Université de Montréal Résine échangeuse d'ions en mode ...

Université de Montréal

Résine échangeuse d’ions en mode biologique pour l’enlèvement de matières organiques

naturelles des eaux de surface

Par

Zhen Liu

Département de chimie, Faculté des arts et des sciences

Thèse présentée en vue de l’obtention du grade de Doctorat (Ph.D.) en chimie

Août, 2021

© Zhen Liu, 2021

Université de Montréal

Unité académique : Département de chimie, Faculté des arts et des sciences

Cette thèse intitulée

Résine échangeuse d’ions en mode biologique pour l’enlèvement de matières organiques naturelles des eaux de surface

Présenté par

Zhen Liu

A été évaluée par un jury composé des personnes suivantes

Kevin J. Wilkinson Président-rapporteur

Sébastien Sauvé

Directeur de recherche

Benoit Barbeau Codirecteur

Jesse Shapiro

Membre du jury

Jean-François Blais Examinateur externe

I

Résumé

La matière organique naturelle (MON) est omniprésente dans les eaux de surface. Bien que

l’exposition à la MON via l’eau potable soit commune et ne soit pas associée à des effets directs

sur la santé humaine, la MON peut avoir des impacts négatifs sur la production d’eau potable,

tels que la contribution aux goûts et odeurs, le développement du biofilm dans les systèmes de

distribution et la formation de sous-produits de désinfection. La résine échangeuse d’ions en

mode biologique (en anglais : Biological ion exchange, BIEX) est un processus prometteur pour

l’enlèvement de la MON des eaux de surface. Il s’agit d’opérer la résine échangeuse d’ions dans

un réacteur à lit fixe avec une régénération peu fréquente de sorte qu’une communauté

microbienne peut se développer sur la surface de résine et ainsi contribuer à l’enlèvement de la

MON par biodégradation. Néanmoins, les mécanismes de l’enlèvement de la MON dans le BIEX

et la faisabilité de son application dans l’usine de production d’eau potable demeurent inconnus.

Ainsi, l’objectif de cette thèse est 1) de comprendre et favoriser l’application du BIEX pour

l’enlèvement de la MON des eaux de surface et 2) de résumer les stratégies qui peuvent alléger

la gestion de la saumure engendrée par la régénération de résines. Les résines en forme chlorure

et bicarbonate ont été d’abord évaluées pour l’application du BIEX où le pilote de BIEX a été

alimenté par l’eau de surface pendant 9 mois sans régénération. Les résultats ont démontré que

l’échange d’ions est le mécanisme dominant pour le BIEX, i.e., la MON échange avec les ions

préchargés (i.e., chlorure et bicarbonate) et les ions préretenus (i.e., sulfate). En plus, les résines

colonisées ont été prélevées du pilote et testées en laboratoire où les résines colonisées ont été

mises en contact avec les composés de modèles (i.e., micropolluants organiques). Les résultats

ont démontré que la biodégradation contribuait à l’enlèvement de micropolluants organiques sur

les résines colonisées, mais le degré de biodégradation dépend des caractères de micropolluants

organiques et la communauté microbienne sur les résines. Ensuite, le BIEX a été évalué en

parallèle du charbon actif biologique (CAB) en filtration secondaire dans l’usine de production

d’eau potable de Sainte-Rose. Les résultats ont démontré que bien que le BIEX ait réalisé un

enlèvement du carbone organique dissous (COD) plus élevé par rapport à celui du CAB, il a une

II

perte de charge plus significative et le rétrolavage de BIEX s’avère être plus complexe par rapport

à celui de CAB. Finalement, une revue de littérature a été menée afin d’identifier les stratégies

sur l’opération de résine et la gestion de saumure, et ainsi d’alléger la gestion de la saumure

engendrée par la régénération de résines échangeuses d’ions. En somme, cette thèse permet de

comprendre les mécanismes de l’enlèvement de la MON dans le BIEX, évaluer la faisabilité de son

application dans l’usine de production d’eau potable ainsi qu’identifier les stratégies qui peuvent

alléger la gestion de la saumure engendrée par la régénération de résines échangeuses d’ions.

Mots-clés : matière organique naturelle, sous-produits de désinfection, micropolluants, résine

échangeuse d’ions, adsorption, biodégradation, BIEX, gestion de saumure, traitement de l’eau

potable.

III

Abstract

Natural organic matter (NOM) is ubiquitous in surface water. Although the exposure to NOM via

drinking water is common and is not associated with direct effects on human health, NOM can

cause negative impacts on drinking water treatment, such as contribution to taste and odors,

development of biofilms in distribution systems and formation of disinfection by-products.

Biological ion exchange (BIEX) is a promising process for the removal of NOM from surface waters.

It involves operating the ion exchange resin in a fixed bed reactor with infrequent regeneration

so that a microbial community can develop on the resin surface and thus contribute to the

removal of NOM by biodegradation. However, the mechanisms for the removal of NOM in BIEX

and the feasibility of its application in the drinking water plant remain unknown. Therefore, the

general objective of this thesis is 1) to understand and promote the application of BIEX for the

removal of NOM from surface water and 2) to summarize the strategies that can alleviate the

management of the brine generated by the regeneration of resins. Chloride and bicarbonate-form

resins were first evaluated for the BIEX application where the BIEX pilot was fed with surface

water for 9 months without regeneration. The results demonstrated that ion exchange is the

dominant mechanism in BIEX, i.e., NOM exchanges with preloaded ions (i.e., chloride and

bicarbonate) and pre-retained ions (i.e., sulfate). In addition, the colonized resins were harvested

from the pilot and tested in the laboratory where the colonized resins were in contact with the

model compounds (i.e., organic micropollutants). The results demonstrated that biodegradation

contributes to the removal of organic micropollutants on colonized resins, but the extent of

biodegradation depends on the characteristics of the organic micropollutants and the microbial

community on the resins. Then, BIEX was evaluated in parallel with biological activated carbon

(BAC) at the second-stage filtration of the Sainte-Rose drinking water treatment plant. The results

demonstrated that although BIEX achieved higher dissolved organic carbon (DOC) removal

compared to BAC, it had a more significant pressure drop and the backwash of BIEX filters was

proved to be more complex compared to that of BAC. Finally, a literature review was carried out

to identify strategies on resin operation and brine management, and thus alleviate the

management of the brine generated by the regeneration of ion exchange resins. Overall, this

IV

thesis allows understanding the mechanisms for the removal of NOM in BIEX, evaluating the

feasibility of its application in drinking water production plants as well as identifying the strategies

that can alleviate the management of the brine generated by the regeneration of ion exchange

resins.

Keywords: natural organic matter, disinfection by-products, micropollutants, ion exchange

resins, adsorption, biodegradation, BIEX, brine management, drinking water treatment.

V

Table des matières

Résumé ............................................................................................................................... I

Abstract ............................................................................................................................ III

Table des matières ............................................................................................................ V

Liste des tableaux ........................................................................................................... XII

Liste des figures ............................................................................................................. XIV

Liste des sigles et abréviations ..................................................................................... XVIII

Remerciements .............................................................................................................. XXI

Chapitre 1 – Introduction ................................................................................................... 1

1.1 Matière organique naturelle ....................................................................................................... 1

1.1.1 Définition et classement ............................................................................................................................. 1

1.1.2 Impact en eau potable ................................................................................................................................ 4

1.1.3 Mesure et caractérisation ........................................................................................................................... 5

1.1.4 Source et occurrence .................................................................................................................................. 8

1.2 Procédés de traitement pour l’enlèvement de la MON .............................................................. 10

1.2.1 Coagulation ............................................................................................................................................... 11

1.2.2 Adsorption ................................................................................................................................................ 11

1.2.3 Filtration membranaire ............................................................................................................................. 12

1.2.4 Processus d’oxydation avancée ................................................................................................................ 13

1.2.5 Biodégradation ......................................................................................................................................... 14

1.2.6 Échange d’ions .......................................................................................................................................... 16

VI

1.2.7 Résine échangeuse d’ions en mode biologique ........................................................................................ 22

1.3 Problématiques, objectifs, hypothèses et structure de la thèse ................................................. 23

1.3.1 Problématiques ......................................................................................................................................... 23

1.3.2 Objectifs et hypothèses ............................................................................................................................ 24

1.3.3 Structure de la thèse ................................................................................................................................. 25

Références ...................................................................................................................................... 27

Chapitre 2 – Operating bicarbonate-form versus chloride-form ion exchange resins without

regeneration for natural organic matter removal ............................................................. 38

Abstract .......................................................................................................................................... 38

2.1 Introduction .............................................................................................................................. 39

2.2 Methods and materials .............................................................................................................. 41

2.2.1 Resin preconditioning ............................................................................................................................... 41

2.2.2 Pilot location and source water characteristics ........................................................................................ 42

2.2.3 Pilot design and operation ........................................................................................................................ 42

2.2.4 Analytical methods ................................................................................................................................... 42

2.2.5 Calculation of BIEX loadings ...................................................................................................................... 43

2.2.6 Statistical analysis ..................................................................................................................................... 44

2.3 Results and discussion ............................................................................................................... 44

2.3.1 BIEX loading during pilot study ................................................................................................................. 44

2.3.2 NOM removal ........................................................................................................................................... 45

2.3.3 BDOC removal ........................................................................................................................................... 47

2.3.4 THM and HAA precursors removal ........................................................................................................... 49

2.3.5 NOM hydrophobicity ................................................................................................................................ 50

2.3.6 NOM fractionation .................................................................................................................................... 51

VII

2.3.7 Implications for IX operation for NOM removal ....................................................................................... 53

2.4 Conclusions ............................................................................................................................... 54

2.5 Supplementary materials ........................................................................................................... 55

References ...................................................................................................................................... 58

Chapitre 3 – Removal of organic micropollutants from surface waters by biological ion exchange

resins ............................................................................................................................... 64

Abstract .......................................................................................................................................... 64

3.1 Introduction .............................................................................................................................. 65

3.2 Materials and methods .............................................................................................................. 67

3.2.1 Biological ion exchange (BIEX) resins characteristics ................................................................................ 67

3.2.2 Raw water characteristics ......................................................................................................................... 67

3.2.3 Target micropollutants ............................................................................................................................. 67

3.2.4 Batch tests ................................................................................................................................................ 69

3.2.5 Analytical methods ................................................................................................................................... 70

3.2.6 Data analysis ............................................................................................................................................. 70

3.2.6.1 Micropollutant removal kinetics ....................................................................................................... 70

3.2.6.2 Suspect screening of transformation products ................................................................................. 71

3.3 Results and discussion ............................................................................................................... 72

3.3.1 Organic micropollutant concentrations in the raw water ........................................................................ 72

3.3.2 Micropollutant removal during batch tests .............................................................................................. 73

3.3.3 Suspect screening of transformation products ......................................................................................... 78

3.3.4 Implications on the application of ion exchange resins ............................................................................ 80

3.4 Conclusions ............................................................................................................................... 81

VIII

3.5 Supplementary materials ........................................................................................................... 81

References ...................................................................................................................................... 90

Chapitre 4 – Biological ion exchange as an alternative to biological activated carbon for drinking

water treatment .............................................................................................................. 95

Abstract .......................................................................................................................................... 95

4.1 Introduction .............................................................................................................................. 96

4.2 Materials and Methods ............................................................................................................. 98

4.2.1 Pilot location and source water characteristics ........................................................................................ 98

4.2.2 Pilot plant design and operation ............................................................................................................... 99

4.2.3 Analytical methods ................................................................................................................................. 101

4.3 Results ..................................................................................................................................... 102

4.3.1 Head loss accumulation .......................................................................................................................... 102

4.3.2 NOM removal ......................................................................................................................................... 103

4.3.3 Removal of THM and HAA precursors .................................................................................................... 106

4.3.4 Impact of BIEX on inorganic anions ........................................................................................................ 106

4.3.5 Removal of ammonia .............................................................................................................................. 108

4.4 Discussion ................................................................................................................................ 109

4.4.1 Head loss in the BIEX filter ...................................................................................................................... 109

