7 h and 24 h of stirring. Adsorption isotherms were performed with initial Pb concentrations between 10 and 1000 mg L À1 . Two common adsorption isotherm models (Langmuir and Freundlich) were used to fit the experimental data. The details of these adsorption isotherm equations are provided in the Supplementary Information.

Results and discussion

Characterization of raw and chemically modified biochars

The pH of suspensions of the raw and modified biochars is given in Table 1. A higher pH of the biochar compared to the raw feedstock was observed on both the organic fraction of municipal solid waste and sewage sludge digestate samples. The pH of MSW sem and SS sem was 8.9 and 6.5, respectively. This can be due to the presence of alkali compound (i.e. calcite) in the organic fraction of municipal solid waste biochar, while this component was not detected in sewage sludge digestate derived biochars (Fig. 1(S)). Chemical modification with KOH induced a significant shift to higher pH values, while a marginal pH variation was observed after H 2 O 2 treatment (Table 2).

Table 1 shows the content of Ca, Mg, Na and K contained in the raw organic digestates (i.e. RMSW and RSS) and the derived biochars (i.e. MSW sem and SS sem ). It was noticed that higher cationic macroelements were found in MSW sem compared to SS sem , particularly for Ca which was almost 4efold higher in the former than in the latter (Table 1). In addition, the chemical characteristics such as ash and metal(loid) composition of RMSW, MSW sem , RSS and SS sem are reported in Table S2. The XRD characterization demonstrates kaolinite, quartz and calcite were present in the biochars (Fig. S1). The presence of these mineral phases on the biochar surface could facilitate the sorption of metals (Al-Degs et al., 2006;Jiang et al., 2009). Additionally, biochar displays the presence of Al, Fe and Mn (Table 1) that could be partially in metal oxide form, i.e. Al 2 O 3 , Fe 2 O 3 and MnO. Thus, this could also promote interaction between the metal ions and biochar (Agrafioti et al., 2014;Wang et al., 2015).

FTIR spectra of MSW sem and SS sem show variations of the surface functional groups according to different biochar types (Fig. S2(a)).

Table 1

The pH and concentrations of Ca, K, Mg, Na, Al, Fe and Mn in the raw organic fraction of municipal solid waste digestate (RMSW), its derived biochar (MSW sem ) with semicontinuous washing, raw sewage sludge digestate (RSS) and its derived biochar (SS sem ) with semi-continuous washing. After alkali modification, stronger absorbance peaks corresponding to vibration of OeH functional groups (3425 cm À1 ), eCH 3 stretching of long chain aliphatic groups (2924 cm À1 ) and vibration of the CeC skeleton and CeO stretching (1033 cm À1 ) were found, especially on SS-KOH sem (Fig. S2(b)). Similar observations were reported by Petrovic et al. (2016) after alkali treatment of biochars. In addition, the FTIR spectra became more intense after H 2 O 2 modification, particularly at 1640 cm À1 which is assigned to the vibrations of C]O bands (ester) (Fig. S2(b)). Such surface functional groups can take part in the sorption mechanisms for metals (Chen et al., 2011;Pituello et al., 2014).

Biochar pH in water

Ca ðg kg À1 Þ K ðg kg À1 Þ Mg ðg kg À1 Þ Na ðg kg À1 Þ Al ðg kg À1 Þ Fe ðg kg À1 Þ Mn ðmg kg À1 Þ RMSW 8.0 ± 0
From Table 2, the BET specific surface areas of SS-H 2 O 2 sem and SS-KOH sem were 3.7 and 3.0 m 2 g À1 , respectively. These values increased after continuous washing with ultrapure water, probably due to the elimination of the ash content or organic matter from biochar pore sites.

Effect of biochar washing after chemical modification

After several batch-washings (semi-continuous mode) of KOH and H 2 O 2 etreated sewage sludge biochar, a subsequent column washing (continuous mode) was performed to highlight possible releases of organic or inorganic compounds from the biochar. , respectively, along the column washing time. During the initial stage of washing (0e5 h), increases in pH were noticed for both KOH and H 2 O 2 etreated sewage sludge biochar. However, after 23 h, the pH of SS-KOH sem started to decrease steadily before it became stable at pH 8.4 (after 64 h). The pH values remained almost unchanged at 6.5 (±0.1) for SS-H 2 O 2 sem beyond 23 h. On the other hand, decreasing trends of conductivities throughout the manipulation were observed on both chemicallymodified biochars (Fig. 1 (a,b)). The values dramatically dropped from 309 to 391 mS cm À1 (at the beginning) to 4e6 mS cm À1 (at the end of operation), which was comparable to the conductivity of ultrapure water from the inlet flow (~3 mS cm À1 ). À and DOC concentrations occurred continuously throughout the period of biochar washing (Fig. 1 (c,d)). Such evolutions are consistent with the conductivity trends (Fig. 1(a,b,c,d)).

For instance, the high release of PO 4 3À (55 mg L À1 ), CO 3 2À (99 mg L À1 ) and HCO 3 À (1418 mg L À1 ) from KOH-treated biochar was related to the high conductivity (391 mS cm À1 ) (Fig. 1(a,c)) at the initial washing stage. Declining trends of the released compounds were observed along the lower conductivity level (Fig. 1(a,c)). Similarly, SS-H 2 O 2 sem showed the same behavior as SS-KOH sem , but with much lower PO 4

3À

, CO 3 2À , HCO 3 À and DOC concentrations (Fig. 1 (b,d)). By comparing KOH and H 2 O 2 treatment, it appears that after KOH treatment the biochar released more organic compounds, PO 4 3À and HCO 3 À /CO 3 2À compared to the H 2 O 2 treatment. The DOC released was about 10-fold higher for SS-KOH sem than for SS-H 2 O 2 sem (at 1 h) (Fig. 1(c andd)). This is likely due to the property of KOH to modify the biochar surface. A similar behavior could be observed on the FTIR spectra, with a significant increase of transmittance for SS biochar treated with KOH (Fig. S2(b,c)). These results clearly illustrate that a semi-continuous washing of chemically-treated biochar up to a stable pH was not sufficient to prevent the release of organic matter and ions. Indeed, in the batch reactor an equilibrium of released elements was achieved and the pH is usually used as an indicator for biochar washing (Huang et al., 2017;Regmi et al., 2012;Wu et al., 2017). However, to directly control the release of ions such as PO 4 3À and HCO 3 À /CO 3 2À from the biochar, conductivity is proposed as a more accurate parameter rather than the pH. Therefore, a continuous washing with conductivity monitoring is recommended.

Adsorption kinetics

Fig. 2 shows the effect of contact time on the Pb removal capacity of MSW sem , MSW-H 2 O 2 sem , MSW-KOH sem , SS sem , SS-H 2 O 2 sem and SS-KOH sem . The time required to reach the equilibrium state was different for each biochar type (Fig. 2). For sewage sludge based biochars, the equilibrium time was achieved after 5 h for raw SS sem , 7 h for SS-H 2 O 2 sem and 2 h for SS-KOH sem , while MSWebased biochars showed faster kinetics compared to SS biochars in which the equilibrium time was reached within 5 h for MSW sem , 2 h for MSW-KOH sem and 1 h for MSW-H 2 O 2 sem . The data of the kinetic experiments were fitted using pseudofirst-order and pseudo-second-order equations. The corresponding kinetic parameters are given in Table S3. The pseudo-second order kinetic model was the most suitable to fit the experimental data for all biochars (Table S3). Considering the parameter k 2 , it could be concluded that MSW-H 2 O 2 sem exhibited the fastest kinetic rate followed by MSW-KOH sem , SS-KOH sem , MSW sem , SS sem and SS-H 2 O 2 sem (Table S3). In addition, the equilibrium adsorption capacities determined by the second order model (Q e ) were in agreement with the experimental ones (Q e,exp ) (Table S3).

Comparison of the different chemical modifications for biochar indicated that the alkali treatment (KOH) could induce a faster Pb removal rate, especially for the organic fraction of municipal solid waste (MSW) biochar. This was supported by a previous study from Regmi et al. (2012) on a switchgrass biochar treated with alkali reagent. In contrast, biochar modified with an oxidizing agent (H 2 O 2 ) may not always increase the kinetic rate for Pb sorption. In this case, SS-H 2 O 2 sem reached equilibrium slightly slower than SS sem .

Different kinetic rates for Pb adsorption were observed when comparing SS sem and MSW sem , the former requiring a longer time to reach sorption equilibrium. This was mainly due to differences in biochar properties resulting from the different origin of these biochars. FTIR spectra show stronger peaks of Oecontaining functional groups in MSW sem compared to SS sem especially at 1033 cm À1 (Fig. S2(a)), which could induce a faster kinetic rate for metals (Bogusz et al., 2017). In addition, MSW sem biochar contained calcite on its surface (Fig. S1(a)) and had higher cationic macroelements (i.e. Ca, K, Mg and Na) content than SS sem biochar (Table 1) which could promote cationic metal (i.e. Pb) binding onto MSW sem biochar. This is consistent with the study of Zhang et al. (2017) on the use of celery biochars with high amounts of alkaline minerals for cationic metal removal from aqueous solutions. However, the physical properties of biochar like porosity and specific surface area should be also considered as they play a role during the adsorption process. The results from Fig. 3(a) demonstrate that chemical treatment of biochar could enhance the Pb sorption capacity from 6.5 mg g À1 on SS sem to 25.1 mg g À1 on SS-H 2 O 2 sem . However, it was completely different for SS-KOH sem due to unexpected low Pb adsorption capacities at low equilibrium Pb concentration (0e50 mg L À1 ). As a result, the Langmuir model failed to describe the behavior of this sorption curve. One hypothesis is that a large amount of organic matter released from SS-KOH sem (Fig. 1(c)) inhibits the Pb sorption onto the biochar by forming soluble Pb-ligand complexes, resulting in lower Q e values at the low C e range (0e50 mg L À1 ).

Adsorption isotherms

Fig. 3(b) shows that the Pb sorption capability was improved by KOH and H 2 O 2 treatment of MSW biochar. It increased from 72.9 mg g À1 for MSW sem to 90.0 mg g À1 and 106.3 mg g À1 for MSW-H 2 O 2 sem and MSW-KOH sem , respectively. This implies that chemical treatment could be effectively used to promote Pb sorption onto the organic fraction of municipal solid waste-based biochar. Fig. 4(a) shows the Pb adsorption isotherms on sewage sludge digestateebased biochars fitted with the Langmuir model (except KOH sem due to non-fitting). Fig. 4(b) and (c) compare the adsorption isotherms obtained for two different washing systems: batch semi-continuous versus batch followed by column continuous washing. From Fig. 4(b), a significant enhancement of Pb sorption by KOH con (continuous washed) was observed compared to KOH sem (semi-continuous washed) particularly at low C e (0e50 mg L À1 , highlighting the importance of biochar washing after chemical treatment. These results are consistent with the hypothesis of Pb sorption inhibition due to the formation of soluble complexes with organic matter released by the biochar. Such compounds were removed during the subsequent column washing, leading to a higher sorption capacity for KOH con compared to KOH sem , especially at low Pb equilibrium concentrations. In addition, the Langmuir adsorption model was then able to fit with experimental data of KOH con (Table 3). However, no significant differences in Pb sorption capacities between KOH sem and KOH con were observed at higher C e (400e1000 mg L À1 ) (Fig. 4(a)). This implies that the presence of organic matter in aqueous solutions significantly affects the Pb sorption onto biochar only at low Pb ion concentrations, especially <50 mg L À1 . In contrast, similar adsorption capacities for Pb by both H 2 O 2 sem and H 2 O 2 con were achieved, even at low Pb equilibrium concentrations (0e30 mg L À1 , Fig. 4(c)). This is because H 2 O 2 treatment induced a small release of inorganic and organic matter from the modified biochar. As a result, batch washing was sufficient and no improvement was obtained with the subsequent column washing. Fig. 4(d) and (e) compare the adsorption isotherms at low C e for Pb by H 2 O 2 and KOH treated sewage sludge biochars on semicontinuous washing and continuous washing, respectively. The conclusions totally differ according to the washing procedure: the semi-continuous washing indicates that sorption is more important after H 2 O 2 (Fig. 4(d)), whereas the continuous washing leads to the opposite conclusion (Fig. 4(e)). This is the consequence of the uncompleted efficiency of batch washing to remove the released compounds for biochar after KOH treatment.

Q m ðmg g À1 Þ K L ðL mg À1 Þ R 2 K F ðmg g À1 ÞðL mg À1 Þ 1=n n R 2
Considering the maximal sorption capacity (Q m ) of each biochar (Table 3), chemical treatments of biochar with KOH and H 2 O 2 show an improved Pb sorption capacity, particularly on sewage sludge digestate biochar. Results show an enhanced Q m for Pb by almost 2efold for SS-H 2 O 2 con and by 12efold for SS-KOH con with respect to the unmodified SS biochar. The use of KOH for chemical treatment thus appears more efficient than H 2 O 2 . However, a proper biochar washing procedure after chemical treatment is highly required to prevent the altered isotherm shape.

Possible Pb sorption mechanisms

Previous studies have illustrated the correlation of sorption for Cu 2þ , Cd 2þ and Pb 2þ with the oxygen-contained functional groups obtained from FTIR analysis (Petrovic et al., 2016;Regmi et al., 2012). The interactions between Pb and available surface functional groups on biochar can occur through surface complexation. In addition, possible cation exchange may take place. This is consistent with the observed release of Ca 2þ and Mg 2þ in solution after Pb sorption (Fig. 5).

Fig. 5(a) compares the cations ( Ca 2þ and Mg 2þ ) released by KOH-modified sewage sludge biochars from semi-continuous and continuous washing. The results show that the concentration of Ca 2þ and Mg 2þ released from SS-KOH sem was relatively low (<3 mg L À1 ) at initial Pb concentrations between 0e203 mg L À1 . This was in correlation to a low sorption capacity for Pb by SS-KOH sem (see section 3.4). In contrast, a huge release of Ca 2þ and Mg 2þ by SS-KOH con was observed at adsorption equilibrium (Fig. 5(a)). The released Ca 2þ and Mg 2þ from this biochar was related to the amount of Pb adsorbed onto the biochar surface (on molecular mass basis) (Fig. 5(b)). These findings are in accordance with previous studies of biochars produced from grape pomace and water hyacinth for Pb and Cd removal, respectively (Petrovic et al., 2016;[START_REF] Zhang | Efficiency and mechanisms of Cd removal from aqueous solution by biochar derived from water hyacinth (Eichornia crassipes)[END_REF].

Moreover, the characterization of biochar shows the presence of a mineral phase that could also be involved in Pb sorption (Fig. S1). Nevertheless, previous links between the release of cations from biochar and the Pb sorption capacity show that sorption of Pb by the mineral phase in the biochar is not the main mechanism involved in sorption. In addition, Pb precipitation onto sludge biochar may contribute to the sorption mechanism. The results from the simulation shows that at initial Pb concentrations of 10 and 970 mg L À1 (Fig. S3), Pb precipitation onto biochar will likely occur when the final pH values are above 6.0 and above 5.0, respectively. Since the final pH in the solutions of both SS-KOH sem and MSW-KOH sem were above these values, Pb precipitation partly occurred on both biochar types during the sorption experiments.

Practical implications

The conversion of organic digested sludge obtained from wastewater treatment plants through pyrolysis is considered as a promising approach for sludge management. This corresponds to a huge reduction of the sludge quantity as well as a decrease of metal mobility in the sludge. It also promotes the environmental sustainability strategy due to recycling of the organic wastes and adding more value to the obtained products (i.e. biochar).

Chemical modification of biochar has been suggested as an economical technique to improve the sorption performance for Pb. From this study, treatment of biochar with chemical reagents affected the sorption kinetics for Pb. In this case, KOH could induce faster Pb removal rates compared to the unmodified biochars, while it was not systematic for biochars treated with H 2 O 2 . Furthermore, the chemically-modified biochars reveal a significant Pb sorption enhancement, but with a caution on proper-washed biochar especially with KOH treatment. Indeed, we observed that this treatment led to an inhibition of Pb sorption at low Pb initial concentration linked to the release of complexing compounds (e.g. phosphate, (bi)carbonate and organic matter). Thus, a sufficient biochar washing is required to remove these compounds prior to practical applications to ensure an improvement of the Pb sorption features and avoid the contamination of water bodies by the released products. Future research on the effect of biochar washing after chemical treatment is required to better understand the role of released organic compounds and other alkali anion ions on Pb sorption by sludge-based biochar.

Conclusions

This study demonstrated the capability of raw digested sludge biochars and chemically-modified biochars in Pb removal from water through adsorption kinetic and isotherm studies. Lead sorption data could be well-described by the Langmuir isotherm model and followed the pseudo-second-order kinetic model. The KOH treatment was found more efficient than the H 2 O 2 treatment to improve Pb sorption, with an improvement of both the sorption kinetics and capacity. In the case of sewage sludge biochar, the sorption capacity was about 2efold and 10efold higher after H 2 O 2 and KOH treatment, respectively. However, special attention should be paid to the washing procedure after chemical modification to avoid an inhibition of metal sorption due to a significant release of soluble compounds from an improperly washed modified biochar.