4.4.2 NOM removal mechanisms in the BIEX filter .......................................................................................... 110

4.4.3 Ammonia release in the BIEX filter ......................................................................................................... 114

4.5 Conclusion ............................................................................................................................... 114

4.6 Supplementary materials ......................................................................................................... 115

References .................................................................................................................................... 118

IX

Chapitre 5 – Alleviating the burden of ion exchange brine in water treatment: from operational

strategies to brine management .................................................................................... 123

Abstract ........................................................................................................................................ 123

5.1 Introduction ............................................................................................................................ 123

5.2 Ion exchange brine characteristics ........................................................................................... 125

5.3 Ion exchange operational strategies to facilitate brine management ....................................... 130

5.3.1 Resin selection strategies ....................................................................................................................... 130

5.3.2 Resin contactor configuration ................................................................................................................. 132

5.3.3 Novel ion exchange operational mode ................................................................................................... 132

5.3.4 Novel ion exchange regeneration strategies .......................................................................................... 133

5.3.4.1 Segmented regeneration ................................................................................................................ 133

5.3.4.2 Alternative regenerants to NaCl ..................................................................................................... 134

5.3.4.2.1 Alternative regenerants for cation exchange resins ............................................................... 134

5.3.4.2.2 Alternative regenerants for anion exchange resins ................................................................ 135

5.3.4.3 No-chemical-addition regeneration ................................................................................................ 136

5.3.4.3.1 Biological regeneration ........................................................................................................... 136

5.3.4.3.2 Electrochemical regeneration ................................................................................................. 137

5.3.4.3.3 Thermal regeneration ............................................................................................................. 138

5.4 Ion exchange brine management ............................................................................................. 139

5.4.1 Ion exchange brine reuse ........................................................................................................................ 139

5.4.1.1 Direct reuse ..................................................................................................................................... 139

5.4.1.2 Treatment strategies for reuse ....................................................................................................... 140

5.4.1.2.1 Treatment strategies for hardness-laden brine ...................................................................... 141

5.4.1.2.2 Treatment strategies for arsenic-laden brine .......................................................................... 142

5.4.1.2.3 Treatment strategies for chrome-laden brine ......................................................................... 142

X

5.4.1.2.4 Treatment strategies for NOM-laden brine ............................................................................ 143

5.4.1.2.5 Treatment strategies for trace organic pollutants-laden brine ............................................... 144

5.4.1.2.6 Treatment strategies for nitrate- and perchlorate-laden brine .............................................. 145

5.4.2 Ion exchange brine disposal .................................................................................................................... 148

5.4.2.1 Direct disposal ................................................................................................................................. 149

5.4.2.2 Treatment strategies for disposal ................................................................................................... 150

5.5 Discussion and Conclusions ...................................................................................................... 151

References .................................................................................................................................... 154

Chapitre 6 – Conclusions et perspectives ........................................................................ 170

6.1 Conclusions ............................................................................................................................. 170

6.2 Perspectives ............................................................................................................................ 173

Annexe A – Biological ion exchange as an alternative to biological activated carbon for natural

organic matter removal: impact of temperature and empty bed contact time (EBCT) .... 176

Abstract ........................................................................................................................................ 176

A.1 Introduction ............................................................................................................................ 177

A.2 Materials and Methods ........................................................................................................... 179

A.2.1 Feed water .............................................................................................................................................. 179

A.2.2 Filtration media ...................................................................................................................................... 180

A.2.3 Bench-scale systems ............................................................................................................................... 180

A.2.4 Analytical Methods ................................................................................................................................. 181

A.2.5 Data analysis ........................................................................................................................................... 182

A.3 Results and discussion ............................................................................................................. 183

A.3.1 DOC removal ........................................................................................................................................... 183

XI

A.3.2 Impact of temperature and EBCT on DOC removal ................................................................................ 185

A.3.3 Impact of temperature on DOC removal kinetics ................................................................................... 186

A.3.4 Implication on the operation of biofilters for NOM removal ................................................................. 190

A.4 Conclusion .............................................................................................................................. 191

References .................................................................................................................................... 191

XII

Liste des tableaux

Tableau 1.1 - Fractions de la MON par rapport à la polarité et la charge (Health Canada, 2020). 3

Tableau 1.2 - Carbone organique total (COT) dans les eaux brutes pour différentes provinces et

territoires canadiens (Canada Health, 2020). ................................................................................. 9

Table S2.1 - Raw water physicochemical characteristics (April. 2019 – January. 2020). .............. 56

Table S2.2 - Source water natural organic matter (NOM) fractions expressed in percentage of

dissolved organic carbon (DOC). ................................................................................................... 57

Table 3.1 - Organic micropollutants selected for the present study. ........................................... 68

Table 3.2 - Micropollutant concentrations in the raw water prior to resin harvesting. Values and

error bars respectively correspond to average and standard deviation of monthly measurements.

...................................................................................................................................................... 72

Table 3.3 - Rate constants (k) and coefficients of determination (R2) for the kinetic models of

micropollutant removal during batch tests. Fitting with R2 lower than 0.25 was designated as not

available. ....................................................................................................................................... 77

Table 3.4 - Estimated concentrations of theobromine and 1-hydroxyibuprofen during the batch

test. ............................................................................................................................................... 79

Table S3.1 - Raw water characteristics ......................................................................................... 82

Table S3.2 - Potential transformation products of selected micropollutants based on the

literature. ...................................................................................................................................... 85

Table 4.1 - Influent water characteristics (average ± standard deviation) during the pilot study

period (April 11, 2018 to January 07, 2019) ................................................................................. 99

Table 4.2 - Media characteristics for IX, three different GAC and BAC used in this study ......... 101

Table 5.1 - Physicochemical characteristics of ion exchange brine following a single regeneration

.................................................................................................................................................... 127

Table A1 - Feed water physicochemical characteristics. Values and confidence interval

respectively correspond to average and standard deviation of measurements during the study

period. ......................................................................................................................................... 179

XIII

Table A2 - Normalized DOC removal rate constants, normalized residual DOC concentrations and

temperature activity coefficients for BIEX and BAC filters. The confidence interval as well as values

in the paratheses corresponds to the standard error of the estimated parameters. ................ 188

XIV

Liste des figures

Figure 1.1 - Structures moléculaires hypothétiques pour l’acide humique et l’acide fulvique

(Stevenson, 1982; Buffle, 1977). ..................................................................................................... 2

Figure 1.2 - Aperçu des processus de traitement pour l’enlèvement de la MON. ....................... 11

Figure 1.3 - Configuration de réacteur pour la résine échangeuse d’ions. (A) : Procédé MIEX; (B) :

Procédé SIX; (C) Procédé FIX; (D) Contacteur à lit fixé. ................................................................. 19

Figure 2.1 - Cumulative loading (eq/L of resin) in the (A) bicarbonate-form BIEX (BIEX-B) and (B)

chloride-form BIEX (BIEX-C) for four anions and dissolved organic carbon (DOC) throughout the

study period (≈ 24 000 BV or 250 days). No regeneration occurred during this period. ............ 45

Figure 2.2 - (A) Normalized effluent dissolved organic carbon concentrations (DOC/DOC0) of

bicarbonate-form BIEX (BIEX-B) and chloride-form BIEX (BIEX-C) filters throughout the study

period (≈ 24 000 BV or 250 days). (B) conceptional displacement of different solutes in the BIEX-

B filter in primary ion exchange, secondary ion exchange and exhaustion phases. NOM was broken

up into three conceptional fractions with an affinity sequence of NOM3>SO42->NOM2>HCO3-

>NOM1. ......................................................................................................................................... 47

Figure 2.3 - Biodegradable dissolved organic carbon (BDOC) removal in bicarbonate-form BIEX

(BIEX-B) and chloride-form BIEX (BIEX-C) filters. .......................................................................... 48

Figure 2.4 - Natural organic matter (NOM) reactivities (expressed as µg disinfection by-products

per mg Cl2 consumed) in raw water, bicarbonate-form BIEX (BIEX-B) and chloride-form BIEX (BIEX-

C) effluents from 5 300 to 11 000 bed volumes (i.e., secondary IX). ............................................ 50

Figure 2.5 - Normalized effluent specific ultraviolet absorbance at 254 nm (SUVA/SUVA0) of


period (≈ 24 000 BV or 250 days). ............................................................................................... 51

Figure 2.6 - Removals of natural organic matter (NOM) fractions in bicarbonate-form BIEX (BIEX-

B). .................................................................................................................................................. 52

Figure S2.1 - Average heterotrophic biomass measured by 14C glucose respiration test as a

function of media depth and empty bed contact time (EBCT) of the bicarbonate-form BIEX (BIEX-

XV

B) and the chloride-form BIEX (BIEX-C) filters at 10 000 BV. Error bars represent the maximum and

the minimum biomass value of triplicate analysis. ....................................................................... 55

Figure S2.2 - Disinfection by-product concentrations (B) in raw water, chloride-form BIEX (BIEX-

C) and bicarbonate-form BIEX (BIEX-B) effluents from 5 300 to 11 000 bed volumes (i.e., secondary

IX). ................................................................................................................................................. 55

Figure S2.3 - Removals of natural organic matter (NOM) fractions in chloride-form BIEX (BIEX-C).

...................................................................................................................................................... 56

Figure S2.4 - Ammonia nitrogen concentration in raw water, chloride-form BIEX (BIEX-C) and

bicarbonate-form BIEX (BIEX-B) effluents from June to August (4 600 – 12 000 bed volumes).

Water temperature: 13-26 °C. ...................................................................................................... 56

Figure 3.1 - Micropollutants removal during batch tests under biotic and abiotic conditions.

Trendlines correspond to the kinetic model (pseudo-1st or pseudo-2nd order) fitted to the data.

Kinetic parameters of the fitted models are summarized in Table 3. Error bars were omitted due

to overlapping with data points. ................................................................................................... 76

Figure 3.2 – Normalized micropollutant concentrations after 24 h batch test. Error bars

correspond to the minimum and maximum values between test groups and controls. CAF:

caffeine; E2: Estradiol; IBU: ibuprofen; NAP: naproxen; THI: thiamethoxam; ATZ: atrazine; PFOA:

perfluorooctanoic acid; PFOS: perfluorooctanesulfonic acid. ...................................................... 78

Figure S3.1 - Micropollutant recovery of different types of filters. .............................................. 83

Figure S3.2 - Chromatographic separation at m/z 181.07255 (positive mode). ........................... 87

Figure S3.3 - Isotopic pattern at m/z 181.07255 (positive mode). ............................................... 87

Figure S3.4 - Fragmentation spectra at 4.01 min for m/z 181.07255 with specific fragments of

theobromine. ................................................................................................................................ 88

Figure S3.5 - Chromatographic separation at m/z 221.11666 (negative mode). ......................... 89

Figure S3.6 - Isotopic pattern at m/z 221.11666 (negative mode). .............................................. 89

Figure S3.7 - Fragmentation spectra at 5.09 min for m/z 221.11666 with specific fragments of 1-

hydroxyibuprofen. ........................................................................................................................ 90

XVI

Figure 4.1 - Schematic overview of the pilot plant consisting of five parallel columns: BIEX, three

GAC filters and one BAC filter. The pilot is fed with coagulated/settled/filtered/ozonated surface

waters from the Sainte-Rose drinking water treatment plant. .................................................. 100

Figure 4.2 - Head loss accumulation for the BIEX, three GAC and the BAC filter after (i) 30 min of

operation following a backwash (D1), (ii) seven days of operation (D7) and (iii) fourteen days of

operation (D14). ......................................................................................................................... 103

Figure 4.3 - Normalized DOC concentrations (Ceffluent/Cinfluent) in the effluents of the BIEX filter,

three GAC filters and one BAC filter for the study period from April 11, 2018 to January 07, 2019.

.................................................................................................................................................... 104

Figure 4.4 - Distribution of total DOC removal (0.87, 0.76, 0.73, 0.71, 0.52 mg C/L) in the upper

(30 cm) vs. lower layer (150 cm) for BIEX, GAC1, GAC2, GAC3 and BAC, realized at about 21 000

BV. ............................................................................................................................................... 105

Figure 4.5 - Distribution of (A) THM-UFC and (B) HAA5-UFC precursors in the influent and effluents

of the filter media under investigation. Period: 18 000 to 32 000 BV. Numbers indicate the average

concentrations. ........................................................................................................................... 106

Figure 4.6 - Evolution of the anion concentrations on the resin during 34 000 BV. ................... 108

Figure 4.7 - Normalized ammonia concentration in the effluents of the BIEX filter, three GAC filters

and one BAC filter for the study period from April 11, 2018 to January 7, 2019. ...................... 109

Figure 4.8 - Displacement of NOM fractions in the BIEX filter a) virgin IEX; b) NOM3, sulfate and

NOM2 replace chloride while NOM1 is nonexchangeable; c) NOM3 and sulfate replace NOM2

leading to the DOC release in the BIEX filter; d) NOM3 replaces sulfate, which explains the long-

term performance of NOM removal in the BIEX filter. N.B. The anion on each band presents the

dominant species but not the only one. ..................................................................................... 112

Figure S4.1 - Normalized UVA254 (UVAout/UVAin) in the effluents of the BIEX filter, three GAC filters

and one BAC filter for the study period from April 11, 2018 to January 7, 2019. ...................... 115

Figure S4.2 - Ammonia removal as function of media depth (or EBCT), profile realized in warm

waters (week 21, » 20 000 BV, T = 21°C) for BIEX filter, three GAC filters and one BAC filter. .. 116

Figure S4.3 - A thick layer of biomass formed on the top of BIEX media, which proved to be difficult

to breakdown during the backwash. .......................................................................................... 117

XVII

Figure S4.4 - Schematic presentation of a) virgin IX with loosely-packed and uniform morphology

at the beginning of test; 2) inter-chain or intra-chain ion bridging induced by multivalent anions

after long period of operation; ................................................................................................... 118

Figure 5.1 - Ion exchange (IX) plant design workflow integrated with IX operational strategies and

IX brine management. ................................................................................................................ 153

Figure A1 - Bench-scale biofiltration system (single filter illustrated). ....................................... 181

Figure A2 - Typical normalized DOC in the BIEX and BAC filter effluents for the different

temperature considered. Results presented for Port 3 corresponding to an EBCT of 30 min. .. 184

Figure A3 - Average DOC removal in BIEX and BAC filters during steady state (after about 40 days

of operation) for the different conditions investigated. Error bars correspond the standard error

of averages during the steady state period. ............................................................................... 186

Figure A4 - Normalized DOC vs. EBCT at different temperatures. Note that the normalized DOC

data is the same as in Figure A3 but presented with equation A1 fitted to the data. Confidence

intervals were omitted because of overlapping with data symbols. .......................................... 188

Figure A5 - Rate constants vs temperature. Note that the rate constants are the same as those

listed in table 2 but presented with equation 2 fitted to the data. Confidence intervals were

omitted because of overlapping with data symbols. .................................................................. 190

XVIII

Liste des sigles et abréviations

AIX : Anion exchange resins

ATP : Adenosine triphosphate

BB : Building blocks

BDOC : Biodegradable dissolved organic carbon

BIEX : Biological ion exchange

BP : Biopolymers

CIX : Cation exchange resins

DBP : Disinfection by-products

DOC : Dissolved organic carbon

EBCT : Empty bed contact time

EEM : Excitation-emission matrice

HAA : Haloacetic acids

HS : Humics

IX : Ion exchange resins

LC-OCD : Liquid chromatography – organic carbon detector

LMW acids : low molecular weight acids

LMW neutrals : low molecular weight neutrals

MWCO : Molecular weight cut-off

MIEX : Magnetic ion exchange

NOM : Natural organic matter

XIX

SBA : Strong base anion exchange

SUVA : Specific ultraviolet absorbance

THM : Trihalomethane

UV : Ultraviolet

WBA : Weak base anion exchange

WTP : Water treatment plants

CAB : Charbon actif biologique

CAG : Charbon actif en grain

CAP : Charbon actif en poudre

COA : Carbone organique assimilable

COD : Carbone organique dissous

CODB : Carbone organique dissous biodégradable

COP : Carbone organique particulaire

COT : Carbone organique total

DCO : Demande chimique en oxygène

MOB : Matière organique biodégradable

MON : Matière organique naturelle

MF : Microfiltration

NF : Nanofiltration

OI : Osmose inverse

SPD : Sous-produits de désinfection

UF : Ultrafiltration

XX

À mon épouse Xiameng Feng, ma mère Xiuzhi Zhang et mon père Yuping Liu.

XXI

Remerciements

Tout d’abord, je tiens à exprimer mes reconnaissances à mon directeur Dr. Sébastien Sauvé et co-

directeur Dr. Benoit Barbeau. Pour Dr. Sébastien Sauvé, je vous remercie de m’avoir donné le

ticket pour ce voyage formidable à l’Université de Montréal. On dit « 知遇之恩 » en chinois, c’est-

à-dire que je serais toujours reconnaissant pour votre gentillesse de reconnaître ma valeur et me

recruter dans votre équipe. Pour Dr. Benoit Barbeau, je vous remercie de m’avoir guidé et

encouragé pendant ce voyage impeccable. On dit « 教诲之恩 » en chinois, c’est-à-dire que je

serais toujours reconnaissant pour vos conseils et critiques, ceux qui m’ont bénéficié au cours de

mon doctorat et me bénéficieront certainement à l’avenir.

Ensuite, j’aimerais exprimer mes remerciements à Dr. Pierre Bérubé et Dr. Madjid Mohseni de

l’Université de Colombie Britannique pour vos suggestions sur mes projets de recherche.

J’aimerais particulièrement remercier Dr. Pierre Bérubé de m’avoir inclus dans votre projet de

recherche et m’avoir fait confiance pour la rédaction de l’article.

En plus, j’aimerais remercier nos collaborateurs, l’usine de production d’eau potable de Pont-Viau

et de Sainte-Rose, pour leurs soutiens sur le terrain. Je remercie également le programme

FONCER-TEDGIEER (Technologies environnementales de décontamination et gestion intégrée des

eaux et effluents résiduaires) pour la bourse doctorale.

Par ailleurs, je remercie Dr. Kim Lompe, Dr. Isabelle Papineau et Dr. Morgan Solliec pour leurs

suggestions et critiques sur mes plans expérimentaux et mes articles. Vos professionnalismes et

rigueurs sur la recherche m’ont inspiré énormément. C’était un grand plaisir et honneur d’avoir

travaillé avec vous. D’ailleurs, je remercie Yves Fontaine, Mireille Blais, Tetiana Elyart, Julie

Philibert, Jacinthe Mailly, Gabriel St-Jean et Dr. Sung Vo Duy pour leurs soutiens techniques, que

ce soit sur le terrain ou au laboratoire, ainsi que leurs patiences pendant nos communications. Je

remercie également les stagiaires Élise Renault, Pauline Morasse et Maxime Combes pour leurs

aides pour la maintenance des pilotes et les travaux au laboratoire.

XXII

Pour mon épouse Xiameng Feng, je suis très chanceux de t’avoir avec moi le longue du chemin de

ma vie. Tu n’es pas seulement ma conjointe mais aussi ma confidente. Je te remercie pour tes

accompagnements, encouragements et confiances. Grâce à toi, cela me fait plus d’envie

d’imaginer l’avenir.

Pour mes parents, je vais vous écrire en chinois. 感谢父母的养育之恩，感谢父母在物质和精

神上对我一如既往的支持。每当我遇到困难或是人生的重要选择时，你们都能给予我鼓励，

打消我的疑虑，让我更加坚定地走下去。是你们不服输、不放弃、坚毅、强大的内心造就

了我现在独立、上进的人格。我自认为并不是一个聪明绝顶的人，但是我是一个爱“做梦”

并且愿为之默默努力的人。感谢父母在我“做梦”的时候，给予我力量和勇气，让我无所

畏惧，大胆地去追梦。

1

Chapitre 1 – Introduction

1.1 Matière organique naturelle

1.1.1 Définition et classement La matière organique naturelle (MON) est définie comme une matrice complexe des substances

organiques présentes dans les eaux, les sédiments et les sols. Les composés de MON varient

énormément en termes de taille, d’hydrophobicité et de structure. Par exemple, chaque composé

de MON peut comprendre une combinaison unique de groupes fonctionnels, y compris

benzéniques, phénoliques, carboxyliques, hydroxyles, amines, estériques, nitrogénés (Gjessing,

1976). La composition de la MON varie d’abord géographiquement parce que la concentration et

les caractéristiques de la MON dépendent de son origine laquelle est dictée par les conditions

géographiques (dont l’hydrographie, la pédologie et la végétation) ainsi que les conditions

climatiques (Fabris et al., 2008). En plus, la composition de la MON varie temporellement pour

une même location dû à la variation climatique saisonnière (Simith and Kamal, 2009). Par

exemple, les précipitations, les ruissellements de fonte des neiges, les inondations et les

sécheresses contribuent à la variation saisonnière de la composition de la MON (Delpla et al.,

2009).

Étant donné la complexité de la composition de la MON, les composés de MON sont rarement

étudiés individuellement. En revanche, la MON est communément étudiée de façon collective

afin d’identifier des caractéristiques propres à la MON. Premièrement, la MON est constituée en

de substances non-humiques et de substances humiques. Les composés de substances non-

humiques ont des formules moléculaires distinctes avec des identités uniques, tels que glucides,

graisses, cires, alcanes, peptides, acides aminés, protéines, lipides et acides organiques (Adusei-

Gyamfi et al., 2019), alors que les substances humiques n’ont pas de formule moléculaire unique.

L’analyse élémentaire démontre que les substances humiques sont composées principalement

de carbone (55%-57%), oxygène (34%-36%), hydrogène (4%-6%), azote (0,9%-3%) et sulfure

(0,4%-1,8%) (Rice and MacCarthy, 1991; Pettit, 2004; Bravo et al., 2017). Les groupes

2

fonctionnels, tels que carboxyle, phénol et alcool, sont les plus fréquemment présents dans les

substances humiques. Les substances humiques peuvent être d’avantage divisées en trois

fractions : 1) l’humine (masse molaire : 100 kDa -10 000 kDa) est insoluble dans l’eau quel que

soit le pH; 2) l’acide humique (masse molaire : 0,5 kDa-100 kDa) est soluble à pH > 2 et 3) l’acide

fulvique (masse molaire : 0,2 kDa-10 kDa) est soluble dans toutes les conditions de pH (Sillanpää,

2014). Des exemples d’acide humique et d’acide fulvique sont illustrés à la Figure 1.1. Les acides

humiques et les acides fulviques sont omniprésents dans les eaux de surface, et constituent 10%-

30% de la MON dissous dans l’eau de mer, 70%-90% de la MON dissous dans l’eau de terre

humide, 40%-90% de la MON dissous dans les cours d’eau et environ 50% de la MON dissous dans

l’eau de lacs (Xue and Sigg, 1999; Thurman, 2012; Lipczynska-Kochany, 2018).

Figure 1.1 - Structures moléculaires hypothétiques pour l’acide humique et l’acide fulvique

(Stevenson, 1982; Buffle, 1977).

En deuxième lieu, la MON peut être fractionnée par rapport à sa polarité (i.e., hydrophobe et

hydrophile) et sa charge (i.e., acide, neutre et base), ce qui permet de fractionner la MON aux six

catégories (Tableau 1.1). En général, la fraction hydrophobe de la MON consiste en des anneaux

aromatiques avec des doubles liaisons conjugués ou des structures phénoliques alors que la

fraction hydrophile est composée de carbones aliphatiques ou de structures azotées, tels que

acides aminés, sucres et protéines.

3

Tableau 1.1 - Fractions de la MON par rapport à la polarité et la charge (Health Canada, 2020).