XRD and FTIR analysis

X-ray diffractometry (XRD) was employed to identify the crystalline structures presented in raw and modified biochars. The samples were grounded to less than 100 µm particle size and were characterized using a powder diffractometer (AXS D8, Bruker, Germany) with Cu Kα radiation at 1.54 Å wavelength, over the 2θ range from 10° to 32°, for 10 seconds per step, at 40 kV voltage and 40 mA current with a Solx (Si-Li) detector. The crystalline compounds present in biochar were identified by search-match software with reference from the American Mineralogist Crystal Structure Database.

Fourier transform infrared spectroscopy (FTIR) were recorded to identify the functional groups present on biochar surfaces. Each sample was mixed with KBr in a ratio of 1 mg sample (particle size 100 µm) per 200 mg KBr and the pellet was prepared by compression under vacuum (Jouraiphy et al., 2005). The analysis was performed using a Shimadzu IRAffinity-1 Spectrometer with a deuterated-triglycine sulfate (DTGS) detector. The number of scans was 12 with a resolution of 16 cm -1 and a frequency in a range from 4000 to 800 cm -1 .

with slight shifts on the XRD spectra. For instance, the XRD peak of CaCO3 was shifted from 29.36° (MSW sem ) to 29.23° (MSW-KOH sem ) and 29.20° (MSW-H2O2 sem ). On the other hand, only Al2Si2O5(OH)4 and SiO2 were detected in SS sem , SS-KOH sem and SS-H2O2 sem (Fig. S1(b)).

The results showed no significant differences of XRD spectra between the raw and chemical treated sewage sludge biochars. Kaolinite and carbonates present in the biochar can be partly involved in adsorption of Pb onto the biochar surface. The peak band of 3425 cm⁻ 1 referred to the stretching vibration of O-H functional groups (H-bonded) of alcohols and phenols and was present in all biochars (Fig. S2(a)). The bands at 2924 cm⁻ 1 was due to the vibration of C-H stretching of long chain aliphatic groups (Chen et al., 2011) and was also found in all biochars (Fig. S2(a)). The bands of 1660-1600 cm⁻ 1 were assigned to the stretching vibrations of C=C and were observed only in SS sem , SS-KOH sem and SS-H2O2 sem (Fig. S2(b)). Since sewage sludge digestate is generally composed of cellulose, hemicellulose and lignin (Jouraiphy et al., 2005), thus the spectra of the amides band are broad bands of 1690-1640 cm⁻ 1 . The broad bands between 1442-1384 cm⁻ 1 was assigned to -CH2 scissoring and were observed in raw and modified biochars from sewage sludge (Fig.

FTIR spectra of biochars

S2(b))

. A strong intensity of -CH2 scissoring was found at a peak of 1427 cm⁻ 1 in MSW sem , MSW-KOH sem and MSW-H2O2 sem (Fig. S2(c)). The stretching of carboxylate groups and C-O of alcohols appeared in the sample between 1300 and 1000 cm⁻ 1 . The peaks at 1033 cm⁻ 1 were assigned to C-C skeleton vibration and C-O stretching vibration and were found in all biochars (Fig. S2(a,b,c)). The FTIR spectra provide the qualitative information of functional groups present on the surfaces of biochars. The results demonstrated the differences in surface functional groups between MSW sem and SS sem corresponding to different feedstock type and pyrolysis temperature. From Fig. S2(b), it is obvious that there were similar FTIR spectra between SS sem , SS-KOH sem , SS-H2O2 sem with a small shift at 1635 cm -1 (SS sem ) to 1597 cm⁻ 1 (SS-KOH sem ).

Adsorption kinetics

The simulation of the sorption kinetics for Pb(II) by raw and modified biochars was performed using pseudo-first-order (Eq. 1) and pseudo-second-order (Eq. 2) kinetic models, which are illustrated as followed:

log(𝑄 𝑒 -𝑄 𝑡 ) = log𝑄 𝑒 - 𝑘 1 𝑡 2.303 (1) 𝑡 𝑄 𝑡 = 1 𝑘 2 𝑄 𝑒 2 + 𝑡 𝑄 𝑒 (2)
where Qt and Qe are Pb(II) adsorption capacity (mg g⁻ 1 ) at time t (h) and equilibrium, k1 (h⁻ 1 ) and k2 (g(mg h)⁻ 1 ) are the rate constants for pseudo-first-order and pseudo-second-order models, respectively (Liu & Zhang, 2009). The pseudo-first-order curves were plotted via log(Qe -Qt) versus time t in which the values of k1 can be obtained from the slope corresponding to Eq. ( 1).

The pseudo-second-order model was also applied to test the experimental data from the plot of t/Qt against t, from which Qe and k2 values can be calculated from the slope and interception of this plot, respectively.

Adsorption isotherms

Langmuir and Freundlich models are presented in equation ( 3) and ( 4), respectively:

𝐶 𝑒 𝑄 𝑒 = 1 𝐾 𝐿 𝑄 𝑚 + 𝐶 𝑒 𝑄 𝑚 (3) ln𝑄 𝑒 = ln𝐾 𝐹 + 1 𝑛 ln𝐶 𝑒 ( 4
)
where Ce is the equilibrium Pb solution concentration (mg L⁻ 1 ), Qe and Qm are the Pb(II) adsorption capacity at equilibrium and at maximum capacity (mg g⁻ 1 ), n is the Freundlich constant which indicates the favorability of adsorption, and KL and KF are the adsorption constants for the Langmuir and Freundlich isotherm models, respectively. c dl refers to below detection limit. d Values reported as means (n=3) followed by standard deviation.

Evolution of pH during Pb adsorption isotherms

Table S2 shows the metal(loid) contents of RMSW, MSW sem , RSS and SS sem . Metals such as Cr, Cu, Pb and Zn were retained in the biochars during pyrolysis, while As (metalloid) was volatilized at the temperature below 400 °C (Table S2), this was also affirmed by previous studies from Duan et al. (2017). As expected, Cu and Zn were the main components in both the solid digestates and biochars (Table S2), which was in correlation to the studies from Pituello et al. (2014). Additionally, no metals exceeded the limit values of metal(loid)s present in the sludge based on the French guidelines (Directive 86/278/EEC).

Table S3.

Parameters of pseudo-first-order and pseudo-second-order kinetics for lead sorption by SS sem , SS-H2O2 a NF refers to no satisfactory fit by the model.

Introduction

Arsenic (As) and cadmium (Cd) contamination in water creates a pressing toxicity towards aquatic organisms. Arsenic is mostly present as two inorganic species, arsenite (As(III)) and arsenate (As(V)), whereas a divalent cadmium ion (Cd(II)) dominantly prevails in water. Due to the toxicity of As and Cd, numerous treatments have been dedicated to removing these elements from polluted water. Among them, sorption is a cost-effective and simple treatment for metal(loid) removal from water (Ding et al., 2016;Ofomaja et al., 2014;Peng et al., 2018;Wang et al., 2016;Zhou et al., 2017). In recent years, biochar, a black solid char obtained from pyrolysis (> 250 °C) of biomass under oxygen-limiting conditions (Ahmad et al., 2014), has been considered as a potential alternative to activate charcoal in sorption treatments. Sewage sludge digestate (SSD) obtained from wastewater treatment plants is currently considered as a promising material to produce a lowcost biochar (Agrafioti et al., 2013;[START_REF] Rajapaksha | Pyrolysis condition affected sulfamethazine sorption by tea waste biochars[END_REF]Yuan et al., 2015;Zielińska and Oleszczuk, 2015).

During the pyrolysis step of biochar preparation, the organic matter from biomass undergoes structural modifications such as oxidation and conversion of aliphatic forms to aromatic forms (Zama et al., 2017). The resulting biochar generally harbors negative and positives charges on its surface depending on the neighboring pH. Consequently, at circumneutral pH, where negative charges are predominant, it has been applied to sorb cationic metals like lead (Pb), zinc (Zn) and cadmium (Cd) from polluted water (Ding et al., 2016;Ho et al., 2017;Park et al., 2017;Wongrod et al., 2018). Unfortunately, the sorption ability of SSD biochar for metal(loid)s, particularly As oxyanions, is relatively low compared to activated carbon or paper mill sludge biochar (Table S1, supplementary information). Therefore, biochar modifications like physical activation and chemical treatment have been applied to improve its sorption ability (Li et al., 2017;Rajapaksha et al., 2016;Sizmur et al., 2017). Even though there are many studies on the adsorption capacity of raw and modified biochars [START_REF] Liu | Enhancement of As(V) adsorption from aqueous solution by a magnetic chitosan/biochar composite[END_REF][START_REF] Zuo | Enhanced removal of Cd(II) from aqueous solution using CaCO 3 nanoparticle modified sewage sludge biochar[END_REF], an evaluation of the modification of the biochar properties is yet to be done. Indeed, the biochar surface properties are important parameters that influence the behavior of a sorbent to sorb anionic and cationic pollutants.

The present work aimed to study the consequences of chemical treatments of SSD biochar using KOH or H 2 O 2 on its arsenic and cadmium sorption capacities. Batch experiments were performed to evaluate the sorption kinetics and capacity of raw and modified biochars for As(III, V) and Cd(II). Several surface properties, such as the pH of point of zero charge (pH PZC ), the cation exchange capacity (CEC) and the Brunauer-Emmett-Teller (BET) surface area were determined to access information on, respectively, the net surface charge, cation exchange capacity and available specific surface area of the SSD biochars. These parameters can be used to discuss the sorption characteristics of the biochar. The influence of chemical modifications of the biochar onto sorption of redox-sensitive elements was also investigated through the determination of arsenic redox transformation (i.e. As(III) and As(V)) during sorption experiments. Finally, the consequences of biochar washing on all these parameters were considered by comparing the results obtained with a KOH-modified biochar submitted either to only a batch washing or to a subsequent continuous column washing.

Materials and methods

Chemical reagents

Cadmium (Cd(NO 3 ) 2 .4H 2 O, 99% w/w, Merck), arsenate (Na 2 HAsO 4 .7H 2 O, 98% w/w, Sigma-Aldrich) and arsenite (AsNaO 2 , 98% w/w, Merck) were used to prepare 10 mM of Cd(II), As(V) and As (III) stock solutions, respectively. The ultrapure water (18.2 MΩ, MilliQ Gradient A10, Millipore SAS 67120, Molsheim, France) was used to wash biochars, as well as to prepare and dilute all solutions and samples.

Biochar preparation and chemical treatment

Pellets of domestic sewage sludge digestate (SSD) were obtained from a wastewater treatment plant (WWTP) located in Limoges, France (size: 285,000 equivalent population). The main treatment processes of the WWTP include activated sludges with anoxic treatment and addition of iron(III) chloride (FeCl 3 ) to, respectively, remove nitrogen(N) and phosphorus (P). Particularly, the SSD sample was collected in the storage tank after dewatering and drying.

The sample was heated at 350 °C for 15 min under slow pyrolysis technology to produce SSD biochar and more information was described in our previous work (Wongrod et al., 2018). The SSD biochar was chemically treated for 2 h with a 10% H 2 O 2 solution (100 g: 1 l) or with 2 M KOH solution (100 g: 2.5 l) (Wongrod et al., 2018).

The raw and chemically-modified biochars were subsequently washed as follows: a triple batch washing in a row followed by a continuous column washing (Wongrod et al., 2018). These washing steps were performed to remove competitive ions (e.g. PO , Ca 2+ and Mg 2+ ) and organic matter released by the biochar during the chemical treatment. These released compounds can react with elements and thus interfere with the biochar-metal(loid) sorption system. In addition, clogging of the biochar pore sites by small particles can be eliminated through these washing steps. To determine if the washing procedure influences the metal(loid) sorption, a KOH-modified biochar was washed only with a batch procedure. The raw SSD biochar, its H 2 O 2 and KOH-modified SSD biochar are denoted as, respectively, BSS, BH 2 O 2 and BKOH. The KOH-modified biochar washed with only the batch procedure is labeled as BKOH bat .

Biochar characterization

X-ray powder diffraction (XRD) and Fourier transform infrared spectroscopy (FTIR) of the SSD and its derived biochar were analyzed to observe the changes of, respectively, crystalline structures and surface functional groups after biochar production. The total amount of metal and metalloid in the SSD and its derived biochar was determined using acid digestion from the EPA Method 3050B (EPA, 1996). The details of the digestion procedures are described by Wongrod et al. (2018).

The pH (pH-meter LPH 330T, Tacussel, France) and electrical conductivity (EC) (CDM 210 conductivity-meter, Radiometer, Denmark) of the biochar were measured in water. The surface area of biochar was measured using N 2 sorption by the Brunauer-Emmett-Teller (BET) method at 77 K (3Flex, Micromeritics, USA). Pretreatment of biochar was performed by drying at 105 °C for 5 h prior to the BET surface area (S BET ) analysis.

The pH of point of zero charge (pH PZC ) of the biochar was determined from zetametry by determining the zeta potential at different pH ranges (Mahmood et al., 2011). For the determination of the cation exchange capacity (CEC), 2 g of biochar (dry weight) were mixed with 40 ml of 0.05 N cobalt hexamine trichloride ([Co(NH 3 ) 6 ]Cl 3 ) solution (99% w/w, Sigma-Aldrich) in polypropylene tubes (Aran et al., 2008). The details of the biochar preparation for pH PZC and CEC measurement are provided in the supplementary information.

Sorption experiments

In sorption kinetic experiments, 10 μM As(III), As(V) and Cd(II) were prepared separately from stock solutions of 1 mM As(III), As(V) and Cd(II) (as AsNaO 2 , NaHAsO 4 .7H 2 O and Cd(NO 3 ) 2 , respectively). Raw and modified biochars (600 mg) were added to 150 ml of metal (loid) solution. The initial pH was adjusted to 5 using HNO 3 and NaOH. The solutions were agitated at 20 ( ± 2) °C at 180 rpm for 24 h using an orbital shaker (KS 501 digital, IKA™, USA). The sorption experiments were performed in duplicate. Samples of solution were collected over time. The total collected volume was less than 5% of the initial volume to prevent any alteration of the sorption equilibrium. Sampled solutions were filtered through Lab Logistics Group (LLG) cellulose acetate syringe filters (0.45 μm) and their As and Cd concentrations were determined by graphite furnace atomic absorption spectrometry (240Z, Agilent Technologies, USA) at λ 193.7 nm and λ 228.8 nm for As and Cd, respectively. The pseudo-first-order (PFO) and the pseudo-secondorder (PSO) kinetics models were used to describe the experimental data. The intraparticle diffusion model was also studied to access information on the mechanism of the sorption processes (Zhao et al., 2015a).

Adsorption isotherms were performed with initial concentrations ranging from 10 to 300 μM for As(III, V) and from 10 to 4000 μM for Cd (II). After 24 h of agitation, the residual concentrations of As were determined by graphite furnace atomic absorption spectrometry as detailed previously, whereas the residual concentrations of Cd were determined by Microwave Plasma Atomic Emission Spectroscopy (MP-AES) (Agilent 4100, USA) at λ 228.8 nm. Two common adsorption isotherm models (Langmuir and Freundlich) were used to fit the experimental data. Arsenic speciation was determined by High-Performance Liquid Chromatography (HPLC) coupled to Atomic Fluorescence Spectrometry (AFS) with Hydride Generation (HG) (PS Analytical Millennium Excalibur, P S Analytical, UK). For the chromatographic conditions, a Hamilton PRP-X100 column was used with a phosphate buffer solution as the mobile phase at a rate of 1 ml min -1 . Detailed information can be found in Wan et al. (2014).

Statistical analysis

Results are reported as means ± standard deviation (n = 3) unless otherwise stated. Uncertainties of sorption parameters were obtained by non-linear regression using Statistica software (v6.1, StatSoft). t-test with two-tailed distribution at a statistical significance level of P ≤ 0.01 were used for data comparison.

Results and discussion

Influence of chemical treatment on biochar characteristics

The XRD analysis showed that quartz was present in both SSD and BSS with insignificant modification of the surface properties after biochar pyrolysis (Fig. S1). FTIR spectra demonstrated similar binding between atoms of both the SSD and BSS (Fig. S2). The absorbance peaks of hydroxyl (-OH) groups (3425 cm -1 ), -CH 3 stretching of aliphatic groups (2924 cm -1 ), C=C stretching vibrations (1600 cm -1 ), and C-C skeleton and C-O stretching (1033 cm -1 ) were mainly detected on the SSD and BSS (Fig. S2). These functional groups can be involved in the sorption for metal(loid)s (Pituello et al., 2014).

The pH values of the suspensions of the raw and chemically-modified biochars are given in Table 1. The KOH treatment induced a change from 6.4 ( ± 0.1) for BSS to 8.4 ( ± 0.1) for BKOH and 10.1 ( ± 0.1) for BKOH bat , whereas no significant difference (P ≤ 0.01) was observed after H 2 O 2 treatment (Table 1). Results also showed no significant changes of the electrical conductivity (EC) among each biochar suspension (i.e. BSS, BH 2 O 2 and BKOH), except for BKOH bat where a 50fold higher value 324 ( ± 2) μS cm -1 was observed. The high EC value of BKOH bat is probably due to a significant release of weakly sorbed ions (e.g. PO 4

3-, HCO 3 -, CO 3 2-, Ca 2+ and Mg 2+ ) from BKOH bat as previously reported by Wongrod et al. (2018). A previous study published by Yuan et al. (2015) is in agreement with our finding, as the authors have reported a very high EC of 7750 μS cm -1 in the sewage sludge biochar that was not washed with ultrapure water after pyrolysis (Table 1). Fig. 1 shows the evolution of the zeta potential according to pH values for BSS, BH 2 O 2 , BKOH and BKOH bat . The corresponding pH PZC are given in Table 1. Results demonstrated similar pH PZC values for both BSS (∼2.7) and BH 2 O 2 (∼2.9) biochars, whereas a slightly higher value was obtained for BKOH (∼3.4). Such limited variations of pH PZC (2.7-3.4) are not prone to significantly modify the pH on the As and Cd sorption, except if the manipulations are performed at a pH lower than 5 (i.e. pH at sorption experiment, see section 2.4). In contrast, the more negative values of zeta potential observed for BH 2 O 2 and BKOH bat (Fig. 1) could indicate higher amounts of negative charges at the surfaces of these two modified biochars, resulting in higher sorption capacities towards cationic metals.

The pH PZC results showed that all biochars carried net negative charges at neutral pH (Fig. 1). At present, the pH PZC of biochar has been rarely reported in the literature and there is a lack of information to understand its variability among different chemical treatments. However, our results are in agreement with a straw-derived biochar with a pH PZC ∼1.9 (Qiu et al., 2009) and a paper mill sludge with pH PZC ranging from 4.0 to 6.0 (Guo et al., 2018).

The BET surface area (S BET ) of the raw and modified biochars is given in Table 1. A significant increase of the S BET was observed for BH 2 O 2 and BKOH, from 0.4 ( ± 0.1) to 5.7 ( ± 0.1) and 7.9 ( ± 0.1) m 2 g -1 , respectively. This increasing S BET is an indicator of higher porosity on the modified biochar and thus a growing chance for metal(loid)s to sorb onto the biochar surface (e.g. by complexation) (Mohan et al., 2014). Our findings are similar to those determined for sewage sludge biochars (4.0-14.3 m 2 g -1 ) (Agrafioti et al., 2013;Yuan et al., 2015) and pineapple-peel-derived biochars (0.7-2.1 m 2 g -1 ) (Wang et al., 2016). Nevertheless, these values are still far lower than those reported for activated carbons (1215.0-1316.0 m 2 g -1 ) (Gong et al., 2015).

The CEC values determined for the different biochars are presented in Table 1. Results showed a slight increase of CEC after treating the biochar with H 2 O 2 from 1.9 ( ± 0.1) to 3.0 ( ± 0.1) cmol + kg -1 , whereas a significantly increase (6 times) after the KOH treatment (P ≤ 0.01) was reported. This higher CEC of BKOH could induce an enhanced sorption ability towards cationic metals. Previously published data showed variable high ranges of the CEC of the biochars from 45.7 cmol + kg -1 to 483.4 cmol + kg -1 on wood and straw-based biochars, respectively (Ding et al., 2016;Jiang et al., 2014).

The variations in the S BET and CEC after chemical treatments of the biochars suggest that H 2 O 2 and KOH-modified biochars provided higher amounts of active pore sites, surface functional groups and exchangeable cations that could enhance the sorption of metallic elements on the biochar surface. For anion sorption, the increase of S BET could influence the arsenic sorption capacity (Table 1).

Concerning the effect of the washing procedures after KOH treatment, the results showed significantly higher values of both pH and conductivity on BKOH bat than BKOH (Table 1), which were attributed to a release of ions such as PO 4 3-and HCO 3 -/CO 3 2-from the biochar (Wongrod et al., 2018). As discussed previously (Wongrod et al., 2018), the insufficient batch washing after biochar modification led to the release of ions (e.g. PO , Ca 2+ and Mg 2+ ) and organic compounds from the biochar which interacted with metals and modified the sorption behavior. In the case of As(V), PO 43-is expected to be a strong competitor, reducing its sorption onto the biochar. The pH PZC , S BET and CEC values were also significantly affected by the uncomplete batch washing, resulting in an alteration of the biochar surface properties.

Effect of chemical treatment and washing on Cd(II) sorption

The results of the kinetic study of the Cd(II) sorption fitted better to the pseudo-second-order (PSO) (Fig. 2a) than intraparticle diffusion (Fig. 2b) models. All values for the kinetic model-fitting can be found in Table 2 and Table S2 (supplementary information). Results showed that the sorption of Cd(II) by all biochars were well-described by the PSO kinetic model (Table 2 and Fig. 2a), literally suggesting that the sorption mainly occurred via simple chemical reaction (Nethaji et al., 2013). Both the KOH and H 2 O 2 -treated SSD biochars showed faster removal kinetics than the raw biochar, reaching the equilibrium in less than 5 h for BH 2 O 2 and BKOH, and after 10 h for BSS (Fig. 2a). These results are consistent with a previous study with faster kinetics on the KOHmodified switchgrass biochar than the raw biochar (Regmi et al., 2012). Fig. 2b shows that the plots did not pass through the origin as the C values are not equal to zero (Table 2). This indicates some degree of boundary layer and the intraparticle diffusion was not the only rate controlling step, but other mechanisms like film diffusion and external mass transfer may also play important roles in the sorption (Nethaji et al., 2013).

Sorption isotherms are shown in Fig. 3, and the corresponding parameters can be found in Table 3 and Table S3 (supplementary information). Cd sorption was better described by the Langmuir model

Table 2

Constants for adsorption kinetics of As(III), As(V) and Cd(III) by biochars with the pseudo-second-order and intraparticle diffusion models.

Pseudo-second-order model

Biochar

As(III) As(V) Cd(II) (Table S3), suggesting the possibility of a monolayer sorption of Cd onto the homogeneous biochar surfaces (Zhao et al., 2015b). The Langmuir maximal sorption capacity, Q m , increased by a 15-fold-20fold factor after treatment with H 2 O 2 and KOH, respectively (Table 3). This can be linked to the previous observation that the chemical treatments not only increase the S BET , but also the CEC, resulting in an increase of the sorption sites and exchangeable cations. The higher Q m value of BKOH compared to BH 2 O 2 is consistent with the more pronounced improvement of both S BET and CEC with the KOH treatment (Table 1).

k 2 a Q e b R 2 k 2 Q e R 2 k 2 Q e R 2
k i d C R 2 k i C R 2 k i C R 2
The effect of the washing procedure applied after KOH treatment on the adsorption kinetics and isotherms are shown in Figs. 2 and3 and Tables 2 and3. The sorption kinetic constant for biochar with the KOH treatment was lower without the column washing step, but the value was still higher than that for the raw biochar. This finding could result from the reduction of S BET for the insufficient batch washed BKOH bat . A slightly higher Q m was observed for BKOH bat compared to BKOH, meaning that the sorption of this cationic metal was not strongly affected by the washing procedure. The release of ions and organic compounds from the BKOH bat did not appear to modify the Cd(II) behavior, in contrast to what was observed in a previous study with Pb(II) (Wongrod et al., 2018). This could be due to a higher affinity of Cd(II) for biochar sites (i.e. higher K L values than Pb(II)) and/or weaker interactions between Cd(II) and dissolved compounds in the solutions. For example in our findings, the K L values of BKOH were 0.032 l μmol -1 and 0.007 l μmol -1 for Cd(II) and Pb(II), respectively (Wongrod et al., 2018).

Arsenic sorption 3.3.1. Effect of chemical treatment of biochar on As(V) sorption

Fig. 4a shows the effect of chemical treatments on the adsorption kinetics for As(V) onto BSS, BH 2 O 2 and BKOH. All experimental data fitted with a pseudo-second-order kinetic model (R 2 ≥ 0.98, Table S2, supplementary information) and corresponding parameters are given in Table 2. These results showed that the sorption kinetics were unchanged for BH 2 O 2 , with a sorption equilibrium reached after 10 h, while it was reached after 24 h for BKOH.

The sorption isotherms for As(V) are shown in Fig. 4b. Similar to Cd, the Langmuir model provided the best fit for As(V) sorption (Table S3, supplementary information). The corresponding sorption parameters are given in Table 3. No significant evolution was observed for the values of K L , indicating that the chemical treatments did not modify the affinity of the biochar for As(V). In contrast, the sorption capacity was enhanced by both H 2 O 2 and KOH-treated biochars (except BKOH bat ), and the Q m values being, respectively, about 4 times and 5 times higher than the raw biochar.

These results indicate that H 2 O 2 treatment only improved the sorption capacity, whereas KOH treatment improved the sorption capacity but induced slower sorption kinetics for As(V) (Tables 2 and3). In both cases, the enhanced sorption ability for As(V) by the chemicallymodified biochars could be linked to the increase of the S BET , which was slightly higher for BKOH than BH 2 O 2 (Table 1). These results are in agreement with previous studies (Ding et al., 2016;Wu et al., 2017), that showed that the surface area plays an important role in the sorption of elements onto biochar. The slightly increased pH pzc of BKOH (Table 1) may result in a little increase of the positive charges at the surface of the biochar, and thus enhance the sorption ability for the anionic As(V).

Comparison of biochar washing

The effect of the washing procedure applied after KOH treatment on the adsorption kinetics and isotherms for arsenic are presented in Fig. 5a-d. The corresponding sorption parameters are given in Tables 2 and3. The experiments were performed for As(V) and As(III) to determine if the washing procedure affects similarly both arsenic species, taking into account that, in the experimental conditions, As(III) is neutral whereas As(V) is negatively-charged.

It appeared that the washing procedures of KOH-biochars had a low impact on the sorption kinetics for As(III), k 2 values increasing from 0.20 ( ± 0.05) h -1 to 0.38 ( ± 0.07) h -1 with the addition of the subsequent column washing step (Fig. 5a). In contrast, the addition of Table 3 Parameters determined from adsorption isotherms for As(III), As(V) and Cd(II) by biochars with the Langmuir model.

Biochar

As(III) As(V) Cd(II) this step strongly affected the sorption kinetics for As(V) (Fig. 5b), inducing a 3-fold decrease of k 2 (Table 2). The sorption isotherms showed an opposite trend: the column washing step only marginally modified the Q m of the biochar for As(III), but induced an almost 10-fold improvement for As(V). This strong difference of the Q m values for As(V) can be attributed to the PO 4 3-released from BKOH bat , as highlighted in a previous study (Wongrod et al., 2018). Indeed, the PO 4 3-may strongly compete with As(V) for sorption onto the biochar. It has been previously established that excessive PO 4 3-concentrations could mobilize As(V) from the sorbent (Neupane et al., 2014). This is likely due to the similar chemical properties between PO 4 3-and As(V) since both chemicals are negatively-charged oxyanions. Moreover, the leaching of dissolved organic carbon (DOC) from the biochar (Wongrod et al., 2018) could accelerate the mobilization of As, as observed in a previous study in a biochar-soil system (Kim et al., 2018). Concerning the sorption constant K L , there was no significant difference (P ≤ 0.01) according to the washing procedures for As(V) and As(III), indicating that no evolution in the affinity of biochar for both arsenic species was observed. When comparing K L values between As(III) and As(V), it appears that the latter has less affinity for BKOH and BKOH bat (Table 3).

K L a Q m b R 2 K L Q m R 2 K L Q m R 2 BSS c - - - 0
It could be hypothesized from these results that arsenic species interact differently with the biochar surface or with different types of surface functions.

Evolution of arsenic redox speciation

The arsenic redox speciation in solution was determined during As (III) and As(V) sorption experiments in order to determine whether the biochar was able to induce redox conversions. During the sorption of As (III) onto BKOH bat , up to 70% of arsenic in solution was As(V), despite only As(III) was initially added. A partial oxidation (< 10%) was also observed in the presence of BKOH and no As(V) was detected in the controls. In contrast, no reduction of As(V) to As(III) was noticed during As(V) sorption experiments. These results show that biochar is able to modify the arsenic redox state in solution and this could be strongly increased with an insufficient washing. These findings are in agreement with a previous study [START_REF] Fakour | Effect of humic acid on as redox transformation and kinetic adsorption onto iron oxide based adsorbent (IBA)[END_REF] that reported an increased oxidation of As(III) in the presence of organic matter in solution.

Conclusions

Chemical treatment of biochar is considered to enhance the sorption performance for both oxyanions and cationic metals. The treatment with KOH or H 2 O 2 strongly affects the sorption kinetics for As(III and V) and Cd(II). For As(V), KOH treatment induced a slower kinetic sorption compared to the raw biochar, while no significant variation was noticed for the H 2 O 2 -modified biochar. In contrast, both treatments showed faster sorption for Cd(II). All treatments induced a strong sorption capacity enhancement, particularly with alkali treatment. This is consistent with the observed evolutions of biochar properties such as CEC, S BET and, to a lesser extent, pH ZPC . Unfortunately, as previously highlighted, without the implementation of a careful washing procedure, this treatment also generated a strong release of inorganic and organic compounds from the biochar that can alter the evaluation of its sorption properties. This issue clearly affected As(V) sorption due to the competition with the released PO 4 3-.

When working with a redox-sensitive element, such as arsenic, it was observed that the biochar may also play an important role in the oxidation of As(III) to As(V). This phenomenon was enhanced by the insufficient washing process. The potential chemical transformation of arsenic should be carefully considered in future research in order to better understand the role of chemically modified biochar on the removal of such compounds from contaminated water.

Measurement of pH of point of zero charge (pHPZC) of biochar

The pHPZC of raw and chemically-modified biochars was determined from zetametry by determining the zeta potential at different pH ranges (Mahmood et al., 2011). Firstly, 40 mL of background electrolyte solution (0.01 M NaNO3) containing 0.2 g of biochar was equilibrated with continuous magnetic stirring at 20 (± 2) °C for 40 min. The pH of the mixture was adjusted in the range of 2.0 -10.0 by using HNO3 (0.1, 0.5 and 1 M) and NaOH (0.1, 0.5 and 1 M). After 40 min of stirring, the pH was measured and readjusted to correct the pH value until each solution reached a stable pH. The suspension was then filtered with a 0.45 µm cellulose acetate syringe filter before determining the zeta potential (Zetasizier Nanoseries (Nano-ZS), Malvern, UK). This filtration was performed to prevent the settlement of biochar particles in the measurement cell of the zetameter and to avoid the clogging of big particles in the device. We hypothesize that all biochar particles display the same behavior, neglecting the particle size.

Measurement of the cation exchange capacity (CEC) of biochar

For the determination of the CEC, 2 g of biochar (dry weight) were mixed with 40 mL of 0.05 N cobalt hexamine trichloride ([Co(NH3)6]Cl3) solution (99% w/w, Sigma-Aldrich) in polypropylene tubes (Aran et al., 2008). The suspensions were shaken for 1 h using an orbital shaker (KS 501 digital, IKA TM , USA) at 60 rpm, then centrifuged at 7000 g for 10 min (Rotina 420, Hettich, Germany). The supernatant was filtered through Whatman polyethersulfone syringe filters (0.2 µm). The pH (Präzisions-pH-meter E510, Metrohm AG, Switzerland) and absorbency at 472 nm (UV-vis spectrophotometer, Lambda 365, PerkinElmer, USA) of each sample were measured immediately. All experiments were performed in triplicate and the results are reported as mean values. The CEC (meq 100 g -1 or cmol⁺ kg -1 ) can be calculated as follows using Eq. (1) (Aran et al., 2008)

: CEC = [ Abs Co -Abs sample Abs Co ] × 50 × V m × 100 (1),
where AbsCo and Abssample refer to the absorbency (at 472 nm) of 0.05 N [Co(NH3)6]Cl3 and of sample (supernatant after filtration), respectively. V and m correspond to solution volume in liter and biochar dry mass in gram, respectively.

X-ray powder diffraction (XRD) analysis

X-ray diffractometry (XRD) was employed to identify the crystalline structures presented in raw and modified biochars. The samples were grounded to less than 100 µm particle size and were characterized using a powder diffractometer (AXS D8, Bruker, Germany) with Cu Kα radiation at 1.54 Å wavelength over the 2θ range from 10° to 32°, for 10 seconds per step, at 40 kV voltage and 40 mA current with a Solx (Si-Li) detector. The crystalline compounds present in biochar were identified using the X'Pert HighScore software for data analysis with the reference from the International Center for Diffraction Data (ICDD).

Fourier-transform infrared spectroscopy (FTIR) analysis

FTIR spectra were recorded to identify the functional groups present on biochar surfaces. Each sample was mixed with KBr in a ratio of 1 mg sample (particle size 100 µm) per 200 mg KBr and the pellet was prepared by compression under vacuum (Jouraiphy et al., 2005). The analysis was performed using a Shimadzu IRAffinity-1 Spectrometer with a deuterated-triglycine sulfate (DTGS) detector. The number of scans was 12 with a resolution of 16 cm -1 and a frequency in a range from 4000 to 800 cm -1 .

FTIR spectra of raw sewage sludge digestate and its derived biochar a Values reported as means (n=3) on dry mass basis followed by standard deviation.

b dl refers to value below detection limit.

Adsorption kinetics

The sorption kinetic data for arsenic and cadmium by biochars was calculated using the pseudofirst-order (PFO) (Eq. 2), pseudo-second-order (PSO) (Eq. 3) and intraparticle diffusion (Eq. 4) kinetic models, which are illustrated as followed:

log(Q e -Q t ) = logQ e - k 1 t 2.303 (2), t Q t = 1 k 2 Q e 2 + t Q e (3), Q t = k i t 1 2 ⁄ + C (4),
where Qt and Qe are arsenic or cadmium adsorption capacity (µmol g -1 ) at time t (h) and at equilibrium, k1 (h⁻ 1 ), k2 (g(µmol h) -1 ) and ki (µmol (g h 1/2 ) -1 ) are the rate constants for the PFO, PSO and intraparticle diffusion kinetic models, respectively (Liu and Zhang, 2009;Zhao et al., 2015). C is the intraparticle diffusion constant related to the boundary layer diffusion. All uncertainties of sorption kinetic parameters were obtained by non-linear regression using Statistica software (v6.1, StatSoft).