Fraction Classe de composés

Hydrophobe

Acides Acides forts : acides humiques et fulviques, acides alkyl monocarboxyliques et

dicarboxyliques de haute masse moléculaire, acides aromatiques

Acides faibles : Phénols, tannins, acides alkyl monocarboxyliques et dicarboxyliques de

moyenne masse moléculaire

Bases Protéines, amines aromatiques, amines alkyl de haute masse molaire

Neutres Hydrocarbures (par exemple, terpénoïdes), aldéhydes, méthylcétones et alcools

alkyliques de masse moléculaire élevé, éthers, furanes, pyrrols

Hydrophile

Acides Acides hydroxyles, sucres, sulfoniques, acides alkyl monocarboxyliques et

dicarboxyliques de base masse moléculaire

Bases Acides aminés, purines, pyrimidines, alkylamines de base masse moléculaire

Neutres Protéines, glucides (par exemple, polysaccharides, alcools alkyliques de faible poids

moléculaire, aldéhydes et cétones), cellulose et dérivés de la cellulose

En troisième lieu, la MON peut être classée par rapport aux tailles moléculaires en utilisant la

technique de filtration membranaire en série ou la chromatographie d’exclusion stérique. Huber

et al. (2011) ont ainsi suggéré de répartir la MON en cinq fractions en utilisant la chromatographie

d’exclusion stérique couplée à un détecteur de carbone organique en ligne (en l’anglais : liquid

chromatography-organic carbon detector, LC-OCD). Les auteurs ont ensuite caractérisé les cinq

fractions, i.e., 1) les biopolymères (masse molaire >>20,000 Da), tels que polypeptides,

polysaccharides, protéines et sucres aminés; 2) les substances humiques (masse molaire : environ

1000 Da), y compris les acides humiques et fulviques; 3) les blocs de construction (masse molaire :

300-500 Da), qui sont les hydrolysats des substances humiques; 4) les acides de faible masse

moléculaire (masse molaire : <300 Da) et 5) les composés neutres de faible masse moléculaire

(masse molaire : <300 Da), tels que alcools, aldéhydes, cétones, sucres et acides aminés. Cette

méthode de fractionnement est souvent utilisée pour la caractérisation de la MON pour le

traitement de l’eau dû à l’utilité des informations obtenues pour optimiser les procédés de

traitement.

4

Finalement, les composés de MON se distinguent par rapport à leur biodégradabilité. Les

composés, tels que glucides, acides aminés et protéines sont facilement biodégradables tandis

que les substances humiques sont difficiles à biodégrader. Pettit (2004) a conclu que les

substances humiques peuvent même persister dans l’environnement pour des siècles sans

biodégradation significative.

1.1.2 Impact en eau potable Bien que l'exposition à la MON soit commune et ne soit pas associée à des effets directs sur la

santé humaine, la présence et les caractéristiques de la MON auront des impacts importants sur

la production de l'eau potable. Tout d'abord, la MON peut entraîner une augmentation des

plaintes de consommateurs car elle peut contribuer à la couleur, aux goûts et aux odeurs

indésirables de l'eau potable (Edzwald, 2010; Thurman, 2012). En deuxième lieu, la MON peut

contribuer indirectement aux impacts sur la santé humaine. Par exemple, la MON forme des sous-

produits de désinfection (SPD) réglementés et non réglementés lorsqu’elle réagit avec les

désinfectants (Tian et al., 2013). Certains sous-produits de désinfection (SPD) se sont avérés être

cancérogènes pour les humains et néfastes pour les écosystèmes (Krasner et al., 2006). De plus,

la MON favorise le développement de biofilms dans les systèmes de distribution en fournissant

de nutriments, ce qui permet de protéger de organismes pathogènes (e.g., légionelles) et ainsi

détériorer la qualité de l’eau (Hijnen et al., 2018). En troisième lieu, la MON peut également

perturber les procédés de traitement de l’eau. Par exemple, la MON peut forcer à augmenter la

dose de coagulant requise pour le traitement de l’eau et ainsi augmenter la production de boues.

La MON peut également réduire la durée de fonctionnement de filtres et donc augmenter la

fréquence de rétro-lavage et ainsi diminuer l’efficacité de filtre. Elle peut entraîner le colmatage

de membranes, augmenter la pression transmembranaire et la consommation d’énergie pour

l’opération de membranes. Enfin, une fraction de la MON entre en compétition avec les

micropolluants organiques lors de la mise en œuvre des procédés d’adsorption sur charbon actif.

5

1.1.3 Mesure et caractérisation Paramètres généraux. Bien que les nombreux composés organiques qui contribuent à la MON ne

puissent pas être mesurés directement, il existe un certain nombre de paramètres généraux qui

peuvent être utilisés pour fournir une indication de la concentration de la MON. Les paramètres

généraux les plus couramment utilisés comprennent le carbone organique total (COT), le carbone

organique dissous (COD), la demande chimique en oxygène (DCO), l'absorbance ultraviolette (UV)

et la couleur.

Le carbone organique total (COT) est la somme du carbone organique particulaire (COP) et dissous

(COD). Une définition opérationnelle largement acceptée du COD est le carbone organique dans

un échantillon d'eau filtrée à travers un filtre de 0.45 µm. Le COT et le COD sont les paramètres

les plus pratiques à utiliser dans le domaine du traitement de l’eau (Edzwald, 2010).

Généralement, toutes les méthodes de quantification du carbone organique dans l'eau

impliquent l'oxydation. Avant l’invention d’analyseurs de COT, un agent oxydant était ajouté et la

quantité d'agent utilisée exprimait la concentration de carbone présent (i.e., DCO). Dans les

techniques modernes, divers types d'oxydation sont utilisés (combustion, rayonnement, agents

oxydants et oxydation supercritique) et le CO2 engendré est mesuré par spectroscopie infrarouge

(IR) (Matilainen et al., 2011).

La spectroscopie d'absorption ultraviolette (UV) est la mesure de l'atténuation d'un faisceau de

lumière après son passage à travers un échantillon ou après réflexion sur une surface

d'échantillon. Bien que les absorptivités molaires varient en raison de la gamme de chromophores

présents dans la structure de la MON, toute longueur d'onde de 220 à 280 nm est considérée être

approprié pour les mesures de la MON (Matilainen et al., 2011). Les différentes longueurs d'onde

peuvent identifier de différents chromophores de la MON. Par exemple, l'absorbance à 220 nm

est associée à la fois aux chromophores carboxyliques et aromatiques, tandis que l'absorbance à

254 nm est typique pour les groupes aromatiques avec divers degrés d'activation (Korshin et al.,

2009). UV254 a été identifié comme une mesure de substitution potentielle pour la mesure du

DOC malgré la tendance à surreprésenter le caractère aromatique.

6

Bien que la couleur puisse être mesurée à l’aide de comparateurs visuels, elle est plus

couramment mesurée à l'aide de méthodes spectrophotométriques. Par exemple, les chercheurs

ont utilisé l'absorbance de la lumière visible à 420 nm comme mesure de la couleur organique

(Ekström et al., 2011 ; Weyhenmeyer et al., 2014). Cependant, une longueur d'onde comprise

entre 450 nm et 465 nm a été proposée comme méthode spectrophotométrique standard (APHA

et al., 2017). La présence de particules en suspension (par exemple, argile, oxydes de fer et de

manganèse) peut donner aux eaux une couleur apparente et un échantillon filtré à travers un

filtre de 0.45 µm est défini de manière opérationnelle comme la couleur vraie. La comparaison

des résultats de couleur vraie et apparente peut aider les opérateurs à déterminer si les plaintes

de couleur sont liées à la MON ou à des particules.

Techniques de caractérisation. Bien que les paramètres généraux soient souvent faciles à

mesurer, les informations sur le caractère de la MON présente dans l’eau demeurent à

investiguer. À ce jour, diverses techniques sont disponibles afin d’explorer le caractère de la MON.

Le SUVA (UV absorbance spécifique) est défini comme l'absorbance UV d'un échantillon donné à

254 nm divisée par la concentration en COD de l'échantillon. Ce rapport décrit l'hydrophobie et

l'hydrophilie de la MON dans l'eau; un SUVA >4 indique un caractère principalement hydrophobe

et surtout aromatique, alors qu'un SUVA <2 illustre un caractère principalement hydrophile

(Edzwald, 2010). Le calcul de SUVA est largement utilisé pour évaluer le caractère de la MON car

il est facile et peu coûteux à déterminer, et il est un bon indicateur pour les changements de la

qualité de l'eau de source (Westerhoff et al., 1999; Imai et al., 2001; Weishaar et al., 2003;

Reckhow et al., 2007).

La spectroscopie de fluorescence est aussi une technique en plein essor pour caractériser la MON.

Il s’agit d’exciter les molécules d'analyte par irradiation à une certaine longueur d'onde, et le

rayonnement émis est mesuré à une autre longueur d'onde. La fluorescence peut donner un

aperçu des caractéristiques chimiques de la MON, car elle est fonction de la structure moléculaire

et de groupes fonctionnels (Hudson et al., 2008). Spécifiquement, une conformation moléculaire

particulière appelée le fluorophore est caractérisée par des longueurs d'onde d'excitation et

d'émission spécifiques. Ces fluorophores sont utiles pour décrire la composition structurelle des

7

substances humiques (Baker et al., 2008 ; Zhang et al., 2008 ; Bieroza et al., 2009). Parmi les

spectroscopies de fluorescence, la spectrophotométrie EEM (i.e., excitation-emission matrice) à

fluorescence tridimensionnelle est une technique de plus en plus populaire en raison de la

simplicité de la mesure et la possibilité de l’utiliser comme outil de détection en temps réel. L'EEM

est un spectre 3D dans lequel l'intensité de fluorescence peut être présentée en fonction de la

longueur d'onde d'excitation et d'émission (Valencia et al., 2013). Le spectre EEM visualise une

gamme de fluorophores différents avec une longueur d'onde d'excitation et d'émission allant de

∼200 à ∼500 nm, ce qui est plus révélateur que la technique traditionnelle à balayage unique

(Spencer et al., 2007). La fluorescence EEM fournit des informations précieuses sur l'élimination

de différentes fractions de MON dans le traitement de l'eau (Jeong et al., 2013; Lee et al., 2013;

Sanchez et al., 2014). L'eau brute typique contient deux pics de fluorescence majeurs, décrits

comme des maxima de fluorescence de type humique et de type protéine (Baghoth et al., 2009)

ou trois pics appelés les fluorophores de type de tryptophane, fulvique et humique (Spencer et

al., 2007 ; Baker et al., 2008 ; Hudson et al., 2008).

Les composés de la MON peuvent être fractionnés à l'aide de résines échangeuses d’ions

disponibles dans le commerce (e.g., XAD-8, XAD-4) tel que mentionné précédemment.

Cependant, la mesure des six fractions de MON (tableau 1.1) demande beaucoup de temps et de

travail (Minor et al., 2014). Les composés de la MON peuvent également être physiquement

fractionnés, tel que mentionné précédemment, en fonction de la différence de taille moléculaire

en utilisant une membrane ou la chromatographie d'exclusion stérique (Koudjonou et al., 2005).

Les fractions peuvent être ensuite analysées par une détection du carbone organique et/ou de

l'azote organique (i.e., LC-OCD-OND) (Huber et al., 2011). Le fractionnement par membrane est

souvent complexe à opérer, sans mentionner que le colmatage des membranes reste un obstacle

pour la caractérisation de la MON.

Tests biologiques. Plusieurs indicateurs ont été développés afin de décrire la teneur en MON

biodégradable (MOB) dans l'eau. Le carbone organique assimilable (COA) évalue le potentiel de

l'eau pour soutenir la repousse microbienne dans un système de distribution d'eau potable (Van

Der Kooij et al., 1982) tandis que le COD éliminé par l'activité microbienne est défini comme le

carbone organique dissous biologique (CODB) (Joret et Lévi, 1986). Le COA est la partie du CODB

8

qui est le plus facilement converti en biomasse par les bactéries, tandis que le CODB est la partie

du carbone organique qui est utilisées par des micro-organismes hétérotrophes pour les activités

métaboliques, tels que la production d'énergie, la croissance de la biomasse (Terry et Summers,

2018). Dans les eaux de surface, le ratio COA/COT varierait de 0,2 % à 38,3 % et avait une médiane

de 2,8 % tandis que le ratio CODB/COT varierait de 1 % à 72 % avec une médiane à 20 % (Terry et

Summers, 2018).