Adsorption isotherms

Langmuir and Freundlich models are presented, respectively, in Eqs. ( 4) and ( 5): ),

C e Q e = 1 K L Q m + C e Q m ( 5
lnQ e = lnK F + 1 n lnC e (6),
where Ce is the equilibrium arsenic or cadmium solution concentration (µM). Qe and Qm are the arsenic or cadmium adsorption capacity at equilibrium and at maximum capacity (µmol g⁻ 1 ), respectively. KL and KF are the adsorption constants for the Langmuir and Freundlich isotherm models, and n is the Freundlich constant which indicates the favorability of adsorption, respectively.

Table S3. Constants for the pseudo-first-order (PFO), pseudo-second-order (PSO) and intraparticle diffusion kinetic models for As(III, V) and Cd(II) sorption by different types of biochars (BSS: biochar from sewage sludge digestate, BH2O2: H2O2-modified biochar, BKOH and BKOH bat : KOH-modified biochars, respectively, with batch and column washing and with batch washing). a k1 and k2 units: h -1 and g(µmol h) -1 , respectively.

Biochar As(III) PFO PSO Intraparticle diffusion k1 a Qe b R 2 k2 a Qe R 2 ki C R 2 BSS - - - - - - - - - BH2O2 - - - - - - - - - BKOH 0
b Qe unit: in µmol g -1 .

Indeed, organic forms of As such as monomethylarsonic acid (MMA) and dimethylarsinic acid (DMA) present intermediate toxicity, whereas inorganic As species, i.e. arsenite (As(III)) and arsenate (As(V)), are known as the most toxic ones (Hughes et al., 2011).

Regarding the acute toxicity of As, efforts have been made to efficiently remove As from polluted water streams by using several treatment techniques such as oxidation, coagulation-flocculation, ion exchange, phytoremediation, membrane separation and adsorption (Jadhav et al., 2015;Jain and Singh, 2012;Singh et al., 2015). Among these methods, sorption is well known as a cost effective approach to remediating metal(loid)s polluted water (Ahmad et al., 2014). Biochar is a type of charcoal obtained from pyrolysis of biowaste materials in an oxygen-limited environment (Novotny et al., 2015). Because of the lowcost and the great abundance of biowaste feedstocks (e.g. sewage sludge digestate obtained from wastewater treatment), biochars are considered as alternatively potential sorbents for metal(loid)s removal from water (Mohan et al., 2014). Most of the sewage sludge biochars were used to sorb cationic metals like lead (Pb(II)) and chromium (Cr(VI)) [START_REF] Ifthikar | Highly efficient lead distribution by magnetic sewage sludge biochar: sorption mechanisms and bench applications[END_REF]Lu et al., 2012;Zhang et al., 2013), and to a lesser extent to metalloids like arsenic (As) and antimony (Sb). Thus, the development of sewage sludge biochar with chemical modification to improve As sorption (Sizmur et al., 2017) as well as the investigation of As speciation is of great interest. To date, As redox species repartition during the adsorption process onto biochar, and the role of biochar on redox modifications of As are not known. Hence, it is worthwhile to investigate the redox transformation of As(III) during its sorption onto biochars, since As(III) is more toxic and weakly bound to solid material than As(V) (Manning et al., 2002). Therefore, the As(III) oxidation induced by biochars is an important reaction that can possibly decrease As toxicity in As polluted water bodies.

Arsenic speciation in solid-phase samples can be accessed via X-ray absorption near edge structure (XANES) spectroscopy (Niazi et al., 2018a(Niazi et al., , 2018b)). However, the use of this technique is limited due to the very high operation cost and the accessibility to synchrotron facilities. Solid-liquid extraction followed by separation techniques, for instance, liquid chromatography (LC) coupled to spectrometric detection techniques such as atomic fluorescence spectroscopy (AFS) or inductively coupled plasma mass spectrometry (ICP-MS), could be an effective and more accessible technique to determine the As species sorbed onto the solid-phase samples.

The present study focused on the determination of the inorganic As redox distribution in raw sewage sludge digestate (SSD) biochar and the H 2 O 2 and KOH modified biochars before and after As(III) sorption using an analytical approach based on a solid-liquid extraction. At present, no study has reported on the implementation of an extraction method to recover As(III) and As(V) from the solid-phase of biochar. An extraction method using phosphoric and ascorbic acids as extracting agents was thus investigated before studying the possibility of redox transformations of As(III) during sorption experiments onto the biochar. The main objectives of this work were to: (1) validate the extraction procedure for As(III) and As(V) speciation in the biochars; (2) investigate the sorption ability for As(III) by the raw and chemicallymodified biochars; and (3) report on the redox conversions of arsenic during the sorption experiments.

Material and methods

Biochar production and chemical modification

Sewage sludge digestate (SSD) was collected from a wastewater treatment plant (WWTP) located in Limoges (France), after its dewatering and drying processes. Biochar was produced from the SSD sample under slow pyrolysis conditions (at 350 °C for 15 min) (Wongrod et al., 2018a). The SSD biochar (100 g) was then modified with a 1 L of 10% H 2 O 2 solution (modified from Xue et al., 2012) or with a 2.5 L of 2 M KOH solution (modified from Jin et al., 2014). For each biochar modification, the mixed biochar-solution was continuously stirred at room temperature for 2 h.

Raw and modified biochars were further washed with ultrapure water (18.2 MΩ, MilliQ Gradient A10, Millipore SAS 67120, Molsheim, France) in a batch system (triplicate in a row) followed by a continuous column washing (Wongrod et al., 2018a). These washing steps were performed to eliminate releasable organic compounds and inorganic ions (i.e. PO 4 3-, HCO 3

- , CO 3 2-
, Ca 2+ and Mg 2+ ) from the prepared biochars, particularly after chemical modification. The KOH modified biochar was also submitted to only a triple batch washing (Huang et al., 2017;Regmi et al., 2012;Wongrod et al., 2018a;Wu et al., 2017) to study the influence of washing steps onto As sorption and redox transformations.

The raw, H 2 O 2 and KOH modified SSD biochars are labeled as BSS, BH 2 O 2 and BKOH, respectively. The KOH modified SSD biochar with only batch washing is denoted as BKOH bat .

Biochar characterization

The pH of the biochar was measured using a pH-meter (LPH 330 T, Tacussel, France) after stirring 1 g biochar in 20 mL deionized water for 5 min and allowed it to settle for 15 min. The electrical conductivity (EC) of biochar was measured using a conductivity meter at 20 °C (CDM 210,Radiometer,Denmark).

The Brunauer-Emmett-Teller surface area (S BET ) of biochar was measured using N 2 sorption at 77 K (3Flex, Micromeritics, USA) after pretreatment of biochar by drying at 105 °C for 5 h. The pH of point of zero charge (pH PZC ) of the biochar was determined from the zeta potential at different pH ranges (Mahmood et al., 2011). The cation exchange capacity (CEC) of the biochar was determined by using a cobalt hexamine trichloride solution (99% w/w, Sigma-Aldrich) (Aran et al., 2008).

Sorption experiments

A 13.33 mM stock solution of arsenite (As(III)) was prepared from AsNaO 2 (98% w/w, Merck) and was diluted to 50 µM prior to As(III) sorption and arsenic speciation (i.e. As(III) and As(V)) experiments. Biochar (0.15 g) was separately added to 37.5 mL of As(III) solution in a polyethylene tube to obtain an initial 146 µg of As(III) in the solution. The initial pH was adjusted to 5.0 ( ± 0.5) by adding 0.01 M HNO 3 or NaOH. The sorption experiments were performed in triplicate at 20 ( ± 2) °C at 180 rpm for 24 h using an orbital shaker (KS 501 digital, IKA™, USA).

The significant presence in solution of releasable dissolved compounds (RDC) from BKOH bat (Wongrod et al., 2018a) could affect arsenic speciation. To determine whether the arsenic oxidation was mainly induced by the biochar itself or by these RDC, a control experiment was carried out. For this purpose, 1 g BKOH bat in 250 mL ultrapure water was stirred at 180 rpm for 24 h at room temperature (20 °C). The resulting solution was filtered with a 0.2 µm polyethersulfone (PES) membrane filter to remove the solid biochar. The resulting solution thus only contains the RDC. A control experiment was then performed by replacing the BKOH bat by this solution of RDC during sorption kinetics experiments.

In all these experiments, samples were collected at 1, 3, 7, 16 and 24 h. The total sampled volume was less than 5% of the initial volume to avoid any disturbance on the sorption equilibrium. Solutions were filtered through a 0.2 µm PES syringe filter and stored in a dark cold room at 4 °C until analysis.

Arsenic redox state distribution in biochars

Total As and its redox state distribution in raw and chemicallymodified biochars before and after As(III) sorption experiments were determined through solution analysis, extraction and acid digestion as described in the following subsections.

Arsenic analysis in solution and deduction of sorbed arsenic

Total arsenic analysis was performed using graphite furnace atomic absorption spectrometry (GF-AAS) (240Z, Agilent Technologies, USA) at λ 193.7 nm. The standard calibrations of As were prepared in the ranges of 10-50 µg L -1 . The detection limit was estimated from the mean of the blank and standard deviation, and the analyte solution was measured to produce a signal of at least 3 times higher signal than the noise level.

The arsenic redox distribution was assessed by liquid chromatography coupled to atomic fluorescence spectroscopy (LC-AFS) with hydride generation (HG) (PS Analytical Millennium Excalibur, PS Analytical, UK). The chromatographic separation was performed using a Hamilton PRP-X100 column with a phosphate buffer solution as the mobile phase at pH 6.9 and at a flow rate of 1 mL min -1 . The details of the experimental conditions for speciation analysis can be found in a previous study reported by Wan et al. (2014). To avoid any transformation between As(III) and As(V), arsenic speciation on both liquid and extracted solutions was performed within 5 h after sample recovery.

The amount of As sorbed onto the biochar was calculated from differences between initial and final concentrations of arsenic during sorption experiments. The sorbed As onto biochar can be calculated following Eq. ( 1):

= - × Q C C V (
)

sol i f .
(1) where Q sol. is the amount of arsenic sorbed onto biochar (µg) (based on 0.15 g of biochar in this study), C i and C f are, respectively, initial and final equilibrium concentrations of arsenic in the solution (µg L -1 ), and V is the total volume of the solution (L).

Extraction of As(III) and As(V) from biochar

Due to the lack of a soft extraction procedure for arsenic from biochar in the literature, this study designed one extraction method based on the extraction procedures described by Thomas et al. (1997), Montperrus et al. (2002) and [START_REF] Zhang | Efficiency and mechanisms of Cd removal from aqueous solution by biochar derived from water hyacinth (Eichornia crassipes)[END_REF]. The extractant was prepared from 0.3 M phosphoric acid (H 3 PO 4 , 85% w/w, Carlo ERBA) with or without the addition of 0.5 M ascorbic acid (99.5% w/w, Fluka). Ascorbic acid was used to prevent the oxidation of As(III) in the solution during the extraction (Xu et al., 2015).

After the sorption experiments, 0.15 g of biochar was quickly rinsed 3 times with ultrapure water and transferred into a Teflon digestion tube with 25 mL of the extracting solution. The extraction was operated using microwave assistance (Multiwave GO, Anton Paar, France) by heating at 80 °C for 20 min. The resulting solution was then centrifuged (Multifuge X3 FR, Thermo Fisher Scientific) at 3400 rpm for 20 min. The supernatant was collected and filtered through a 0.2 µm PES syringe filter and diluted to a constant volume of 50 mL. Samples were analyzed for As speciation using LC-AFS as mentioned in Section 2.4.1.

Biochar acid digestion for total arsenic measurements

After the sorption experiments, the biochar (about 0.15 g) was rinsed with ultrapure water and transferred into a digestion tube (Teflon). Concentrated nitric acid (HNO 3 , 69.5% w/w, Panreac ITW) (6 mL) was added to the digestion tube. Hydrogen peroxide (H 2 O 2 , 30% w/w, Carlo ERBA) (6 mL) was then slowly added into the digestion tube with a caution regarding its strong oxidizing property. The digestion tube was left overnight (12 h) to oxidize and digest organic matter in biochar prior to a microwave acid digestion (in order to avoid overpressure due to the high initial release of CO 2 ). Finally, 3 mL of concentrated hydrochloric acid (HCl, 37% w/w, VWR) were added before starting the digestion that was operated at 180 °C for 4 h (4 cycles) using a microwave digestion (1 cycle per 1 h) to sufficiently digest all biochar particles. After digestion, the resulting solution was recovered and the dilution in a 50 mL volume was made with ultrapure water. It was filtered through a 0.2 µm PES syringe filter before analysis. Total arsenic content was analyzed by GF-AAS as previously mentioned in Section 2.4.1. Regarding the analysis of total arsenic in the biochars, the As content that is already present in each biochar (before sorption) was deduced from the sorbed As after the As(III) sorption experiments. This deduction was made to avoid errors in the experimental results.

Statistical analysis

The biochar characteristics and As(III) sorption experiments were carried out in triplicate. Results are reported as the mean value followed by standard deviation. Constants for sorption parameters were obtained by the non-linear regression using Statistica software (v6.1, StatSoft). Statistical analysis of the experimental data was performed by the t-test with two-tailed distribution at a statistical significance level of p ≤ 0.01.

Results and discussion

Characterization of raw and chemically-modified biochars

The total As content in the raw and chemically-modified biochars is given in Table 1. Results show a significant (p ≤ 0.01) decrease of the As content after treating the raw biochar with H 2 O 2 or KOH from 47 ( ± 1) µg g -1 for BSS to 29 ( ± 1) and 13 ( ± 1) µg g -1 , respectively, for BH 2 O 2 and BKOH. These reductions can be due to the ability of KOH to dissolve ash contents (Lin et al., 2012;Liou and Wu, 2009;Liu et al., 2012) and H 2 O 2 to oxidize organic matter present in the solid phase of the biochars (Xue et al., 2012). As a result, the As bound to the biochar matrix could be chemically altered and thus the As contents were lower in the chemically-modified biochars compared to the raw biochar (Table 1).

Considering BKOH and BKOH bat , an almost 2 times decrease of As, from 23 ( ± 2) µg g -1 for BKOH bat to 13 ( ± 1) µg g -1 for BKOH, was found after the subsequent continuous column washing.

The pH values of all biochar suspensions are also provided in Table 1. The pH of BSS (6.4 ± 0.1) increased to 8.4 ( ± 0.1) and 10.1 ( ± 0.1) for BKOH and BKOH bat , respectively, whereas H 2 O 2 induced no significant change (p ≤ 0.01) on the pH suspension of biochar. The electrical conductivity (EC) of biochar suspensions remained in similar ranges from 4.0 ( ± 0.1) to 4.1 ( ± 0.3) and 6.2 ( ± 0.5) µS cm -1 for Note: all values reported are mean of triplicate followed by standard deviation, except pH PZC and zeta potential. a S BET refers to Brunauer-Emmett-Teller surface area of biochar.

b CEC refers to cation exchange capacity of biochar.

BSS, BH 2 O 2 and BKOH, respectively. However, a significantly higher EC value (324 ± 2 µS cm -1 ) for BKOH bat was found, compared to BKOH (Table 1). This highlights the ability of BKOH bat to release a relatively high amount of dissolved ions as previously reported by Wongrod et al. (2018a).

The pH PZC and zeta potential (at pH 5) values of the raw and chemically-modified biochars are given in Table 1. Results showed similar pH PZC ranges (2.7-3.4) among the raw and modified biochars. Currently, there is still a lack of information regarding pH PZC of biochars reported in the literature. Nevertheless, these findings are consistent with previous studies from Qiu et al. (2009) who reported a pH PZC of 1.9 on straw biochar and Petrovic et al. (2016) with pH PZC values of 4.5 and 6.0, respectively, on grape pomace biochar and its KOH modified biochar. Moreover, the negative zeta potential values of the biochars implied that all biochars carried net negative charges at pH 5, particularly for the BH 2 O 2 (-16.5 mV) and BKOH bat (-17.1 mV). These negatively charges may hinder the sorption ability for arsenic oxyanions at pH 5.

The BET surface area (S BET ) of all biochars is reported in Table 1. The S BET was significantly improved by 7-20 times after H 2 O 2 or KOH modification of the biochar compared to the non-treated biochar. In this case, the S BET increased from 0.4 ( ± 0.1) (the raw biochar) to 5.7 ( ± 0.1) and 7.9 ( ± 0.1) m 2 g -1 for BH 2 O 2 and BKOH, respectively. This implies that H 2 O 2 and KOH induced higher porosity on the modified biochars, thus As sorption can be enhanced for these biochars (Mohan et al., 2014). These S BET results are in agreement with those reported for sewage sludge biochars (4.0-14.3 m 2 g -1 ) (Agrafioti et al., 2013;Yuan et al., 2015), but lower than biochar made from wood (475 m 2 g -1 ) (Niazi et al., 2018b). Nevertheless, they remain very low compared to activated carbon which displays specific surface area beyond 1000 m 2 g -1 in most cases (Gonzalez-Garcia, 2018) (e.g. 1215.0-1316.0 m 2 g -1 for activated carbon prepared from coal) (Gong et al., 2015).

The CEC values for all biochars are also given in Table 1. Results demonstrate a similar CEC of the biochar after H 2 O 2 modification, i.e. from 2.0 ( ± 0.1) to 3.0 ( ± 0.1) cmol + kg -1 , whereas the CEC values were much higher (7-10 times) for BKOH (13.4 ± 0.1 cmol + kg -1 ) and BKOH bat (20.8 ± 0.1 cmol + kg -1 ) (Table 1). Nevertheless, in the literature much higher CEC values are reported for wood and straw derived biochars, ranging from 45.7 to 483.4 cmol + kg -1 (Ding et al., 2016;Jiang et al., 2014).

The biochar properties reported in Table 1 highlight that the total As content, zeta potential and S BET are the key indicators that were significantly affected by the biochar treatment. Due to its lower negative charge and its higher S BET , BKOH can potentially sorb more As compared to other biochars. Furthermore, the biochar washing procedure is an important parameter to be considered after the chemical treatment, since the biochar properties significantly change between BKOH and BKOH bat .

3.2.

As redox state evolution during sorption kinetics by BKOH bat and its dissolved solutions Fig. 1a shows the As redox state distribution in the solution during adsorption experiments for As(III) onto BKOH bat . The results demonstrate that while As was gradually sorbed onto BKOH bat (corresponding to the decrease of total As over time), the oxidation of As(III) to As(V) occurred continuously. To elucidate whether this oxidation was promoted by compounds released from the biochar into the solution or by the biochar itself, a control experiment was performed by replacing the biochar by a solution of released dissolved compounds (RDC) from BKOH bat . The experimental data and corresponding results are shown in Fig. 1b. From the results, a slight oxidation of As(III) with a final proportion of 9% As(V) in the RDC solution was found, compared to 43% oxidation with the presence of biochar. This demonstrates that the oxidation of As(III) to As(V) was mainly induced by the biochar material itself and to a lesser extent by the dissolved compounds being released from the BKOH bat . At present, there is still a lack of information on the role of biochars and the released dissolved compounds towards the As(III) oxidation. Nevertheless, the findings are supported by Niazi et al. (2018aNiazi et al. ( , 2018b) ) who found the oxidation of As(III) sorbed onto biochars prepared from Japanese oak wood and perilla leaf. However, Dong et al. (2014) showed the high potential oxidation of As (III) (up to 25%) induced by the dissolved organic matter from sugar beet tailing and Brazilian pepper derived biochars.

Extraction procedure for As redox state distribution in biochars

In order to assess the redox state distribution of As sorbed onto biochar by LC-AFS, it was first necessary to validate an extraction procedure that allows a quantitative As recovery without any conversion of As species, especially As(III) being oxidized to As(V).

The importance of adding ascorbic acid in the extraction solution to stabilize As(III) was assessed by applying the extraction procedure to a biochar sample with and without spiking with a known amount of As (III) just before performing the extraction step. In this study, the BKOH bat was selected due to its release of dissolved compounds and its ability to oxidize As(III) as previously reported by Wongrod et al. (2018aWongrod et al. ( , 2018b)).

Table 2 shows the recoveries of the arsenic species and total As after the extraction using H 3 PO 4 with or without addition of ascorbic acid onto the BKOH bat sample, and with and without As(III) spiking just before the extraction. Results show that all spiked arsenic was recovered by the extraction procedure. The results also highlight that when using only 0.3 M H 3 PO 4 , a significant amount of spiked As(III) was oxidized. Indeed, the addition of 12.5 µg of As(III) to the biochar before performing the extraction step resulted in an increase of 8.8 ( ± 0.7) µg and 3.5 ( ± 0.3) µg for As(III) and As(V), respectively. Thus, 29% of spiked As(III) was converted into As(V). In contrast, no oxidation of As(III) was observed during extraction with H 3 PO 4 and ascorbic acid: the difference between the As(III) amount before and after spiking is 12.6 ( ± 0.7) µg, which is in agreement with the amount of As(III) added (12.5 µg). Therefore, the addition of ascorbic acid is essential to ensure As(III) stability during the extraction step. The findings are also in agreement with the study of Xu et al. (2015) who reported the efficiency of using both phosphoric and ascorbic acids as extracting agents to quantitatively recover As(III) and As(V) from fly ash.

To estimate the extraction yield of As sorbed onto the biochar after the As(III) sorption experiments, the biochars were recovered and submitted to both acid digestion and extraction procedures (see Section 2.4). The corresponding results are presented in Table 3. For BSS and BH 2 O 2 , the amount of total sorbed As was too low to allow an accurate determination after the acid digestion and/or extraction. As a result, no extraction yield was calculated. For BKOH bat , the total amount of sorbed As was quantitatively recovered by the extraction procedure. In the case of BKOH, most of the sorbed As was recovered but about one fourth could not be extracted under the applied conditions. This difference in the extraction efficiency may result from different sorption mechanisms of As from one biochar type to the other. Nevertheless, the majority of As could be extracted from the biochars and thus the identification of the As species in the extracted solutions gives the main As species sorbed onto the biochars.

Determination of As(III) and As(V) sorbed onto biochars

Different types of biochars were exposed in the solution containing an initial As(III) amount of 146 µg. After 24 h of As(III) sorption, the biochars were recovered and the exposure solutions were analyzed to determine the remaining amount of As and to assess its distribution between As(III) and As(V). The biochars were separately submitted to acid digestion for the determination of the total quantity of sorbed As and to the extraction for the assessment of the redox distribution of sorbed As. The amounts of As(III), As(V) and total As in the exposure solutions and sorbed onto the biochars (determined after acid digestion and extraction steps) are shown in Table 4.

Results showed that the total As (i.e. As(III) + As(V)) remaining in the final exposure solutions of BSS and BH 2 O 2 were not significantly different from the initial As(III), indicating that no As was sorbed onto these biochars. In contrast, As was removed from the exposure solutions by both BKOH and BKOH bat , corresponding to an As sorption efficiency of 67% and 50%, respectively. These results are in agreement with the sorption yield estimated with the amount of As quantified in the digestate-based biochars (Table 4).

The arsenic redox distribution in the final exposure solution shows almost no As(III) oxidation for both BSS and BH 2 O 2 as the final amount of As(III) remained the same as the initial As(III) amount (Table 4). For the KOH treated biochars, the assessment of the redox distribution only concerned the unsorbed As. In the case of BKOH, no significant oxidation of As(III) was observed, whereas for BKOH bat about 28% of the remaining As was oxidized to As(V) (Table 4). These findings are in agreement with a previous study from Wongrod et al. (2018b) with a large oxidation of As(III) in BKOH bat and a partial oxidation in BKOH during As(III) sorption. As previously discussed (Fig. 1a and1b), the majority of As(III) oxidation was induced by the biochar solid compounds.

Considering As(III) and As(V) extracted from the biochars after the sorption experiment, a similar distribution of the As species was found for BKOH and BKOH bat (Table 4). Arsenic was mainly sorbed as As(III) (90-92%) and only 8-10% was sorbed as As(V) onto the KOH biochars. In the case of BKOH, the redox distribution of As sorbed onto the biochar is quite similar to the distribution observed in the final exposure solution. However, for BKOH bat , the evolution of the As species in solution and on the biochar was different: As(V) represents only 8% of the sorbed arsenic whereas it corresponds to 28% of the dissolved As(V) (Table 4). This suggests that the As(III) sorption onto the KOH biochar was favored compared to As(V) sorption. This finding is consistent with results previously reported by Wongrod et al. (2018b) when comparing As(III) and As(V) sorption onto such biochars. Another possible explanation would be that As(V) could be reduced into As(III) during the sorption process, as observed by Niazi et al. (2018aNiazi et al. ( , 2018b) ) who studied the speciation of As onto wood and leaf derived biochars using the solid-phase XANES technique. Depending on the nature and chemical treatment applied to the biochar, it could also contain chemical functions that act as electron donors to facilitate the As reduction (Choppala et al., 2016).

Comparison of the two KOH biochars showed that the adsorption capabilities for As(III) were quite similar for both the BKOH (410 µg g -1 ) and BKOH bat (385 µg g -1 ) at As(III) concentration ranges of 0-4000 µM (or 0-54 µg L -1 ) and the initial solution pH at 5.0. Nevertheless, a much higher As(III) sorption was observed onto the Japanese oak wood derived biochar (3,890 µg g -1 ) (concentrations of 27-144 µg L -1 and at pH 7.0 ± 0.1) (Niazi et al., 2018b). This is probably due to different biochar properties between the SSD-based biochars and the Japanese oak wood biochar. For instance, the specific surface area was substantially lower with the KOH modified biochars from SSD (3.0-7.9 m 2 g -1 ) (Table 1) compared to the oak wood biochar (475 m 2 g -1 ) (Niazi et al., 2018b).

In the case of BH 2 O 2 , the amount of sorbed As being low, the results of the As redox distribution were less accurate and should thus be taken with caution. The percentage of As(V) sorbed onto the H 2 O 2 -modified a na refers to not available due to very low values detected by liquid chromatography coupled to atomic fluorescence spectroscopy (LC-AFS) and/or graphite furnace atomic absorption spectrometry (GF-AAS).

biochar appears to be higher than into solution, highlighting a potential oxidation of As(III) during sorption. This phenomenon was also observed by Niazi et al. (2018aNiazi et al. ( , 2018b)), who suggested that redox active species present at the biochar surface such as NO 3 -or FeO(OH) could induce As(III) oxidation. Therefore, the presence of iron (Fe) (65 g kg -1 ) and manganese (Mn) (769 mg kg -1 ) on the SSD biochar, that could be partially in metal oxide forms (Wongrod et al., 2018a), could promote the redox transformation of As(III) to As(V) on metal oxides associated onto the biochars (Vithanage et al., 2017). Such a phenomenon is also reported by several studies (Han et al., 2011;Manning et al., 2002;Wang et al., 2015).

Conclusions

The extraction using phosphoric and ascorbic acids allowed recovering sorbed arsenic, while preserving As(III) oxidation state. Arsenic redox distribution could thus be assessed in both solid and liquid phases. Arsenic was mainly sorbed onto biochar as As(III) but during sorption, As(III) oxidation may occur over time. This oxidation, mainly induced by biochar solid compounds rather than by soluble compounds released from biochar, can strongly vary depending on the chemical treatment and efficiency of washing procedures applied to biochar after chemical treatment. The KOH modification efficiently improves the biochar sorption capacity but also promotes As(III) oxidation, especially with an incomplete washing. a na refers to not available due a value close to limit of quantification.

Figure 2 . 1 :

 21 Figure 2.1: Overview of biochar production technologies (data was obtained and modified from Kambo and Dutta (2015); Novotny et al. (2015); Sohi et al. (2010)).

Figure 2 . 2 :

 22 Figure 2.2: Interaction of biochar with metal(loid) pollutants during the sorption mechanism.

Figure 4 . 1 :

 41 Figure 4.1: Scheme of biochar production from organic fraction of municipal solid waste digestate (OFMSWD) under slow pyrolysis at lab scale.

Figure 4 . 2 :

 42 Figure 4.2: Biochar production from organic fraction of municipal solid waste digestate (OFMSWD) at different pyrolysis temperatures (residence time of 1 h and a heating rate of 15 °C min⁻ 1 ).

  Figure 4.3.

Figure 4 . 3 :

 43 Figure 4.3: Overview of chemical treatments and washing procedure of biochar for metal(loid)s sorption and arsenic redox state distribution on biochar and in solution experiments (Papers I, II and III).

Figures 5 .

 5 Figures 5.1(a) and5.1(b) show the change of XRD spectra between feedstocks and biochars after pyrolysis of the OFMSWD and SSD, respectively. XRD analysis showed that OFMSWD contained the crystalline structures of quartz (SiO2) and calcite (CaCO3) (Figure5.1(a)). Similar trends were also found in its derived biochar with slight shifts on the XRD spectra. On the other hand, only SiO2 was detected in SSD and BSS (Figure5.1(b)). The results showed no significant differences of XRD spectra between the digestates and the biochars. This indicates that the pyrolysis only induced the insignificant modification of surface properties of biochars. Nevertheless, the presence of CaCO3 on the biochar, in this case on the BMSW, may contribute to more sorption of metal elements(Al-Degs et al., 2006; Jiang et al., 2009).

Figure 5 . 1 :

 51 Figure 5.1: XRD spectra of the organic fraction of municipal solid waste digestate (OFMSWD) and its derived biochar produced at 400 °C (BMSW) (a) and sewage sludge digestate (SSD) and its derived biochar produced at 350 °C(BSS) (b) (Q and C refer to quartz (SiO2) and calcite (CaCO3), respectively).

Figures 5 .

 5 Figures 5.2(a) and 5.2(b) illustrate, respectively, the changes of FTIR spectra of the OFMSWD and its derived biochar (BMSW), and the SSD and its biochar (BMSW) obtained before and after pyrolysis. From Figure 5.2(a) and 5(b), the FTIR spectra showed that both the OFMSWD and SSD contained the absorbance peaks of hydroxyl (-OH) groups (3425 cm⁻ 1 ), -CH3 stretching of aliphatic groups (2924 cm⁻ 1 ), and C-C skeleton and C-O stretching (1033 cm⁻ 1 ). Similar surface functional groups were also detected on their derived biochars. Nevertheless, a strong intensity of -CH2 scissoring

Figure 5 . 2 :

 52 Figure 5.2: FTIR spectra of the organic fraction of municipal solid waste digestate (OFMSWD) and its derived biochar produced at 400 °C (BMSW) (a) and sewage sludge digestate (SSD) and its derived biochar produced at 350 °C(BSS) (b).

Figure 5 . 3 :

 53 Figure 5.3: Evolution of the zeta potential as a function of the pH values.

  of metals such as Pb(II) and Cd(II) may be enhanced by the modified biochar (e.g. via complexation)(Mohan et al., 2014) (Paper II). The global negative charge of biochars (from zeta potential) also indicate a better potential for modified biochars to sorb cationic metal ions but this may hinder the sorption ability for oxyanions such as As. In the case of BSS-KOH, the higher values of zeta potential (i.e. the lower negative charge) may allow a better sorption capacity towards As compared to other biochars.As discussed in chapter 4 (section 4.1.2), the SSD biochars were washed with ultrapure water in two approaches: (1) a triple batch washing in a row until the pH became stable and (2) the same batch washing followed by continuous column washing. During the column washing, the pH, EC, PO4 3 ⁻, HCO3⁻/CO3 2 ⁻ and DOC were measured over time until completion to ensure the elimination of all these releasable compounds (Paper I).

Figures 5 .

 5 Figures 5.4(a) and5.4(b) represent the evolutions of pH and EC over time during a continuous column washing of BSS-KOH bat and BSS-H2O2 bat , respectively. The pH values of BSS-KOH bat decreased steadily after 23 h of washing and became stable at pH 8.4 (± 0.1) after 64 h whereas for the H2O2-modifed biochar, it remained almost stable at 6.5 (± 0.1). Similarly, dramatic decreases of EC occurred throughout the manipulations on both modified biochars, which is consistent with the decreasing trends of the PO4 3 ⁻, HCO3⁻/CO3 2 ⁻ and DOC).

Figures 5 .

 5 Figures 5.4(c) and 5.4(d) show the release of PO4 3 ⁻, HCO3⁻, CO3 2 ⁻ and DOC from the BSS-KOH bat and BSS-H2O2 bat , respectively, over time during column washing. The concentrations of PO4 3 ⁻, CO3 2 ⁻, HCO3⁻ and DOC continuously decreased along the washing time(Figure 5.4(c, d)). These reductions are in correlation to the decline of conductivity (Figure5.4(a, b)). For example, the high concentrations of PO4 3 ⁻ (55 mg

Figure 5 . 4 :

 54 Figure 5.4: Evolution of pH and conductivity during continuous washing of BSS-KOH bat (a) and BSS-H2O2 bat (b), and concentration profile of released phosphate, (bi)carbonate and dissolved organic carbon (DOC) for BSS-KOH bat (c) and BSS-H2O2 bat (d) (Paper I).

Figure 5 . 5 :

 55 Figure 5.5: Adsorption kinetics for Pb(II) by BMSW bat , BMSW-H2O2 bat , BMSW-KOH bat , BSS bat , BSS-H2O2 bat and BSS-KOH bat at different contact times fitted with the pseudo-second-order kinetic model (initial pH=5, initial concentration: 482.6 µmol L -1 (100 mg L -1 )) (Paper I).

Figure 5 . 6 :

 56 Figure 5.6: Effect of chemical treatment on sorption kinetics for Cd(II) by the raw and chemicallymodified sewage sludge biochars fitted with the pseudo-second-order kinetic model (initial pH=5, initial concentration: 10 µmol L -1 ) (Paper II).

Figure 5 .

 5 Figure 5.7 shows the influence of chemical treatment on the sorption kinetics for As(V) by the raw, H2O2 and KOH modified biochars submitted to batch and subsequent column washings (i.e. BSS, BSS-H2O2 and BSS-KOH, respectively). No data on As(III) sorption was shown due to a very low or absence of sorption by the BSS and BSS-H2O2. The PFO, PSO and intraparticle diffusion models were used to describe the experimental data and the corresponding values are reported in Table5.5.