1.1.4 Source et occurrence La MON présente dans les eaux de surface peut provenir des processus naturels ou des activités

anthropogéniques.

Tout d’abord, la source naturelle de la MON peut être allochtone ou autochtone. La MON

allochtone est dérivée de l’écosystème terrestre, i.e., la MON présente dans les sols, tels que

l’humus du sol, la litière végétale, la biomasse microbienne, les animaux en décomposition, les

exsudats racinaires, est transportée dans les cours d’eau par le ruissellement et/ou la

précipitation (Kalbitz et al., 2000). La MON allochtone est principalement composée de

substances humiques, et elle a tendance à être de nature hydrophobe et réfractaire à la

biodégradation (Edzwald, 2010). La MON autochtone quant à elle est dérivée de la production et

la décomposition de la biomasse microbienne et végétale dans les sources d’eau (Nguyen et al.,

2002; Zhou et al., 2014). Les algues planctoniques s’avèrent être une contribution importante

pour les grands lacs tandis que les macrophytes sont de contributeurs importants pour les petits

lacs (Bertilsson and Jones, 2003; Wetzel, 2003). La prolifération d’algues et de cyanobactéries

représente une source importante de la MON autochtone, et peut être nocive pour la santé

humaine étant donné que plusieurs cyanotoxines s’avèrent toxiques pour les humains et les

animaux (Watson et al., 2016). La contribution de la MON autochtone à la MON totale du cours

d’eau varie de 5% à 100% selon les conditions géographiques et climatiques (Bertilsson and Jones,

2003; Wetzel, 2003; Tomlinson et al., 2016). Les composés de la MON autochtone consiste en

mono- et polysaccharides, acides aminés, peptides, protéines, acides nucléiques, acides

organiques, lipides et acides gras (Pivokonsky et al., 2006 ; Henderson et al., 2008). La MON

9

autochtone a tendance à être de nature plus hydrophile et plus facilement biodégradable

(Edzwald, 2010).

La source anthropogénique de la MON est constituée principalement des rejets d’eaux usées, des

eaux pluviales, du ruissellement de champs agricoles et des rejets industrielles. La MON

anthropogénique s’avère être de nature plus hydrophile et plus riche en azote (Imai et al., 2001).

Les concentrations de la MON (présentes sous forme de carbone organique total, COT) dans l'eau

brute représentent l'effet net des processus hydrologiques et biogéochimiques dans le bassin

versant ou l'aquifère (Eckhardt and Moore, 1990). Le tableau 1.2 résume le COT des eaux brutes

à travers des provinces et territoires canadiens. En général, le COT varie spatialement tel que

mentionné précédemment, et le COT de l’eau de surface démontre une concentration plus élevée

que celle de l’eau souterraine. C’est parce que la matière organique est assujettie à l’adsorption

et la biodégradation lorsque l’eau souterraine traverse les sols (Thurman, 1985; Aiken and

Cotsaris, 1995; Aitkenhead-Peterson et al., 2003). La MON de l’eau souterraine s’avère être de

nature plus hydrophile et plus réfractaire à la biodégradation (Diem et al., 2013).

Tableau 1.2 - Carbone organique total (COT) dans les eaux brutes pour différentes provinces et

territoires canadiens (Canada Health, 2020).

Provinces canadiennes et territoires Carbone organique total (mg C/L)

Eaux souterraines Eaux de surfaces

Terre-Neuve-et-

Labradora

Nombre d’échantillons 322 833

Médiane 1,2 6,5

Moyenne 2,0 7,0

90ième centile 4,3 11,4

Nouvelle-Écosse


Médiane 1,2 4,6

Moyenne 2,3 5,8


Nouveau-Brunswick Nombre d’échantillons 893 324

10

Médiane 2,0 4,8

Moyenne 2,1 4,8


Québec


Médiane 2,8 6,0

Moyenne 3,1 6,2


Manitoba


Médiane 2,9 10,9

Moyenne 4,0 11,6

90ième centile 8,2 16,2 a : Les données pour Terre-Neuve-et-Labrador sont du carbone organique dissous.

1.2 Procédés de traitement pour l’enlèvement de la MON

Afin de sélectionner, concevoir et exploiter de manière appropriée des procédés de traitement

de l'eau, une compréhension des variations de la concentration et du caractère de la MON est

nécessaire. Pour ce faire, les exploitants de services d’eau devraient idéalement avoir des

connaissances au sujet de la MON présente dans leur source d’eau, y compris 1) la source et

l’occurrence de la MON; 2) les interactions avec d’autre composés dans l’eau (tel que l’ion

bromure) et avec les produits chimiques utilisés pour le traitement; 3) les interactions avec les

procédés de traitement; et 4) les impacts sur le système de distribution (Health Canada, 2020).

Une étude de traitabilité spécifique à la source devrait être menée pour évaluer et comparer les

options de traitement pour l'élimination de la MON dans une source d’eau spécifique (Kastl et al.,

2016). L'étude de traitabilité doit inclure des essais au laboratoire ou à l'échelle pilote en tenant

compte des objectifs concomitants de qualité de l'eau, tels que les risques microbiens, les SPD, la

stabilité biologique et le contrôle de corrosion. De nombreux procédés de traitement peuvent

aider à éliminer la MON de l’eau brute (Figure 1.2) (Sillanpää, 2014). Ces options de traitement

sont revues en détail dans les sections suivantes.

11

Figure 1.2 - Aperçu des processus de traitement pour l’enlèvement de la MON.

1.2.1 Coagulation Mécanismes et type de coagulant. Deux mécanismes sont impliqués pendant la coagulation : 1)

les composés de la MON sont neutralisés par les coagulants positivement chargés et forment ainsi

des précipités insolubles; 2) les composés de la MON peuvent être adsorbés sur les flocs de

coagulants (Edzwald et al., 2011). Le choix du coagulant dépend des caractéristiques de l’eau à

traiter. Les coagulants couramment utilisés comprennent les coagulants à base d’aluminium et

de fer, des floculants de polymères inorganiques, des polyélectrolytes organiques et des

coagulants composites (Sillanpää, 2014).

Avantages et inconvénients. La coagulation est la méthode la plus couramment utilisée pour

l’enlèvement de la MON étant donné son bon rapport d’efficacité-prix dans la plupart des

applications. Cependant, l’application de coagulant doit être soigneusement analysée en fonction

de la source, car la coagulation ne peut éliminer que certaines fractions de la MON. Par exemple,

la MON allochtone a tendance à être de nature hydrophobe et peut être généralement enlevée

par la coagulation, tandis que la MON hydrophile a tendance à être plus difficile à traiter par la

coagulation (Volk et al., 2002; Chow et al., 2004). En effet, pour les sources d’eau riches en MON

neutre et hydrophile, la coagulation sera inefficace (Chow et al., 2006). Par conséquent, il est

important d'effectuer des jar-tests et/ou de tests du potentiel de formation de SPD pour

déterminer la faisabilité de la coagulation pour l'élimination de la MON.

1.2.2 Adsorption Mécanismes et type d’adsorbant. Les composés organiques peuvent être adsorbés sur la surface

d’adsorbants par la force van der Waals et l’interaction thermodynamique (Worch, 2012).

L'adsorbant le plus couramment utilisé pour le traitement de l'eau est le charbon actif, qui peut

Processus de traitement pour l’enlèvement de la MON

Coagulation Adsorption Filtration Membranaire

Oxydation avancée

Échange d’ionsBiodégradation

12

être appliqué sous forme de charbon actif en poudre (CAP) ou en grain (CAG). Le CAP est souvent

appliqué en lit fluidisé alors que le CAG est appliqué en lit fixé. Le charbon actif a une structure

poreuse et peut fournir une grande surface spécifique, ce qui permet d’avoir une capacité de

sorption élevée pour les substances organiques (Simpson, 2008).

Avantages et inconvénients. L’adsorption est une technologie efficace pour l'élimination de la

MON et d'autres polluants organiques présents dans l’eau (i.e., micropolluants organiques). Les

adsorbants, tel que le charbon actif, est facile à opérer et demandent peu de produits chimiques.

Cependant, le charbon actif n'a pas été largement utilisé comme stratégie principale pour le

contrôle de la MON, car la capacité d'adsorption du charbon actif a tendance à s'épuiser

rapidement et la régénération (par exemple régénération par la chaleur) peut être coûteuse

(Prévost et al., 1998). Donc, l’adsorption est utilisée en tant que processus supplémentaire afin

d’améliorer l’élimination de la MON, particulièrement pendant la saison de goûts et d’odeurs en

été (par exemple, pendant la saison de la floraison d’algues).

1.2.3 Filtration membranaire Mécanismes et type de membrane. La filtration membranaire est un processus alternatif pour

l’enlèvement de la MON. Quatre types de membranes à pression sont actuellement utilisées dans

le traitement de l'eau potable : la microfiltration (MF), l'ultrafiltration (UF), la nanofiltration (NF)

et l'osmose inverse (OI). Les membranes sont généralement classées selon le type des substances

qu'elles éliminent, la pression de fonctionnement et la taille des pores ou la masse moléculaire

coupée (en anglais : molecular weight cut-off, MWCO). MF et UF sont appelés membranes à basse

pression et sont utilisées pour l'élimination de particules/agents pathogènes. Le mécanisme

d'élimination prédominant est l'exclusion de taille pour les membranes MF et UF. NF et RO quant

à eux sont appelés membranes à haute pression et sont utilisées pour l'élimination de la MON et

de substances inorganiques (par exemple sodium, chlorure, calcium et magnésium). Le

mécanisme d'élimination prédominant est la différence de diffusivité et la force électrostatique.

En général, plus de 50% des molécules de MON ont une masse moléculaire < 1 kDa et 80% ont

une masse moléculaire < 10 kDa (Sillanpää et al., 2015). Par conséquent, MF (MWCO : > 100 kDa)

13

et UF (MWCO : 1 kDa – 100 kDa) ne peuvent pas suffisamment enlever la MON, et une membrane

NF étanche est nécessaire pour éliminer la MON (MWCO : 200-300 Da) (Sillanpää et al., 2014).

Avantages et inconvénients. La filtration membranaire est facile à opérer et demande peu de

produits chimiques (similaire à l’adsorption). Cependant, il est important de considérer le

potentiel de colmatage car la MON colmate les membranes, ce qui entraîne une augmentation

de pression et une diminution de débit. Un rétrolavage avec/sans produits chimiques est

nécessaire afin de maintenir la performance de membrane. Les indicateurs du potentiel de

colmatage comprennent un faible SUVA, une fraction hydrophile élevée, une concentration

élevée en azote dissous ou une concentration élevée en biopolymères (Lee et al., 2006; Amy,

2008; Croft, 2012; Kimura et al., 2014; Siembida-Lösch et al., 2014).

1.2.4 Processus d’oxydation avancée Mécanismes et type de processus. Les processus d’oxydation avancée consistent à enlever la

MON par la réaction oxydante avec les radicaux. De nombreuses configurations sont disponibles,

y comprise ozone/UV, ozone/H2O2, UV/ H2O2 et la réaction Fenton. Il faut noter que les processus

d'oxydation avancée sont d’abord utilisés pour la désinfection et/ou la dégradation des

contaminants organiques ciblés, et ils n’ont pas été utilisé en tant que stratégie principale pour

l’enlèvement de la MON (Sillanpää et al., 2014).

Avantages et inconvénients. Les processus d’oxydation avancés suivi par un traitement

biologique peuvent diminuer la fraction de la MON biodégradable dans l’eau traitée, ce qui

permet d’augmenter la biostabilité de l’eau pendant la distribution. Cependant, bien que les

processus d'oxydation avancée puissent, en principe, éliminer une fraction de la MON, ils peuvent

également augmenter la formation de SPD ou produire nouveau SPD (Bond et al., 2011). Par

exemple, bien que l’ozonation puisse éliminer une fraction de la MON, elle peut aussi former du

bromate dans une eau contenant du bromure. Donc, l’utilisation des processus d’oxydation

avancée pour l’enlèvement de la MON devrait être évaluée soigneusement avant sa mise en

place.

14

1.2.5 Biodégradation Mécanismes et type de processus. Le traitement biologique a pour but d’enlever la fraction

biodégradable de la MON (i.e., MOB) dont profitent les micro-organismes hétérotrophes. Trois

configurations du traitement biologique sont disponibles, i.e., la filtration par les berges, la

filtration rapide sur média granulaire et la filtration lente sur sable. La filtration par les berges

consiste à localiser des puits d'approvisionnement en eau verticaux ou horizontaux près d'une

rivière afin d’utiliser la berge et l'aquifère adjacent comme filtre naturel et ainsi enlever la MON

par adsorption et biodégradation (Kuehn et Mueller, 2000; Ray et al., 2002). La filtration rapide

(vitesse de filtration : > 4 m/h) consiste à utiliser des filtres à média granulaire (par exemple sable,

anthracite et charbon actif granulaire) sans l’utilisation d’un désinfectant pendant l’opération, et

la MOB peut être biodégradée par la communauté microbienne qui se développe sur la surface

de média. Le charbon actif biologique (CAB) est le type de filtration rapide le plus couramment

utilisé pour l’enlèvement de la MOB. Le CAB est souvent placé après le traitement conventionnel

(i.e., coagulation-floculation-décantation-filtration) en tant que processus de polissage.

D’ailleurs, le CAB est souvent utilisé après l’ozonation (i.e., ozone – charbon actif biologique) étant

donné que l’ozonation peut transformer la MON à la MOB et ainsi rendre la MON plus facilement

biodégradable. La filtration lente sur sable consiste à traiter de l'eau brute qui s'écoule par

gravité à faible vitesse (vitesse de filtration : < 0.5 m/h) à travers un lit de sable. Pendant

l’opération, la croissance microbienne se développe dans le lit de sable et le support de gravier.

En plus, les bactéries et les matières organiques présentes dans la source d'eau s'accumulent au-

dessus du filtre à sable pour former un « schmutzdecke » (une couche de limon riche en

microorganismes et matières organiques). Par conséquent, la croissance microbienne à l'intérieur

du filtre et du schmutzdecke contribuent à l’enlèvement de la MON par l’adsorption et la

biodégradation. (Logsdon et al., 2002).

Facteurs impactant la performance de biodégradation. La performance de biodégradation en

termes d’enlèvement de la MOB dépend de plusieurs facteurs. D’abord, l’efficacité de

biodégradation dépend de la concentration de MOB dans l’eau à traiter. La MON réfractaire à la

biodégradation ne répond pas au traitement biologique sauf si un processus d’oxydation est

utilisé afin de la rendre plus biodégradable. En deuxième lieu, le type de média filtrant aurait un

15

impact sur la performance de biodégradation. Le charbon actif granulaire (CAG) s’est avéré être

plus performant par rapport à l’anthracite et au sable dû à sa structure poreuse, un avantage qui

permet de supporter plus de biomasse (Emelko et al., 2006). En troisième lieu, la température

aurait un impact direct sur la performance de biodégradation étant donné que la croissance

microbienne est fonction de la température. Une baisse de température peut entraîner la baisse

de la performance de biodégradation, un défi important pour l’opération de filtres biologiques

aux pays nordiques. En quatrième lieu, les conditions opérationnelles, tels que le temps de

contact (présenté en empty bed contact time, EBCT) et les processus de rétro-lavage, sont

importants pour la performance de la filtration biologique. Une exposition plus longue de l’eau à

traiter avec la biomasse peut rendre un enlèvement plus significatif (Liu et al., 2017). Le rétro-

lavage implique généralement d’appliquer l’air et l’eau avec ou sans oxydant pour enlever les

particules. Les résultats ont démontré que le filtre biologique est plus performant lorsque

l’oxydant ne présente pas dans l’eau de lavage (Basu et al., 2016). Finalement, les nutriments sont

un facteur vital pour la performance de biodégradation. La croissance microbienne peut être

limitée dû au manque de nutriments. Une formule moléculaire empirique a été proposée pour

les cellules bactériennes (i.e., C60H87O23N12P) (Basu et al., 2016). Les filtres biologiques peuvent

ainsi être bonifiés par l’ajout de nutriments afin de booster la croissance microbienne (Liu et al.,

2017).

Avantages et inconvénients. La filtration biologique est un processus économique, et elle est

facile à opérer. Elle peut non seulement éliminer de la MOB, mais aussi éliminer de micropolluants

organiques et de contaminants inorganiques (par exemple l’azote ammoniacal, le fer et le

manganèse). Cependant, la filtration biologique ne peut pas être utilisée en tant que stratégie

principale pour l’enlèvement de la MON du fait qu’elle peut seulement enlever la partie

biodégradable de la MON (i.e., 5%-20%) (Terry and Summers, 2018). D’ailleurs, sa performance

est fortement impactée par la température, un défi majeur pour les pays nordiques où la

température varie énormément durant l’année.

16

1.2.6 Échange d’ions Contexte en résine échangeuse d’ions. Les résines échangeuses d'ions (en anglais : ion exchange

resins, IX) sont des billes de plastique synthétiques. La structure de IX est constituée de polymères

sur lesquels de groupes fonctionnels sont fixés de façon permanente (Bolto et al., 2002). Comme

les groupes fonctionnels sont chargés, les contre-ions sont attachés aux groupes fonctionnels afin

de maintenir la neutralité électrique. Les contre-ions sont mobiles et peuvent entrer et sortir de

résines. Les résines échangeuses d’ions consistent généralement en deux groupes selon la charge

du contre-ion, i.e., les résines échangeuses de cations (CIX) et les résines échangeuses d'anions

(AIX). Les résines échangeuses de cations sont souvent utilisées pour l'adoucissement de l'eau

tandis que les résines échangeuse d’anions sont utilisées pour enlever la MON et les anions

inorganiques.

Mécanismes de l’enlèvement de la MON. Puisque les molécules de MON contiennent

généralement un groupe d’acide carboxylique ou un groupe phénol dans leurs structures, elles

peuvent perdre de protons lors de l'hydrolyse. Par conséquent, la majorité de la MON est chargée

négativement dans l'eau et est donc échangeable contre les résines échangeuses d’anions

(Cornelissen et al., 2008). Deux mécanismes sont impliqués dans l'enlèvement de la MON par les

résines, i.e., 1) les parties négativement chargées de la MON peuvent échanger avec les contre-

ions sur la résine (i.e., échange d’ions); 2) les parties non-ioniques de la MON peuvent être

adsorbées sur la surface de résine par les forces van der Waals, les liaisons hydrogène et les

interactions thermodynamiques (i.e., adsorption) (Fu et Symons, 1990; Bolto et al., 2002; Tan et

Kilduff, 2007). Cependant, Fu et Simons (1990) ont constaté que l'adsorption n’avait pas lieu de

manière significative pour la fraction de MON de grande masse moléculaire (> 1k Da), et que

l'échange d'ions était le mécanisme dominant pendant le processus d’échange d’ions.

Configuration de réacteur pour la résine échangeuse d’ions. Les contacteurs à lit mobile ont été

proposés pour la résine afin d’enlever la MON de l’eau turbide. Le processus MIEX (en anglais :

Magnetic Ion EXchange) consiste à utiliser des réacteurs à flux complètement mélangé pour

enlever la MON en mode continu (Figure 1.3A). Le processus consiste généralement en : 1) 10-30

min de temps de contact avec une vitesse de rotation lente (vitesse de la pointe < 5 m/sec) et; 2)

le processus de séparation des résines pour le recyclage et la régénération des résines (Slunjski

17

et al., 2000). Cette configuration a reçu des intérêts intensifs depuis 2000, mais il faut noter que

son application est limitée aux résines MIEX, une résine particulière qui est incrustée de particules

d’oxydes de fer (Boyer, 2015). Par ailleurs, une autre configuration de réacteur à lit mobile a été

proposée en 2010 par PWN Technologies aux Pays-Bas (figure 1.3B) (Koreman et Galjaard, 2016).

Dans ce processus, les résines sont injectées à l'eau brute et s’écoulent dans un réacteur à flux de

bouchon (temps de contact : 10-30 min). Ensuite, les résines sont séparées, régénérées et

stockées dans le réservoir de résine avant la prochaine injection. Tous les types de résine

échangeuse d'anions peuvent s’adapter au procédé SIX, une différence majeure par rapport au

procédé MIEX (Koreman et Galjaard, 2016). Finalement, l’échange d'ions en lit fluidisé (en

anglais : fluidized ion exchange, FIX) est une autre configuration de réacteur à lit mobile

(Cornelissen et al., 2009) (figure 1.3C). Dans ce processus, l'eau brute est pompée dans un

réacteur à courant ascendant à un débit constant de sorte que les résines sont fluidisées dans le

contacteur. Ce procédé a été testé avant une membrane d'ultrafiltration, et les résultats ont

indiqué que le procédé FIX pouvait atténuer le colmatage de la membrane causé par la MON

(Cornelissen et al., 2009). Néanmoins, il existe encore peu études sur l’application du FIX à pleine

échelle, et la performance de celle-ci demeure inconnue.

Le réacteur à lit fixe consiste à traiter l’eau brute à flux descendant par la gravité. Les rétrolavages

à contre-courant sont nécessaires afin de réduire l'impact du colmatage. Les résultats de l’analyse

du cycle de vie ont montré que les réacteurs à lit fixe consomment moins d'énergie et de résines

mais nécessitent plus de sels pour la régénération par rapport au procédé MIEX (Amini et al.,

2015). Cependant, les auteurs ont également indiqué que l'entretien des résines pour les

réacteurs à lit fixe est un facteur clé afin de réduire leurs impacts environnementaux. Une bonne

gestion de résines (par exemple une fréquence de régénération appropriée et une bonne

stratégie de rétrolavage) peut rendre le réacteur à lit fixe plus économique et respectueux de

l’environnement par rapport au réacteur à lit mobile.

18

(A) Procédé MIEX

(B) Procédé SIX

Réservoir de résine fraîche

Contacteurs

Eau brute

Décanteur

Eau traitée

RecyclageRégénération

Saumure

Eau brute

Contacteurs Décanteur

Eau Taitée

Contacteur de régénération

Saumure

Réservoir de résine faîche

Injection de résine

19

(C) Procédé FIX (D) Contacteur à lit fixé

Figure 1.3 - Configuration de réacteur pour la résine échangeuse d’ions. (A) : Procédé MIEX; (B) :

Procédé SIX; (C) Procédé FIX; (D) Contacteur à lit fixé.

Facteurs impactant la performance de résines échangeuses d’ions. Les facteurs influençant la

performance de résines échangeuses d’ions en termes d’enlèvement de la MON comprennent les

caractéristiques de l’eau (le caractère de MON, la présence d’anions inorganique), les propriétés

de résines (le caractère de polymère, le type de contre-ion) et les stratégies d’opération. Ces

facteurs sont revus en détail dans les sections suivantes.

Le caractère de la MON peut influencer l'efficacité d'enlèvement de la MON. En premier lieu,

étant donné que la résine enlève la MON principalement par force électrostatique, les molécules

de MON de densité de charge plus élevée peuvent avoir une plus grande affinité sur la résine

(Boyer et al., 2008). En plus, la taille des molécules ou la distribution de masse moléculaire peut

également influencer la performance. Une masse moléculaire élevée peut empêcher les

moléculaires d’entrer dans les pores de résine, un phénomène appelé l'exclusion de taille qui peut

réduire l’efficacité de la résine (Bazri et Mohseni, 2016; Fu et Symons, 1990; Tan et Kilduff, 2007).

Par ailleurs, l’hydrophobicité des composés de MON peut aussi impacter la performance de

résine, particulièrement pour les composés neutres où l’adsorption est le mécanisme

prédominant (Finkbeiner et al., 2019).