Figure 5 . 7 :

 57 Figure 5.7: Effect of chemical treatment on sorption kinetics for As(V) by the raw and chemicallymodified sewage sludge biochars fitted with the pseudo-second-order kinetic model (initial pH=5, initial concentration: 10 µmol L -1 ) (Paper II).

Figures 5 .

 5 Figures 5.8(a),5.8(b) and5.8(c) compare the effect of biochar washing procedures (i.e. batch versus batch and subsequent column washings) after KOH treatment for Cd(II), As(III) and As(V) sorption kinetics, respectively. The model-fitting constants are also given in Table5.6. The PSO kinetic model showed the best-fit to explain the

Figure 5 . 8 :

 58 Figure 5.8: Influence of biochar washing of KOH-modified biochars on sorption kinetics for Cd(II) (a), As(III) (b) and As(V) (c) fitted with the pseudo-second-order kinetic model (initial pH=5, initial concentration: 10 µmol L -1 ) (Paper II).

Figures 5 .

 5 Figures 5.9(a) and 5.9(b) represent the Pb(II) sorption as a function of equilibrium Pb(II) concentrations by batch washed BMSW bat , BMSW-H2O2 bat and BMSW-KOH bat (5.9(a)) and by BSS bat , BSS-H2O2 bat and BSS-KOH bat (with zooms of Ce at low concentration ranges) (5.9(b)), respectively. The constant parameters of model-fittings for Langmuir and Freundlich isotherms for Pb(II) are reported in Table5.7.

Figure 5 . 9 :

 59 Figure 5.9: Sorption isotherms for Pb(II) by BMSW bat , BMSW-H2O2 bat and BMSW-KOH bat (a) and by BSS bat , BSS-H2O2 bat and BSS-KOH bat with zooms of Ce at low concentration ranges (b) as a function of equilibrium Pb(II) concentrations (initial pH=5, 24 h, initial concentration: 0-5000 µM, Langmuir model-fittings except BSS-KOH bat due to non-fitting) (Paper I).

Figures 5 .

 5 Figures 5.10(a) and5.10(b) show, respectively, the Cd(II) and As(V) sorption isotherms by batch and subsequent column washed BSS, BSS-H2O2 and BSS-KOH (Paper II). The constant values corresponding to Langmuir and Freundlich sorption models are given in Table5.8.

Figure 5 . 10 :

 510 Figure 5.10: Effect of chemical treatment and biochar washing on sorption isotherms for Cd(II) (a) and As(V) (b) by the raw and chemically-modified sewage sludge digestate biochars (initial pH=5, 24 h, Langmuir model fitting) (Paper II).

Figures 5 .

 5 11(b) and5.11(c) compare the biochar washing procedures: batch versus batch and subsequent continuous column washings of the KOH and H2O2 modified biochars, respectively, on the Pb(II) sorption isotherms at low Ce of Pb(II).

  Figures

Figure 5 . 11 :

 511 Figure 5.11: Comparison of the Pb(II) sorption isotherms on sewage sludge digestate-based biochars with different washing modes fitted with the Langmuir model (except BSS-KOH bat due to non-fitting) (a).Figure (b) to (e) are zooms of Ce at low concentration ranges (<200 µM); batchwashed BSS-KOH bat versus continuous washed BSS-KOH (b), BSS-H2O2 bat versus BSS-H2O2

  Figure (b) to (e) are zooms of Ce at low concentration ranges (<200 µM); batchwashed BSS-KOH bat versus continuous washed BSS-KOH (b), BSS-H2O2 bat versus BSS-H2O2 (c), and with different chemical treatments; BSS-KOH bat versus BSS-H2O2 bat (d), and BSS-KOH versus BSS-H2O2 (e) (Paper I).

Figure 5 . 12 :

 512 Figure 5.12: Influence of biochar washings of KOH-modified sewage sludge digestate biochars on sorption isotherms for Cd(II) (a), As(III) (b) and As(V) (c) (initial pH=5, 24 h, Langmuir model fitting) (Paper II).

a

  nq refers to not quantified due to very low values detected by liquid chromatography coupled to atomic fluorescence spectroscopy (LC-AFS) with hydride generation (HG) and/or graphite furnace atomic absorption spectrometry (GF-AAS).

Figure 5 .

 5 Figure 5.13: Arsenic redox distribution in solutions during sorption kinetics for As(III) by BSS-KOH bat (a) and for control with only released dissolved compounds (RDC) from BSS-KOH bat (b) (Paper III).

  Fig. 1(a) and (b) represent the evolution of pH and conductivity of SS-KOH sem and SS-H 2 O 2 sem

Fig. 1

 1 (c) and (d) show the changes of phosphate, inorganic carbon and dissolved organic matter released from SS-KOH sem and SS-

Fig. 1 .

 1 Fig. 1. Evolution of pH and conductivity during continuous washing of SS-KOH sem (a) and SS-H 2 O 2 sem (b), and concentration profile of phosphate, (bi)carbonate and dissolved organic carbon (DOC) along continuous washing of SS-KOH sem (c) and SS-H 2 O 2 sem (d).

Fig. 3 (

 3 Fig. 3(a) and (b) show the adsorption isotherms for Pb by semicontinuous washed SS sem , SS-H 2 O 2 sem , SS-KOH sem and by MSW sem , MSW-H 2 O 2 sem and MSW-KOH sem , respectively. Table 3 presents the fitting parameters of the Langmuir and Freundlich isotherm models. All experimental data could be well-described by the Langmuir adsorption isotherms model (R 2 > 0.98), except for SS-KOH sem . The results from Fig. 3(a) demonstrate that chemical treatment of biochar could enhance the Pb sorption capacity from 6.5 mg g À1 on SS sem to 25.1 mg g À1 on SS-H 2 O 2

Fig. 2 .

 2 Fig. 2. Adsorption kinetics for Pb by semi-continuous washed MSW sem , MSW-H 2 O 2 sem , MSW-KOH sem , SS sem , SS-H 2 O 2 sem and SS-KOH sem at different contact times at an initial solution pH of 5 and initial Pb concentration of 100 mg L À1 fitted with the pseudosecond order kinetic model.

Fig. 3 .

 3 Fig. 3. Adsorption isotherms for Pb by semi-continuous washed SS sem , SS-H 2 O 2 sem and SS-KOH sem (a) and by MSW sem , MSW-H 2 O 2 sem and MSW-KOH sem (b) at different equilibrium Pb(II) concentrations at an initial solution pH of 5 fitted with the Langmuir model (except SS-KOH sem due to non-fitting).

Fig. 4 .

 4 Fig. 4. Comparison of the Pb adsorption isotherms on sewage sludge digestate-based biochars with different washing modes fitted with the Langmuir model (except KOH sem due to non-fitting) (a).Figure (b) to (e) are zooms of C e at low concentration ranges (<100 mg L À1 ); semi-continuous washed KOH sem versus continuous washed KOH con (b), H 2 O 2 sem versus

  Fig. 4. Comparison of the Pb adsorption isotherms on sewage sludge digestate-based biochars with different washing modes fitted with the Langmuir model (except KOH sem due to non-fitting) (a).Figure (b) to (e) are zooms of C e at low concentration ranges (<100 mg L À1 ); semi-continuous washed KOH sem versus continuous washed KOH con (b), H 2 O 2 sem versus H 2 O 2 con (c), and with different chemical treatments; KOH sem versus H 2 O 2 sem (d), and KOH con versus H 2 O 2 con (e).

Fig. 5 .

 5 Fig. 5. Comparison of Ca 2þ and Mg 2þ released by SS-KOH sem (semi-continuous washed) and SS-KOH con (continuous washed) (a), and the amount of adsorbed Pb versus the summation of the released cations (Ca 2þ and Mg 2þ ) at Pb adsorption equilibrium (b).

Fig. S2 .

 S2 Fig. S2. Comparison of the FTIR spectra of semi-continuous washed SS sem and MSW sem (a),

Fig. S3 .

 S3 Fig. S3. The evolution of solution pH during Pb adsorption isotherms onto SS-KOH sem (a) and MSW-KOH sem (b).

Fig. 1 .

 1 Fig. 1. Evolution of the zeta potential as a function of the pH values.

Fig. 2 .

 2 Fig. 2. Effect of chemical treatment and biochar washing on adsorption kinetics for Cd(II) by the raw and chemically-modified biochars fitted with the pseudosecond-order kinetic model (a), and with the intraparticle diffusion model (b) (initial pH = 5, initial concentration: 10 μmol l -1 ).

Fig. 3 .

 3 Fig. 3. Effect of chemical treatment and biochar washing on adsorption isotherms for Cd(II) by the raw and chemically-modified biochars (initial pH = 5, 24 h, Langmuir model fitting).

Fig. 4 .

 4 Fig. 4. Effect of chemical treatment on As(V) adsorption kinetics by biochars (a) (initial pH = 5, initial concentration: 10 μmol l -1 , pseudo-second-order kinetic model fitting), and As(V) adsorption isotherms by biochars (b) (initial pH = 5, 24 h, Langmuir model fitting).

Fig. 5 .

 5 Fig. 5. Influence of biochar washing of KOH-modified biochars on adsorption kinetics for As(III) (a) and As(V) (b) (initial pH = 5, initial concentration: 10 μmol l -1 , pseudo-second-order kinetic model fitting), and on adsorption isotherms for As(III) (c) and As(V) (d) (initial pH = 5, 24 h, Langmuir model fitting).

Fig. S2 .

 S2 Fig. S2. FTIR spectra of raw sewage sludge digestate (RSS) and biochar from sewage sludge digestate (BSS) produced at 350 °C under slow pyrolysis.

Fig. 1 .

 1 Fig. 1. Arsenic redox distribution in solution during adsorption kinetics for As(III) by BKOH bat (a) and for the control with only released dissolved compounds (RDC) from BKOH bat (b).

  Before As(III) spiking 0.4 b ± 0.1 0.8 b ± 0.1 --After As(III) spiking (12.5 µg) 13.0 ± 0.7 0.8 b ± 0.1 101 101 a Recovery: amount of spiked As(III) recovered by extraction either as As(III) or total arsenic (i.e. As(III) + As(V)). b Value close to limit of quantification.

  .Wongrod, S., Simon, S., van Hullebusch, E.D., Lens, P.N.L., Guibaud, G. (2018). Changes of sewage sludge digestate-derived biochar properties after chemical treatments and influence on As(III and V) and Cd(II) sorption. International Biodeterioration & Biodegradation 135: 96-102.

	XVI
	List of Publications
	I. Wongrod, S., Simon, S., Guibaud, G., Lens, P.N.L., Pechaud, Y., Huguenot, D.,
	van Hullebusch, E.D. (2018). Lead sorption by biochar produced from
	digestates: Consequences of chemical modification and washing. Journal of
	Environmental Management 219: 277-284.
		hydroxide modified biochar from sewage sludge
		digestate with batch washing
	BSS-KOH	Potassium hydroxide modified biochar from sewage sludge
		digestate with batch and subsequent continuous column washing
	CA	Cellulose acetate
	CEC	Cation exchange capacity
	DMA	Dimethylarsinic acid
	DOC	Dissolved organic compounds
	EC	Electrical conductivity
	FTIR	Fourier transform infrared spectroscopy
	GF-AAS	Graphite furnace atomic absorption spectrometry
	ICP-OES	Inductively coupled plasma optical emission spectroscopy
	LC-AFS	Liquid chromatography coupled to atomic fluorescence
		spectrometry
	MMA	Monomethylarsonic acid
	MP-AES	Microwave plasma atomic emission spectroscopy
	OFMSW	Organic fraction of municipal solid waste
	OFMSWD	Organic fraction of municipal solid waste digestate
	PFO	The pseudo-first-order kinetic model
	PSO	The pseudo-second-order kinetic model
	pHPZC	pH of point of zero charge
	RDC	Release of dissolved compounds
	SBET	Brunauer-Emmett-Teller specific surface area
	SSD	Sewage sludge digestate
	XRD	X-ray diffraction

II

III.

[START_REF] Wongrod | Assessing arsenic redox state evolution in solution and solid phase during As(III) sorption onto chemically-treated sewage sludge digestate biochars[END_REF]

.

Assessing arsenic redox state evolution in solution and solid phase during As(III) sorption onto chemically-treated sewage sludge biochars. Bioresource Technology 275: 232-238.

Table 2

 2

	.1: Comparison of maximum sorption capacities for Pb(II) by biochars produced from
	different organic waste materials.
	Biochar material	Pyrolysis	Qm	Experimental conditions:	Reference
		temperature	(µmol g -1 )	initial pH, biochar
				dosage
	Pine wood	300	18.8	pH 5.0, 4 g L -1	Liu & Zhang
					(2009)
	Rice husk	300	8.6	pH 5.0, 4 g L -1	Liu & Zhang
					(2009)
	Peanut shell	350	254.8	pH 5.0, 4 g L -1	Wang et al.
					(2015)
	Whole sugar beet	600	196.9	pH 5.0, 2 g L -1	Inyang et al.
	digestate

Table 2 .

 2

2 compares the sorption abilities for As(III), As(V) and Cd(II) by biochars obtained from different origin of feedstocks. From Table

Table 2 .2: Comparison

 2

	of the sorption performance of As(III), As(V) and Cd(II) by different types
	of biochars.
	Biochar	Element Qm	Concentration	Solution	Reference
			(µmol g -1 )	ranges (µM)	pH
	Rice straw biochar	As(III)	5.9	13-667	-	Wu et al.
						(2017)
	Pine wood biochar	As(III)	16	0-18	3.5	Mohan et al.
						(2007)
	Oak wood biochar	As(III)	78	0-18	3.5	Mohan et al.
						(2007)
	Rice straw biochar	As(V)	7.3	13-667	-	Wu et al.
						(2017)
	Fe(II)-loaded activated	As(V)	27	7-113	3.0	Özge et al.
	carbon					(2013)
	Red mud-modified	As(V)	79	13-667	-	Wu et al.
	biochar from rice straw					(2017)
	Paper mill sludge	As(V)	303	280-2529	6.5	Yoon et al.
	biochar					(2017)
	Magnetic carbonaceous	As(V)	507	10-1335	5.0	Wen et al.
	tea waste					(2017)
	Carbonaceous nanofiber As(V)	670	100-900	5.0	Cheng et al.
						(2016)
	Paper mill sludge	Cd(II)	369	187-2506	6.5	Yoon et al.
	biochar					(2017)
	H2O2-treated yak	Cd(II)	419	10-1779	-	Wang and Liu
	manure biochar					(2018)
	Graphene oxide	Cd(II)	945	40-450	6.0	Zhao et al.
	nanosheet					(2011)

Table 4

 4

	.1: Overview of the pyrolysis conditions (temperature and residence time) applied to
	solid digestate samples for biochar production (OFMSWD: organic fraction of municipal solid
	waste digestate and SSD: sewage sludge digestate).
	Samples	Temperature (°C)	Residence time (min)	Paper
	OFMSWD	400	60	I
	SSD	350	15	I, II, III

Table 4 .2: Overview

 4

	of chemical treatments and washing procedure of biochar to study the
	sorption ability for Pb(II), Cd(II) and As(III, V) and the influence on its redox state distribution in
	solution and on the biochar during As(III) sorption.

Sample Chemical treatment/ washing procedure Label Metal(loid) element Paper

	BMSW	unmodified / batch	BMSW bat	Pb(II)	I
		H2O2 / batch	BMSW-H2O2 bat
		KOH / batch	BMSW-KOH bat
	BSS	unmodified / batch	BSS bat	Pb(II)	I
		H2O2 / batch	BSS-H2O2 bat
		H2O2 / batch+column	BSS-H2O2
		KOH / batch	BSS-KOH bat
		KOH / batch+column	BSS-KOH
	BSS	unmodified / batch	BSS bat	Cd(II) and As(III, V)	II, III
		unmodified / batch+column	BSS
		H2O2 / batch	BSS-H2O2 bat
		H2O2 / batch+column	BSS-H2O2
		KOH / batch	BSS-KOH bat
		KOH / batch+column	BSS-KOH

Table 4 .3: Analytical

 4 methods and instruments used for biochar characterization in Papers I-III.

	Parameter	Analytical methods and instruments	Paper
	pH	pH-meter LPH 330T, Tacussel, France	I, II, III
	pH of point of zero	Zetametry by the zeta potential at different pH ranges	II, III
	charge (pHPZC)
	Zeta potential	Zetasizier Nanoseries (Nano-ZS), Malvern, UK	II, III
	Electrical	CDM 210 conductivity-meter, Radiometer, Denmark	I, II
	conductivity (EC)
	Dissolved organic	AFNOR standard method (NF EN 1484 Juillet 1997;	I
	carbon (DOC)	AFNOR, 1997a), TOC Analyzer (multi N/C ® 2100S,
		Analytikjena, Germany)
	Phosphate (PO4 3 ⁻)	AFNOR colorimetric method (NF EN 1189 Janvier	I, II
		1997; AFNOR, 1997b), Vis spectrophotometer
		(DR1900, Hach, France) at λ 700 nm
	(Bi)carbonate	ARNOR titrimetric method (NF EN ISO 9963-1 Février	I
	(HCO3⁻/CO3 2 ⁻)	1996; AFNOR, 1996) using 10⁻ 3 M HCl as a titrant
		and phenolphthalein and mixed bromocresol green-
		methyl red as indicators.
	Cation exchange	Cobalt hexamine trichloride method (Aran et al.,	II, III
	capacity (CEC)	2008), Präzisions-pH-meter E510, Metrohm AG,
		Switzerland and spectrometer

Table 4 .3: Analytical

 4 methods and instruments used for biochar characterization in Papers I-III.

	Parameter	Analytical methods and instruments	Paper
	Specific surface	Brunauer-Emmett-Teller (BET) N2 sorption method at	II, III
	area	77 K (3Flex, Micromeritics, USA)
	Crystalline structure	X-ray diffraction (XRD), a powder diffractometer (AXS	I, II
		D8, Bruker, Germany)
	Surface functional	Fourier transform infrared spectroscopy (FTIR),	I, II
	groups	Shimadzu IRAffinity-1 Spectrometer with a
		deuterated-triglycine sulfate (DTGS) detector
	Total metals	Acid digestion (EPA Method 3050B; EPA, 1996),	I, II
	(i.e. Al, Ca, Cd, Cr,	inductively coupled plasma optical emission
	Cu, K, Ni, Pb, Mg,	spectroscopy (ICP-OES) (Optima 8300, PerkinElmer,
	Mn, Na, Fe and Zn)	France) or microwave plasma atomic emission
		spectroscopy (MP-AES) (Agilent 4100, Agilent
		Technologies Inc., USA)
	Total As	Graphite furnace atomic absorption spectrometry	II, III
		(GF-AAS) (240Z, Agilent Technologies, USA)
	Arsenic speciation	Liquid chromatography coupled to atomic	II, III
		fluorescence spectrometry (LC-AFS) with hydride
		generation (HG) (PS Analytical Millennium Excalibur,
		PS Analytical, UK)

Table 5 . 1 :

 51 Properties of the organic fraction of municipal solid waste digestate (OFMSWD), its derived biochar (BMSW), and sewage sludge digestate (SSD) and its derived biochar (BSS) before and after pyrolysis.

	Parameters	OFMSWD	BMSW	SSD	BSS
	Pyrolysis conditions	-	400 °C, 1 h	-350 °C, 15 min
	pH in water	8.0 ± 0.1 a	8.9 ± 0.1	6.0 ± 0.1	6.5 ± 0.1
	Biochar yield (%wt)	na b	74 ± 1	na	na
	Ash (%wt)	52 ± 1	68 ± 1	na	na
	Moisture content (%wt)	81 ± 1 c	na	na	na
	Al (µmol g⁻ 1 )	667 ± 37	519 ± 37	408 ± 37	519 ± 37
	As (µmol g⁻ 1 ) b	dl d	dl	1.05 ± 0.04	0.63 ± 0.02
	Ca (µmol g⁻ 1 )	2.22 ± 0.05	2.87 ± 0.03	0.50 ± 0.03	0.75 ± 0.03
	Cd (µmol g⁻ 1 )	dl	dl	dl	0.08 ± 0.01
	Cr (µmol g⁻ 1 )	1.81 ± 0.02	1.92 ± 0.04	0.81 ± 0.02	0.92 ± 0.02
	Cu (µmol g⁻ 1 )	4.11 ± 0.11	6.86 ± 0.08	6.14 ± 0.16	9.71 ± 0.11
	Fe (µmol g⁻ 1 )	197 ± 18	287 ± 18	1182 ± 18	1182 ± 18
	K (µmol g⁻ 1 )	818 ± 26	563 ± 26	102 ± 26	102 ± 26
	Mg (µmol g⁻ 1 )	127 ± 18	182 ± 18	91 ± 18	146 ± 18
	Mn (µmol g⁻ 1 )	5.88 ± 0.06	7.59 ± 0.06	8.68 ± 0.13	14.02 ± 0.15
	Na (µmol g⁻ 1 )	1392 ± 44	1392 ± 44	174 ± 44	217 ± 44
	Ni (µmol g⁻ 1 )	0.73 ± 0.09	0.65 ± 0.05	0.43 ± 0.02	0.70 ± 0.02
	Pb (µmol g⁻ 1 )	0.49 ± 0.03	0.63 ± 0.06	0.39 ± 0.05	0.41 ± 0.02
	Zn (µmol g⁻ 1 )	10.60 ± 0.20 13.58 ± 0.19	11.56 ± 0.16	15.56 ± 0.20

a Values reported as means (n=3) followed by standard deviation. b na refers to not available. c Values reported on dry basis. d dl refers to below detection limit (i.e. 50 µg L⁻ 1 for As and 5 µg L⁻ 1 for Cd).

Table 5 . 2

 52

	Biochar	Total As	pH in	pHpzc	ZP a at	Electrical	SBET b	CEC c
		(µmol g⁻ 1 )	water		pH 5	conductivity	(m 2 g⁻ 1 )	(cmol +
					(mV)	(µS cm⁻ 1 )		kg⁻ 1 )
	BSS	0.63 ± 0.02 6.4 ± 0.1	2.7	-9.8	4.0 ± 0.1	0.4 ± 0.1	2.0 ± 0.1
	BSS-H2O2	0.39 ± 0.02 6.5 ± 0.1	2.9	-16.5	4.1 ± 0.3	5.7 ± 0.1	2.9 ± 0.1
	BSS-KOH	0.17 ± 0.02 8.4 ± 0.1	3.4	-8.1	6.2 ± 0.5	7.9 ± 0.1	13.4 ± 0.1
	BSS-KOH bat	0.31 ± 0.03 10.0 ± 0.1	2.9	-17.1	324.0 ± 2.4	3.0 ± 0.1	20.8 ± 0.1

: Chemical and physical properties of sewage sludge digestate-derived biochar and its chemically-modified biochars (Papers I, II and III).

a ZP refers to zeta potential. b SBET refers to Brunauer-Emmett-Teller surface area of biochar. c CEC refers to cation exchange capacity of biochar.

Table 5 .3: Constants

 5

	Biochar

for the pseudo-first-order (PFO) and pseudo-second-order (PSO) kinetic model fittings for Pb(II) sorption by different types of biochars (Paper I).

a Qe,exp and Qe refer to equilibrium adsorption capacity obtained from the experiments and from model fittings, respectively. b na refers to not available due to no satisfactory fit by the model.

Table 5 . 4

 54

	Biochar					Cd(II)
		PFO			PSO			Intraparticle diffusion
		k1	Qe	R 2	k2	Qe	R 2	ki	C	R 2
		(h⁻ 1 )	(µmol		(g(µmol	(µmol		(µmol (g
			g⁻ 1 )		h)⁻ 1 )	g⁻ 1 )		h 1/2 )⁻ 1 )
	BSS	0.184 0.94	0.923 1.724	1.93	0.983 0.269	0.956	0.852
					± 0.204	± 0.04		± 0.039	± 0.101
	BSS-H2O2 0.400 0.14	0.881 12.251	2.09	0.985 0.042	1.932	0.787
					± 0.803	± 0.01		± 0.010	± 0.029
	BSS-KOH	1.217 0.19	0.763 12.907	2.14	0.993 0.041	1.973	0.699
					± 0.479	± 0.01		± 0.013	± 0.036
	BSS, BSS-H2O2 and BSS-KOH refer to raw biochar from sewage sludge digestate, and its H2O2 and
	KOH modified biochars with batch and subsequent column washings.
	All sorption kinetic constants and uncertainties for Cd(II) were obtained by non-linear regression using
	Statistica software (v6.1, StatSoft).

: Constants for the pseudo-first-order (PFO), pseudo-second-order (PSO) and intraparticle diffusion kinetic models of Cd(II) sorption by raw biochar from sewage sludge digestate and its chemically-modified biochars (Paper II).

Table 5 . 5 :

 55 Constants for the pseudo-first-order (PFO), pseudo-second-order (PSO) and intraparticle diffusion kinetic models of As(V) sorption by raw biochar from sewage sludge digestate and its chemically-modified biochars (Paper II).

	Biochar					As(V)
		PFO			PSO			Intraparticle diffusion
		k1	Qe	R 2	k2	Qe	R 2	ki	C	R 2
		(h⁻ 1 )	(µmol		(g(µmol	(µmol		(µmol (g
			g⁻ 1 )		h)⁻ 1 )	g⁻ 1 )		h 1/2 )⁻ 1 )
	BSS	0.167 0.40	0.918 0.630	0.62	0.985 0.115	0.065	0.910
					± 0.098	± 0.03		± 0.013	± 0.035
	BSS-H2O2 0.188 0.44	0.948 0.619	0.66	0.985 0.121	0.083	0.913
					± 0.092	± 0.03		± 0.014	± 0.036
	BSS-KOH 0.165 1.84	0.969 0.052	2.35	0.995 0.396	-0.074	0.993
					± 0.006	± 0.09		± 0.012	± 0.030

BSS, BSS-H2O2 and BSS-KOH refer to raw biochar from sewage sludge digestate, and its H2O2 and KOH modified biochars with batch and subsequent column washings, respectively. All sorption kinetic constants and uncertainties for As(V) were obtained by non-linear regression using Statistica software (v6.1, StatSoft).

Table 5 . 6 :

 56 Constants for the pseudo-first-order (PFO), pseudo-second-order (PSO) and intraparticle diffusion kinetic models of Cd(II), As(III) and As(V) sorption by KOH-modified biochars from sewage sludge digestate with different biochar washing procedures (Paper II).

	Biochar

BSS-KOH bat refers to KOH modified biochar from sewage sludge digestate (SSD) with batch washing. BSS-KOH refers to KOH modified biochar from SSD with batch and subsequent column washings. All sorption kinetic constants and uncertainties for Cd(II) and As(V) were obtained by non-linear regression using Statistica software (v6.1, StatSoft).

Table 5 .

 5

	Biochar	Langmuir			Pb(II) Freundlich
		KL	Qm	R 2	KF	n	R 2
		(L µmol⁻ 1 )	(µmol g⁻ 1 )		((µmol g⁻ 1 )(L µmol⁻ 1 ) 1/n )
	BMSW bat	0.002	357.1	0.993	2.557	1.585	0.940
	BMSW-H2O2 bat 0.016	434.7	0.997	25.557	2.555	0.896
	BMSW-KOH bat 0.006	500.0	0.998	10.240	1.875	0.860
	BSS bat	0.014	31.4	0.994	7.802	5.647	0.782
	BSS-H2O2 bat	0.004	121.9	0.984	5.610	2.623	0.956
	BSS-KOH bat	0.001	714.2	0.294	1.991	1.472	0.566
	BMSW

7: Constants for Langmuir and Freundlich isotherm model-fittings of Pb(II) sorption by raw and chemically-modified digestate biochars with batch washing (Paper I). bat , BMSW-H2O2 bat and BMSW-KOH bat refer to raw biochar from the OFMSWD, and its H2O2

Table 5 . 8 :

 58 Constants for the Langmuir and Freundlich isotherm models for Cd(II) and As(V) sorption by raw and chemically-modified biochars from sewage sludge digestate (Paper II).

					Cd(II)
	Biochar	Langmuir KL	Qm	R 2	Freundlich KF	n	R 2
		(L µmol⁻ 1 )	(µmol g⁻ 1 )		((µmol g⁻ 1 )(L µmol⁻ 1 ) 1/n )
	BSS	0.061 ± 0.009	15.49 ± 0.32	0.992 3.163	4.336 0.796
	BSS-H2O2	0.005 ± 0.001	218.73 ± 6.91	0.990 5.893	2.008 0.942
	BSS-KOH	0.032 ± 0.004	306.09 ± 6.07	0.994 23.460	2.570 0.933
					As(V)
	Biochar	Langmuir KL	Qm	R 2	Freundlich KF	n	R 2
		(L µmol⁻ 1 )	(µmol g⁻ 1 )		((µmol g⁻ 1 )(L µmol⁻ 1 ) 1/n )
	BSS	0.015 ± 0.006	1.60 ± 0.23	0.958 0.112	2.185 0.986
	BSS-H2O2	0.012 ± 0.003	7.22 ± 0.67	0.989 0.156	1.456 0.966
	BSS-KOH	0.014 ± 0.003	8.11 ± 0.53	0.994 0.196	1.473 0.967

BSS, BSS-H2O2, and BSS-KOH refer to raw biochar from sewage sludge digestate, and its H2O2 and KOH modified biochars, respectively, with batch and subsequent column washings. All sorption kinetic constants and uncertainties for Cd(II) and As(V) were obtained by non-linear regression using Statistica software (v6.1, StatSoft).

Table 5 .9: Effect

 5

	Biochar	Langmuir			Pb(II) Freundlich
		KL	Qm	R 2	KF	n	R 2
		(L µmol⁻ 1 )	(µmol g⁻ 1 )		(µmol g⁻ 1 )(L µmol⁻ 1 ) 1/n
	BSS-H2O2 bat	0.004	121.9	0.984	5.610	2.623	0.956
	BSS-H2O2	0.007	58.1	0.994	6.030	3.527	0.967
	BSS-KOH bat	0.001	714.2	0.294	1.991	1.472	0.566
	BSS-KOH	0.007	386.6	0.987	38.00	3.462	0.926
	bat refers to biochar washing in batch system with ultrapure water (triplicate in a row).

of biochar washing procedures (batch versus batch and subsequent continuous washings) on the Pb(II) sorption by chemically-modified biochars from sewage sludge digestate fitted with Langmuir and Freundlich isotherm models (Paper I).

Table 5 .10:

 5 Constants for the Langmuir and Freundlich isotherm models for Cd(II), As(III) and As(V) sorption by KOH-modified biochars from sewage sludge digestate with different biochar washing steps (Paper II).

					Cd(II)
	Biochar	Langmuir KL	Qm	R 2	Freundlich KF	n	R 2
		(L µmol⁻ 1 )	(µmol g⁻ 1 )		((µmol g⁻ 1 )(L µmol⁻ 1 ) 1/n )
	BSS-KOH bat 0.009 ± 0.001 372.49 ± 6.71 0.996 9.013	1.849 0.882
	BSS-KOH	0.032 ± 0.004 306.09 ± 6.07 0.994 23.460	2.570 0.933
					As(III)
	Biochar	Langmuir KL	Qm	R 2	Freundlich KF	n	R 2
		(L µmol⁻ 1 )	(µmol g⁻ 1 )		((µmol g⁻ 1 )(L µmol⁻ 1 ) 1/n )
	BSS-KOH bat 0.088 ± 0.007 7.60 ± 0.12	0.996 1.007	2.420 0.880
	BSS-KOH	0.103 ± 0.018 8.74 ± 0.33	0.976 1.059	2.299 0.884
					As(V)
	Biochar	Langmuir KL	Qm	R 2	Freundlich KF	n	R 2
		(L µmol⁻ 1 )	(µmol g⁻ 1 )		((µmol g⁻ 1 )(L µmol⁻ 1 ) 1/n )
	BSS-KOH bat 0.024 ± 0.008 0.85 ± 0.09	0.965 0.044	1.867 0.906
	BSS-KOH	0.014 ± 0.003 8.11 ± 0.53	0.994 0.196	1.473 0.967
	BSS-KOH bat refers to KOH modified biochar from sewage sludge digestate (SSD) with batch washing.
	BSS-KOH refers to KOH modified biochar from SSD with batch and subsequent column washings.
	All sorption kinetic constants and uncertainties for Cd(II) and As(V) were obtained by non-linear
	regression using Statistica software (v6.1, StatSoft).

Table 5 .11:

 5 Comparison of total amount of sorbed arsenic determined after acid digestion or extraction procedure (Paper III).

	Biochar	Total arsenic (µmol)	Recovery from extraction (%)
		Extraction	Acid digestion
	BSS	0.002 ± 0.004	0.013 ± 0.013	nq a
	BSS-H2O2	0.040 ± 0.013	0.080 ± 0.013	nq
	BSS-KOH	0.934 ± 0.040	1.241 ± 0.040	75 ± 4
	BSS-KOH bat	0.840 ± 0.053	0.800 ± 0.040	105 ± 9

Table 5 .

 5 12: Comparison of As speciation and total As in exposition solutions and sorbed onto biochars (initial As(III) amount in exposition

	solution: 1.949 µmol) (Paper III).
	Biochar	Exposition solution			Sorbed As onto biochar
		Initial	Final			Acid digestion		Extraction
		As(III)	As(III)+As(V)	As(III)	As(V)	Total As	Sorption yield	As(III)+As(V)	As(III)	As(V)
		(µmol) (µmol)	(%)	(%)	(µmol)	(%)	(µmol)	(%)	(%)
	BSS	1.949	1.935 ± 0.027	99	1	0.013 ± 0.013 <1	0.002 ± 0.004	na	na
	BSS-H2O2	1.949	1.908 ± 0.040	96	4	0.080 ± 0.013 4	0.040 ± 0.013	76	24
	BSS-KOH	1.949	0.640 ± 0.040	98	2	1.241 ± 0.040 64	0.934 ± 0.040	90	10
	BSS-KOH bat	1.949	0.961 ± 0.067	72	28	0.800 ± 0.040 41	0.840 ± 0.053	92	8

na refers to not available due a value close to limit of quantification.

  The semi-continuously washed H 2 O 2 and KOH-treated biochars are labeled as MSW-H 2 O 2

	sem , MSW-
	KOH sem , SS-H 2 O 2

sem and SS-KOH sem for RMSW and SS based biochar, respectively. The continuously washed H 2 O 2 and KOH-treated sewage sludge biochars are labeled as SS-H 2 O 2 con and SS-KOH con , respectively.