Les anions inorganiques présents dans l’eau, pour leur part, peuvent rivaliser avec les molécules

de MON pour les positions d’échange d’ions sur la résine et ainsi détériorer la performance de la

Eau brute

Eau traitée

Contacteur

Régénérant

SaumureEau traitéeou saumure

Eau bruteou régénérant

Contacteur

Eau de lavage

Eau de lavage

20

résine. Par exemple, nombreuses études ont démontré que l’enlèvement de la MON par IX

diminuait lorsque la concentration de sulfate augmentait (Ates et Incetan, 2013; Boyer et Singer,

2006; Tan et Kilduff, 2007). D’ailleurs, Tan et Kilduff (2007) ont constaté que l’augmentation de

la concentration de sulfate dans l'eau brute pouvait modifier la fraction de la MON enlevée par la

résine, i.e., les résines changent leurs préférences à la MON de haute masse molaire lorsque la

concentration de sulfate augmente. Une explication probable serait que les molécules de MON à

haute masse molaire auraient une densité de charge plus élevée, ce qui les rend plus

favorablement échangées contre les résines (Tan et Kilduff, 2007). En plus, Croué et al. (1999) ont

observé que la présence du bicarbonate et du chlorure dans l'eau peut augmenter l’enlèvement

de la MON en raison de la déshydratation de la MON. Ce phénomène, appelé « l’effet de salting

out », peut changer l’hydrophobicité de molécules de MON et ainsi les rendre plus facile à

adsorber sur la résine. Néanmoins, il faut aussi noter qu’une haute concentration de chlorure ou

de bicarbonate peut diminuer l’enlèvement de la MON dû à la compétition. Finalement, l'effet de

cations sur la performance d’échange d’ions a été peu exploré dans la littérature, sauf que la

présence de fer dans l'eau peut provoquer un colmatage de résine et ainsi diminuer la capacité

de résine (Hongve et al., 1999).

Le caractère du polymère de résine influence également la performance en termes d’enlèvement

de la MON. En premier lieu, de nombreuses études ont suggéré que l’enlèvement de la MON

augmentait avec la teneur en eau dans la structure de résine (Bolto et al., 2002; Cornelissen et

al., 2008). Une explication raisonnable serait qu'une teneur élevée en eau dans la structure de

résine peut favoriser la diffusion des molécules de MON dans les pores de résine. Par ailleurs, le

matériel de polymère peut également impacter l’enlèvement de la MON. La résine polyacrylique

s’avère plus efficace que celle en polystyrène parce qu’elle est plus hydrophile et favorise ainsi la

diffusion des molécules (Bolto et al., 2002; Fu et Symons, 1990). D’ailleurs, la résine d’une

structure macroporeuse est généralement plus performante que celle en gel dû à la facilité de

diffusion dans les pores de résine (Bolto et al., 2002). Enfin, bien que la capacité d'échange d'ions

soit un facteur important, une valeur élevée à elle seule est peu susceptible de favoriser

l’enlèvement de la MON. Cornelissen et al. (2008) ont démontré que la résine avec la capacité

d’échange la plus élevée avait la pire performance pendant leurs tests.

21

Le type de contre-ion impacte également la performance des résines en termes d’enlèvement de

la MON; Un contre-ion plus sélectif contre la résine peut diminuer l’efficacité du traitement. Le

chlorure est le contre-ion le plus conventionnel pour la résine échangeuse d’anions dû fait que :

1) le chlorure est chimiquement inerte dans l'eau; 2) le chlorure a une affinité faible sur la résine

de sorte que les anions ciblés peuvent être favorablement échangés avec le chlorure; 3) la

régénération avec le chlorure de sodium est économique et facile à opérer (Rokicki et Boyer

2011). Cependant, puisque le rejet de la saumure de chlorure pose un risque pour les

écosystèmes et les systèmes d’égout (dû à la salinité élevée) (Matošić et al., 2000 ; Rokicki et

Boyer, 2011), de nombreuses études ont été consacrées à trouver des contre-ions alternatifs. Le

bicarbonate a été proposé en tant que contre-ion alternatif au chlorure pour la résine échangeuse

d’anion, parce que 1) le bicarbonate a une affinité similaire au chlorure sur la résine; 2) le rejet

de la saumure chargée de bicarbonate est plus facile étant donné qu’il est chimiquement bénin

(Hu et Boyer, 2017 ; Maul et al., 2014 ; Ness et Boyer, 2017 ; Rokicki et Boyer, 2011 ; Walker et

Boyer, 2011). Cependant, le bicarbonate de sodium a une efficacité de régénération inférieure à

celle du chlorure de sodium principalement en raison de sa faible solubilité dans l'eau (96 g

NaHCO3/L à 20°C) (Hu et Boyer, 2017 ; Matošić et al., 2000 ; Ness et Boyer, 2017). Par conséquent,

l'enlèvement de la MON par la résine en forme bicarbonate peut diminuer plus rapidement que

celle en forme chlorure au cours du service. D’ailleurs, Jelinek et al. (2004) ont constaté des

précipitations formées dans l'eau traitée lorsque la concentration de calcium était supérieure à

100 mg/L. Par ailleurs, le sulfate a été testé en tant que contre-ion pour la résine échangeuse

d’anion. Puisque le sulfate a une affinité plus grande que le chlorure sur la résine, l’enlèvement

de la MON par la résine en forme sulfate est inférieur à celui par la résine en forme chlorure (33%

vs 42%) (Verdickt et al., 2012). Cependant, l’utilisation de la résine en forme de sulfate peut éviter

l'enlèvement indésirable des anions inorganiques (par exemple : le nitrate, le sulfate et le

bicarbonate).

La stratégie d’opération est aussi un facteur important en ce qui concerne la performance de

résine pour l’enlèvement de la MON. Par exemple, une dose de résine plus grande dans un

réacteur à lit mobile ou un EBCT plus long dans un réacteur à lit fixe peut augmenter l’efficacité

de traitement. Un rétrolavage plus fréquent pour le réacteur à lit fixe peut diminuer le colmatage

22

du filtre et ainsi améliorer l’efficacité de traitement. La stratégie de régénération, telle que la

concentration de régénérant, la fréquence de régénération, aurait également un impact sur

l'enlèvement de la MON. Une régénération plus fréquente peut directement augmenter

l’efficacité de traitement, mais cela va aussi engendrer plus de saumure qui nécessite une gestion

soigneuse.

Avantages et inconvénients. Les processus d’échange d’ions possèdent plusieurs avantages par

rapport à d’autre processus dédiés à l’enlèvement de la MON. D’abord, les processus d’échange

d’ions sont faciles à opérer, particulièrement pour le réacteur à lit fixe. En plus, les processus

d’échange d’ions sont efficaces et peuvent constituer la stratégie principale pour l’enlèvement de

la MON ou ils peuvent être intégrés dans les chaînes de traitement de l'eau existantes. Une revue

de littératures détaillée a été réalisée pour évaluer les études précédentes sur l’enlèvement de la

MON par IX (Levchuk et al., 2018). Par ailleurs, les résines échangeuses d’ions sont régénérables

et leurs régénérations peuvent être effectuées sur site. Cependant, cette régénérabilité entraîne

également l'inconvénient majeur du processus IX, i.e., le processus de régénération peut produire

une saumure qui contient une concentration élevée en chlorure et le rejet de cette saumure est

restreint selon les règlementations environnementales applicables.

1.2.7 Résine échangeuse d’ions en mode biologique En 2017, un nouveau processus appelé « résine échangeuse d’ions en mode biologique » (en

anglais : biological ion exchange, BIEX) a été proposé afin d’enlever la MON de l’eau de surface

(Schulz et al., 2017). Il s’agit d’opérer la résine échangeuse d’ions dans un réacteur à lit fixe avec

une régénération peu fréquente de sorte qu’une communauté microbienne peut se développer

sur la surface de résine et ainsi contribuer à l’enlèvement de la MON par biodégradation. Dans

un test à l’échelle laboratoire, Schulz et al. (2017) ont constaté qu’un filtre de résine biotique

enlevait 60 % de la MON alors qu’un filtre de résine abiotique enlevait 40 % de la MON. Ensuite,

le charbon actif biologique (CAB) et la résine en mode biologique (BIEX) ont été comparés en

termes d’enlèvement de la MON à l’échelle du laboratoire (Winter et al., 2018) et du pilote (Amini

et al., 2018). Les résultats ont démontré que le filtre BIEX avait réalisé un enlèvement plus grand

par rapport au CAB (labo : 56% vs 15 %; pilot : 62% vs 7%). Ces résultats démontrent que le filtre

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BIEX peut être une stratégie pour l’enlèvement de la MON. D’ailleurs, puisque la régénération est

menée moins fréquemment, la quantité de sel et la production de saumure ont été réduits

énormément, un avantage qui peut rendre l’utilisation de résine plus respectueux de

l’environnement par rapport au mode d’application conventionnel. Cependant, les mécanismes

de l’enlèvement de la MON dans le filtre BIEX et sa faisabilité à l’usine de traitement de l’eau

demeure inconnus.

1.3 Problématiques, objectifs, hypothèses et structure de la thèse

1.3.1 Problématiques Bien que certaines études aient été faites sur le BIEX avant commencer cette recherche, il n’y

avait que peu d’études élucidant les mécanismes de l’enlèvement de la MON pour le BIEX. Schulz

et al. (2017) ont démontré que le filtre de résine biologique (i.e., BIEX) enlevait plus de COD (20%)

par rapport à un filtre abiotique. Cependant, pendant leur test, l’azoture (négativement chargé)

a été ajouté dans l’affluent du filtre abiotique afin de réprimer la croissance microbienne, ce qui

pouvait rivaliser avec les molécules de MON et ainsi baisser l’enlèvement de la MON par l’échange

d’ions. Autrement dit, la performance réalisée par le filtre abiotique a été sous-estimée, et la

contribution de la biomasse dans le filtre BIEX demeure incertaine. Dans la recherche

subséquente, les chercheurs ont pu quantifier la biomasse présente dans le BIEX par la méthode

d’ATPmétrie (Adénosine Triphosphate, ATP) (Winter et al., 2018; Amini et al., 2018). Les résultats

ont démontré que la quantité de biomasse était similaire ou même plus grande par rapport au

filtre CAB sur la surface du filtre. Cependant, une preuve directe qui démontre la relation entre

l’activité microbienne et l’élimination de la MON est manquante. Finalement, Amini et al. (2018)

ont supposé que l’échange d’ions secondaires contribue aussi à l’enlèvement de la MON dans le

filtre BIEX, i.e., la MON échange avec les anions inorganiques pré-retenues à la suite de

l’épuisement du chlorure. Cependant, les concentrations d’anions n’ont pas été suivies dans les

études précédentes (Winter et al., 2018; Amini et al., 2018), et les mécanismes d’échanges d’ions

secondaires demeurent inconnus.

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L’intégration du processus d’échange d’ions à la filière du traitement de l’eau demeure un défi

pour l’utilisation de résines. Le procédé MIEX est souvent placé au début de la filière de

traitement étant donné que la résine peut enlever la MON dans l’eau turbide, une stratégie qui

bénéficie aux processus subséquents (par exemple la baisse de la consommation de produits

chimiques). Par ailleurs, Grefte et al. (2013) ont démontré que le MIEX peut être aussi placé au

milieu (avant ozonation) ou à la fin (après la filtration lente par sable) d’une chaîne de traitement

(coagulation, décantation, filtration, ozonation, adoucissement, CAB et filtration lente par sable)

avec des performances similaires. Bien que l’application du MIEX à la filière du traitement de l’eau

ait été bien étudiée, l’application du réacteur à lit fixe a été peu explorée.

Finalement, pendant la conception du processus d’échange d’ions, les ingénieurs et les

chercheurs ne considèrent actuellement que les paramètres opérationnels, et l’impact négatif

causé par la saumure n’est souvent pas anticipé avant mettre le processus en place. De

nombreuses études ont été consacrées à optimiser l’opération de résine et la gestion de saumure

autre que la filtration BIEX afin d’alléger la gestion et l’impact de la saumure engendrée par la

régénération de résines. Néanmoins, une revue de ces mesures et un schéma de conception qui

inclut l’opération de résines et la gestion de saumure sont manquants.