Table 2

 2 The

	Biochar	pH in water	S BET	a À m 2 g À1 Á
	SS-H 2 O 2 sem	6.2 ± 0.1	3.6 ± 0.1
	SS-KOH sem	10.1 ± 0.1	3.0 ± 0.1
	SS-H 2 O 2 con	6.5 ± 0.1	5.7 ± 0.1
	SS-KOH con	8.4 ± 0.1	7.9 ± 0.1

pH and specific surface area of chemically-modified sewage sludge digestate biochar with semi-continuous washing (SS-H 2 O 2 sem and SS-KOH sem ) and with continuous washing (SS-H 2 O 2 con and SS-KOH con ).

a S BET refers to BrunauereEmmetteTeller (BET) surface area of biochar.

Table 3

 3 Effect of biochar modification (H 2 O 2 and KOH) and biochar washing (semi-continuous and continuous) on the Pb adsorption capacity.

	Biochar	Langmuir isotherm model	Freundlich isotherm model

Table S1 . Comparison of maximum adsorption capacities for lead by biochar produced from different waste materials.

 S1

	Biochar material	Pyrolysis	Qm (mg g⁻ 1 ) Isotherm well-fitted model	Expriemental conditions:	Reference
		temperature			initial pH, biochar dosage
	Peanut shell	350	52.8	Freundlich, Langmuir	pH 5, 4 g L⁻ 1	Wang et al. (2015)
	Medicine material residues	350	46.1	Langmuir	pH 5, 4 g L⁻ 1	Wang et al. (2015)
	Pine wood	300	3.9	Langmuir	pH 5, 4 g L⁻ 1	Liu & Zhang (2009)
	Rice husk	300	1.8	Langmuir	pH 5, 4 g L⁻ 1	Liu & Zhang (2009)
	Digested dairy waste	600	51.4	-	2 g L⁻ 1	Inyang et al. (2012)
	Digested whole sugar beet	600	40.8	Langmuir	2 g L⁻ 1	Inyang et al. (2012)
	Sewage sludge digestate	350	6.5	Langmuir	pH 5, 4 g L⁻ 1	This study
	Organic fraction of municipal	400	72.9	Langmuir	pH 5, 4 g L⁻ 1	This study
	solid waste digestate

Table S2 . Chemical characteristics of the organic fraction of municipal solid waste digestate (RMSW), its derived biochar (MSW sem ), sewage sludge digestate (RSS) and its derived biochar (SS sem ).

 S2

	Parameters	RMSW	MSW sem	RSS	SS sem
	Biochar yield (%wt)	NA a	74 ± 1	NA	NA
	Ash (%wt)	52 ± 1	68 ± 1	NA	NA
	Moisture content (%wt)	81 ± 1	NA	NA	NA
	As (mg kg⁻ 1 ) b	dl c	dl	79 ± 3 d	34 ± 12
	Cd (mg kg⁻ 1 )	dl	dl	dl	9 ± 1
	Cr (mg kg⁻ 1 )	94 ± 1	100 ± 2	42 ± 1	48 ± 1
	Cu (mg kg⁻ 1 )	261 ± 7	436 ± 5	390 ± 10	617 ± 7
	Ni (mg kg⁻ 1 )	43 ± 5	38 ± 3	25 ± 1	41 ± 1
	Pb (mg kg⁻ 1 )	101 ± 5	130 ± 12	80 ± 9	85 ± 3
	Zn (mg kg⁻ 1 )	693 ± 13	888 ± 12	756 ± 10	1017 ± 13

a NA refers to not available. b Values reported on dry basis.

sem , SS-KOH sem , MSW sem , MSW-H2O2 sem and MSW-KOH sem .

			Pseudo-first-order (PFO)	Pseudo-second-order (PSO)
	Sample
		Qe,exp	Qe	k1	R 2	Qe	k2	R 2
	SS sem	3.4	3.5	0.536 0.146	3.5	0.270	0.997
	SS-H2O2 sem	8.2	8.8	0.633 0.653	8.5	0.137	0.999
	SS-KOH sem	13.7	13.9	1.780	NF a	14.0	1.057	0.999
	MSW sem	23.7	24.2	4.287 0.990	23.8	0.735	0.999
	MSW-H2O2 sem	24.4	24.5	10.177 0.856	24.5 18.496	1.000
	MSW-KOH sem	23.8	24.0	7.766	0.78	23.8	3.528	1.000

Table 1

 1 pH, electrical conductivity and surface area of the raw and chemically-modified biochars.

	Biochar	pH in water	pH pzc	Electrical conductivity (EC) (μS cm -1 )	S BET a (m 2 g -1 )	CEC b (cmol + kg -1 )	Reference
	BSS c	6.4 ± 0.1	2.7	4.0 ± 0.1	0.4 ± 0.1	2.0 ± 0.1	This study
	c BKOH c BH 2 O 2	6.5 ± 0.1 8.4 ± 0.1	2.9 3.4	4.1 ± 0.3 6.2 ± 0.5	5.7 ± 0.1 7.9 ± 0.1	2.9 ± 0.1 13.4 ± 0.1	This study This study
	BKOH batc	10.0 ± 0.1	2.9	324.0 ± 2.4	3.0 ± 0.1	20.8 ± 0.1	This study
	Sewage sludge (300 °C)	6.7 ± 0.2	na d	7750.0 ± 160.0	14.3 ± 0.2	na	Yuan et al. (2015)
	Sewage sludge (300 °C)	na	na	na	4.0	na	Agrafioti et al. (2013)
	a S BET refers to Brunauer-Emmett-Teller surface area of biochar. b CEC refers to cation exchange capacity of biochar. c BSS: biochar from sewage sludge digestate, BH 2 O 2 : H 2 O 2 -modified biochar, BKOH and BKOH bat : KOH-modified biochars, respectively, with batch and column
	washing, and with batch washing.
	d na refers to not available.

  BSS: biochar from sewage sludge digestate, BH 2 O 2 : H 2 O 2 -modified biochar, BKOH and BKOH bat : KOH-modified biochars, respectively, with batch and column washing, and with batch washing.

	BSS	-	-	-	0.115 ± 0.013	0.065 ± 0.035	0.910	0.269 ± 0.039	0.956 ± 0.101	0.852
	BH 2 O 2	-	-	-	0.121 ± 0.014	0.083 ± 0.036	0.913	0.042 ± 0.010	1.932 ± 0.029	0.787
	BKOH	0.429 ± 0.044 0.981 ± 0.113	0.938	0.396 ± 0.012	-0.074 ± 0.030	0.993	0.041 ± 0.013	1.973 ± 0.036	0.699
	BKOH bat	0.358 ± 0.033 0.219 ± 0.086	0.948	0.050 ± 0.006	0.036 ± 0.015	0.919	0.077 ± 0.018	1.799 ± 0.046	0.749

a k 2 unit: g(μmol h) -1 . b Q e unit: in μmol g -1 . c d k i unit: μmol (g h 1/2 ) -1 .

  BSS: biochar from sewage sludge digestate, BH 2 O 2 : H 2 O 2 -modified biochar, BKOH and BKOH bat : KOH-modified biochars, respectively, with batch and column washing, and with batch washing.

					.015 ± 0.006	1.60 ± 0.23	0.958	0.061 ± 0.009	15.49 ± 0.32	0.992
	c BKOH c BH 2 O 2	-0.103 ± 0.018	-8.74 ± 0.33	-0.976	0.012 ± 0.003 0.014 ± 0.003	7.22 ± 0.67 8.11 ± 0.53	0.989 0.994	0.005 ± 0.001 0.032 ± 0.004	218.73 ± 6.91 306.09 ± 6.07	0.990 0.994
	BKOH batc	0.088 ± 0.007	7.60 ± 0.12	0.996	0.024 ± 0.008	0.85 ± 0.09	0.965	0.009 ± 0.001	372.49 ± 6.71	0.996
	a K L unit: in l μmol -1 .

b Q m unit: in μmol g -1 . c

Table S1 .

 S1 Comparison of the sorption performance of As(V) and Cd(II) by different types of biochars.

	Adsorbent	Element	Maximum	adsorption	Concentration	Solution	References
			capacity (µmol g -1 )	ranges (µM)	pH
	BSS a	As(V)	1.6		10 -300	5.0	This study
	BH2O2 a	As(V)	7.1		10 -300	5.0	This study
	BKOH a	As(V)	8.5		10 -300	5.0	This study
	Fe(II)-loaded activated carbon	As(V)	27		7 -113	3.0	Özge et al. (2013)
	Paper mill sludge biochar	As(V)	303		280 -2529	6.5	Yoon et al. (2017)
	Magnetic carbonaceous tea waste	As(V)	507		10 -1335	5.0	Wen et al. (2017)
	Carbonaceous nanofibers	As(V)	670		100 -900	5.0	Cheng et al. (2016)
	BSS	Cd(II)	16		10 -4000	5.0	This study
	BH2O2	Cd(II)	220		10 -4000	5.0	This study
	BKOH	Cd(II)	318		10 -4000	5.0	This study
	Paper mill sludge biochar	Cd(II)	369		187 -2506	6.5	Yoon et al. (2017)
	H2O2-treated yak manure biochar	Cd(II)	419		10 -1779	-	Wang and Liu (2018)
	Graphene oxide nanosheets	Cd(II)	945		40 -450	6.0	Zhao et al. (2011)

a BSS, BH2O2 and BKOH refer to, respectively, biochar produced from sewage sludge digestate, and its derived biochar with H2O2 and KOH treatments.

Table S2 .

 S2 Chemical characteristics of raw sewage sludge digestate (RSS) and its derived biochar produced at 350 °C under slow pyrolysis (BSS).

	Parameters	RSS	BSS
	pH in water	6.0 ±0.1	6.4 ±0.1
	As (mg kg⁻ 1 )	79 ±3 a	47 ±1
	Cd (mg kg⁻ 1 )	dl b	9 ±1
	Cr (mg kg⁻ 1 )	42 ±1	48 ±1
	Cu (mg kg⁻ 1 )	390 ±10	617 ±7
	Ni (mg kg⁻ 1 )	25 ±1	41 ±1
	Pb (mg kg⁻ 1 )	80 ±9	85 ±3
	Zn (mg kg⁻ 1 )	756 ±10	1017 ±13

Table 1

 1 pH, electrical conductivity and surface area of the raw and chemically-modified biochars.

	Biochar	Total As	pH in water	pH pzc	Zeta potential at pH 5	Electrical conductivity (at 20 °C)	S BET a	CEC b
		(µg g -1 )			(mV)	(µS cm -1 )	(m 2 g -1 )	(cmol + kg -1 )
	BSS	47 ± 1	6.4 ± 0.1	2.7	-9.8	4.0 ± 0.1	0.4 ± 0.1	2.0 ± 0.1
	BH 2 O 2	29 ± 1	6.5 ± 0.1	2.9	-16.5	4.1 ± 0.3	5.7 ± 0.1	2.9 ± 0.1
	BKOH	13 ± 1	8.4 ± 0.1	3.4	-8.1	6.2 ± 0.5	7.9 ± 0.1	13.4 ± 0.1
	BKOH bat	23 ± 2	10.0 ± 0.1	2.9	-17.1	324.0 ± 2.4	3.0 ± 0.1	20.8 ± 0.1

Table 2

 2 Arsenic speciation stability using H 3 PO 4 with or without ascorbic acid (BKOH bat sample, with and without As(III) spiking just before extraction).

Table 3

 3 Comparison of total amount of sorbed arsenic determined after acid digestion or extraction procedure.

	Biochar	Total arsenic (µg)	Recovery from extraction (%)
		Extraction	Acid digestion
	BSS	0.2 ± 1	1 ± 1	na a
	BH 2 O 2	3 ± 1	6 ± 1	na
	BKOH	70 ± 3	93 ± 3	75 ± 4
	BKOH bat	63 ± 4	60 ± 3	105 ± 9

Table 4

 4 Comparison of As speciation and total As in exposure solutions and sorbed onto biochars (initial As(III) amount in exposure solution: 146 µg).

	Biochar	Exposure solution			Sorbed As onto biochar
		Initial	Final			Acid digestion		Extraction
		As(III) (µg)	As(III) + As(V) (µg)	As(III) (%)	As(V) (%)	Total As (µg)	Sorption yield (%)	As(III) + As(V) (µg)	As(III) (%)	As(V) (%)
	BSS	146	145 ± 2	99	1	1 ± 1	< 1	0.2 ± 0.3	na a	na
	BH 2 O 2	146	143 ± 3	96	4	6 ± 1	4	3 ± 1	76	24
	BKOH	146	48 ± 3	98	2	93 ± 3	64	70 ± 3	90	10
	BKOH bat	146	72 ± 5	72	28	60 ± 3	41	63 ± 4	92	8

Acknowledgements

Patrice Fondan eche is acknowledged for technical assistance with the Pb measurement. This research project has received funding from the European Union's Horizon 2020 research and innovation programme under the Marie Sklodowska-Curie grant agreement N 643071.

Acknowledgments

We thank Patrice Fondanèche (Université de Limoges, France) for technical assistance with the arsenic and cadmium analysis. This work was financially supported by the Marie Sklodowska-Curie European Joint Doctorate Advanced Biological Waste-to-Energy Technologies (ABWET) under the European Union's Horizon 2020 research and innovation programme (grant agreement N°643071).

Acknowledgments

Patrice Fondanèche (Université de Limoges, France) is acknowledged for technical assistance with arsenic analysis. This research work was economically supported by the Marie Sklodowska-Curie European Joint Doctorate in Advanced Biological Waste-to-Energy (ABWET) under the European Union's Horizon 2020 framework [grant agreement N°643071].

This dissertation summarizes the research work performed at University of Paris-Est (for Water Education (the Netherlands). All the experimental works were financially supported by the Marie Skłodowska-Curie European Joint Doctorate (EJD) Advanced Biological Waste-to-Energy Technologies (ABWET) programme under the European Union Horizon 2020 framework (grant agreement no. 643071).

I would like to thank my promotor, Prof. Eric D. van Hullebusch, for his guidance through my research works and for his supervision during

Author's Contribution

Paper I:

Suchanya Wongrod planned and performed the experiments on biochar characterization and metal analysis and wrote the manuscript. Gilles Guibaud is the corresponding author. Gilles Guibaud and Stéphane Simon helped in data evaluation and revising the manuscript. Eric van Hullebusch, Piet Lens, Yoan Pechaud and David Huguenot contributed to the planning of the experiments as well as the manuscript revision.

Paper II: Suchanya Wongrod performed the experiments, characterized the biochar properties, metal(loid)s analysis, data interpretation and wrote the manuscript. Stéphane Simon helped in arsenic speciation analysis, data evaluation and is the corresponding author. Gilles Guibaud and Stéphane Simon contributed to the planning of experimental work and commented on the manuscript. Eric van Hullebusch and Piet Lens helped in revising the manuscript. XRD analysis showed that MSW sem contained the crystalline structures of kaolinite (Al2Si2O5(OH)4), quartz (SiO2) and calcite (CaCO3) at 2θ 12.38°, 26.68° and 29.51°, respectively (Fig. S1(a)). Similar trends were also found in its chemically-modified biochar This work seeks to extend the knowledge on the effect of chemical treatment of sewage sludge digestate (SSD)derived biochar for the As(III and V) and Cd(II) sorption ability using potassium hydroxide (KOH) or hydrogen peroxide (H 2 O 2 ). Results showed the increases of the pH of point of zero charge, the Brunauer-Emmett-Teller (BET) surface area and cation exchange capacity (CEC) after chemical treatment of biochar. The sorption ability was enhanced from 1.6 μmol g -1 (As(V)) and 16.1 μmol g -1 (Cd(II)) on raw biochar to 8.5 μmol g -1 (As(V)) and 318.5 μmol g -1 (Cd(II)) on KOH-modified biochar. Furthermore, arsenic redox distribution showed a large oxidation (70%) of As(III) to As(V) in KOH-biochar with batch washing, while a partial oxidation (7%) was observed in KOH-biochar with batch and subsequent column washing. The washing procedures after KOH treatment play an important role on arsenic sorption, due to the release of phosphate (PO 4 3-) as well as organic matter from the biochar that may subsequently lead to the oxidation of As(III) to As(V). Our findings highlight the potential influence of biochar on the redox transformation of As(III) to As(V) and therefore requires a careful assessment while investigating the fate of As in aquatic environments.

Paper III:

II

Conflict of Interest

The authors declare no conflict of interest.

Appendix A. Supplementary data

Supplementary data to this article can be found online at https:// doi.org/10.1016/j.ibiod.2018.10.001. Table S4. Constants for the Langmuir and Freundlich isotherms for As(III), As(V) and Cd(II) sorption by different types of biochars (BSS: biochar from sewage sludge digestate, BH2O2: H2O2-modified biochar, BKOH and BKOH bat : KOH-modified biochars, respectively, with batch and column washing and with batch washing).

Supplementary Information

Biochar As(III)

As(V) b Qm unit: in µmol g -1 .

c KF unit: in (µmol g -1 )(l µmol -1 ) 1/n .

III

Introduction

Elevated concentrations of arsenic (As) in water bodies represent a global environmental and health issue because of its toxic features. The increasing threats of As contamination mainly originate from anthropogenic sources, particularly mining, industrial and agricultural activities (Vithanage et al., 2017). Arsenic toxicity in targeted organisms is strongly linked to its chemical speciation, especially its redox state.

Conflict of interest

The authors declare no conflict of interest.