1.3.2 Objectifs et hypothèses L’objectif général de cette thèse est 1) de comprendre et favoriser l’application de la résine

échangeuse d’ions en mode biologique (i.e., BIEX) pour l’enlèvement de la MON des eaux de

surface et 2) décrire et analyser les stratégies qui peuvent alléger la gestion de la saumure

engendrée par la régénération de résines. Les objectifs spécifiques et les hypothèses

correspondantes sont présentés ci-dessous :

Objectif 1 : Évaluer la performance de l’enlèvement de la MON pour les filtres BIEX en forme

bicarbonate et en forme chlorure.

Hypothèses : 1) les résines opérées en mode biologique enlèvent de la MON par l’échange d’ions

et la biodégradation; 2) La MON peut échanger avec les anions pré-retenus par la suite de

l’épuisement des anions pré-chargées (i.e., bicarbonate et chlorure); 3) La performance du BIEX

en forme bicarbonate est similaire à celle du BIEX en forme chlorure.

25

Objectif 2 : Évaluer l’enlèvement des composés de modèle (i.e., micropolluants organiques) sur

la résine en mode biologique et ainsi valider l’occurrence de biodégradation.

Hypothèses : 1) les composés biodégradables peuvent être biodégradés sur la résine en mode

biologique; 2) les composés négativement chargés peuvent échanger contre la résine en mode

biologique; 3) Les résines en mode biologique n’enlèvent pas de composés neutres et non-

biodégradables.

Objectif 3 : Évaluer la performance du filtre BIEX en filtration secondaire dans l’usine de

production d’eau avec une référence du filtre CAB.

Hypothèses : 1) le filtre BIEX est plus performant que le filtre CAB en termes d’enlèvement de la

MON; 2) le filtre BIEX est plus performant que le filtre CAB en termes d’enlèvement de

précurseurs de sous-produit de désinfection; 3) le filtre BIEX est similaire au filtre CAB en termes

d’enlèvement de l’azote ammoniacal; 4) le filtre BIEX a une plus grande perte de charge par

rapport au filtre CAB.

Objectif 4 : Résumer les stratégies sur l’opération de résine et la gestion de saumure et ainsi

proposer un nouveau schéma de conception pour le processus d’échange d’ions.

Hypothèses : 1) De nombreuses stratégies sur l’opération de résines (autres que le BIEX) existent

dans la littérature qui peuvent réduire l’impact négatif causé par la saumure; 2) De nombreuses

stratégies sur la gestion de saumure existent dans la littérature qui peuvent réduire l’impact

négatif causé par la saumure; 3) Une conception comprenant l’opération de résine et la gestion

de saumure peut réduire l’impact environnemental du processus d’échange d’ions.

1.3.3 Structure de la thèse Le chapitre 1 présente d’abord une revue de littératures en ce qui concerne la matière organique

naturelle et les processus de traitement pour l’enlèvement de la matière organique naturelle.

Ensuite, les problématiques, les objectifs et les hypothèses de cette étude sont présentés à la fin

du chapitre 1.

Le chapitre 2 aborde l’objectif 1 via une étude du pilote à longue terme (9 mois) à l’usine de

production d’eau potable de Pont-Viau (Laval). Les concentrations de COD, CODB et des anions

26

ont été suivies dans les affluents et les effluents du pilote, ce qui permet d’élucider l’effet

d’échange d’ions pour le filtre BIEX. La référence de l’article est présentée ci-dessous :

Liu, Z., Papineau, I., Mohseni, M., Peldszus, S., Bérubé, P. R., Sauvé, S., & Barbeau, B. (2021). Operating Bicarbonate-Form versus Chloride-Form Ion Exchange Resins without Regeneration for Natural Organic Matter Removal. ACS ES&T Water, 2021 1 (6), 1456-1463

Le chapitre 3 aborde l’objectif 2 via une étude en batch qui évalue l’enlèvement de composés de

modèle avec différentes charges et biodégrabilité par la résine biotique et abiotique. L’article est

en révision au journal de Science of the Total Environment.

Le chapitre 4 aborde l’objectif 3 via une étude du pilote à l’usine de production d’eau potable à

Sainte-Rose. La performance du filtre BIEX a été comparée avec les filtres de charbon actif

pendant une longue durée d’opération (9 mois), ce qui permet d’évaluer l’application du BIEX

dans l’usine de production d’eau. La référence de l’article est présentée ci-dessous :

Liu, Z., Lompe, K. M., Mohseni, M., Bérubé, P. R., Sauvé, S., & Barbeau, B. (2020). Biological ion exchange as an alternative to biological activated carbon for drinking water treatment. Water Research, 168, 115148.

Le chapitre 5 aborde l’objectif 4 via une revue de littérature au sujet des stratégies sur l’opération

de résine et la gestion de saumure, ce qui permet d’alléger le fardeau de la saumure produite par

la régénération de résines. La référence de l’article est présentée ci-dessous :

Liu, Z., Haddad, M., Sauvé, S., & Barbeau, B. (2021). Alleviating the burden of ion exchange brine in water treatment: from operational strategies to brine management. Water Research, 205, 117728.

Le chapitre 6 présente les conclusions principales de cette étude et propose des perspectives

pour l’avenir.

L’annexe A aborde une étude sur l’impact de température et EBCT pour les filtres BIEX et CAB.

Les essais ont été menés par les collaborateurs de l’Université de Colombie Britannique, et

l’article a été rédigé par moi. La référence de l’article est présentée ci-dessous :

Liu, Z., Mills, E. C., Mohseni, M., Barbeau, B., & Bérubé, P. R. (2022). Biological ion exchange as an alternative to biological activated carbon for natural organic matter removal: Impact of temperature and empty bed contact time (EBCT). Chemosphere, 288, 132466.

27

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38

Chapitre 2 – Operating bicarbonate-form versus chloride-form

ion exchange resins without regeneration for natural organic

matter removal

Abstract

Ion exchange (IX) is a promising drinking water treatment process for natural organic matter

(NOM) removal. However, standard IX processes require frequent regenerations with

concentrated NaCl solution, producing a brine that requires costly and complicated disposal

methods. To alleviate the burden of IX brine, we previously proposed operating IX with infrequent

regeneration to favor biomass development on the resins and thus benefit NOM removal through

biomass contribution, a process referred to as biological ion exchange (BIEX). The objective of the

present study is to evaluate the performance of BIEX filtration for NOM removal from primary IX

to complete exhaustion using bicarbonate-form and chloride-form IX resins. Parallel pilot-scale

bicarbonate-form and chloride-form BIEX were fed with surface water (dissolved organic carbon

(DOC) = 7 mg C/L) for 9 months without regeneration. The results demonstrated that bicarbonate-

form BIEX achieved a marginally lower DOC removal (median: 49 % vs. 53 %), a higher

biodegradable DOC (BDOC) removal (average: 50 % vs. 33 %) and a similar disinfection by-product

precursor removal compared to chloride-form BIEX. Overall, BIEX filtration using bicarbonate-

form and chloride-form IX resins offers a similar NOM removal efficiency and eases spent brine

management.

Keywords: Biological ion exchange; Bicarbonate-form ion exchange; Biofiltration; natural organic

matter; NOM fractionation; sustainability.

39

2.1 Introduction Natural organic matter (NOM) removal is of paramount importance for drinking water production

as it causes aesthetic problems such as color, odor and taste, leads to the formation of harmful

disinfection by-products (DBP), disrupts water treatment processes (e.g., membrane fouling) and

contributes to biofilm regrowth in distribution systems (Edzwald et al., 2010; Richardson et al.,

2007; Tian et al., 2013; Peleato et al., 2017; Hijnen et al., 2018;). Using strong base ion exchange

(IX) is an efficient option to remove NOM (Kim and Symons, 1991; Tan et al., 2005; Zhang et al.,

2019; MacKeown et al., 2020). IX removes NOM through two main mechanisms which are driven

by the functional groups of NOM moieties. Charged groups can sorb to IX functional groups

through electrostatic interactions whereas uncharged groups may adsorb onto the IX surface

through the Van der Waal force or hydrophobic interactions (Bolto et al., 2002; Fu and Symons,

1990; Li and SenGupta, 2004). IX offers several advantages as a process used for NOM removal.

First, it can target hydrophilic NOM which is not readily removed by conventional processes (e.g.,

coagulation and activated carbon) (Matilainen et al., 2010; Shutova et al., 2020). Second, IX has

proven to be flexible in terms of operational mode, reactor configurations and integration within

existing treatment trains (Greft et al., 2013). It can also be adapted into a broad spectrum of

applications varying from household point-of-use to full scale drinking water treatment plants

(WTP) (Amini et al., 2015). Third, IX can be regenerated on-site using concentrated NaCl solution

(e.g., 12 %w/v) as opposed to activated carbon which is usually regenerated off-site using energy-

intensive thermal processes (Ledesma et al., 2014). Yet, the drawback of IX regenerability is the

production of a highly concentrated chloride brine whose direct disposal to the environment is

restricted due to its high salinity (Levchuk et al., 2018). As a result, despite its ease of use and

great NOM removal efficiency, the implementation of IX processes is limited by the burden of

brine management and disposal (Sillanpää, 2014).

In an effort to prolong the service time of IX and reduce the production of brine, it has been

proposed to operate fixed-bed IX filters over an extended period of time without regeneration to

favor biomass development on the surface of IX beads and thereby benefit NOM removal through

biodegradation, i.e., biological ion exchange (BIEX) (Amini et al., 2018). Lab-scale experiments

indicated that in the absence of regeneration, biodegradation contributed to NOM removal as

40

BIEX achieved a higher dissolved organic carbon (DOC) removal than its abiotic counterpart

operated in parallel (BIEX: 60 % vs. Abiotic-IX: 40 %) (Schulz et al., 2017). BIEX also exceeded NOM

removal efficiency of biological activated carbon (BAC) filters at lab-scale (BIEX: 56 % vs. BAC: 15

%) as well as at pilot-scale (BIEX: 62 % vs. BAC: 5 %) (Amini et al., 2018; Winter et al., 2018).

Although BIEX proved to be less efficient than a conventional IX filter which was regenerated on

a weekly basis (IX: 80 % vs. BIEX: 62 %), this novel mode demonstrated the possibility of

significantly reducing the IX brine production (i.e., 331 days without regeneration) while

maintaining a high NOM removal (Amini et al., 2018).

Another alternative to ease brine management is to resort to the use of bicarbonate salt as an IX

regenerant (i.e., bicarbonate-form IX) as opposed to chloride salt. Bicarbonate brine can be

directly disposed into sewers or, in some cases, the environment as bicarbonate has little impact

on flora and fauna (Rokicki et al., 2011; Maul et al., 2014). Bicarbonate-form IX has a similar

affinity for DOC compared to chloride-form IX and the treated water is less corrosive due to the

release of bicarbonate as opposed to the release of chloride (Nguyen et al., 2010; Walker and

Boyer, 2011; Ishii and Boyer, 2011; Ness and Boyer, 2017). However, the real application of

bicarbonate-form IX remains limited as NaHCO3 is more expensive than NaCl and has a lower

solubility in water/regeneration efficiency which may progressively lower NOM removal

efficiency as the number of regeneration increases (Walker and Boyer, 2011; Hu et al., 2016; Ness

and Boyer, 2017; Hu and Boyer, 2017).

Consequently, it is of great interest to study the operation of bicarbonate-form IX in BIEX mode

as it could not only greatly reduce regeneration frequency but also ease brine disposal. This is of

paramount importance for drinking water utilities, and more important for small communities

located in remote areas which would otherwise depends on chemical/brine transportation.

However, as none of the previous studies assessed bicarbonate-form IX operated in biological

mode (BIEX), its NOM removal efficiency under such operational conditions remains unknown.

Hence, the objective of the present study is to evaluate the NOM removal performance of

bicarbonate-form IX operated in BIEX mode with a benchmark of chloride-form IX operated in

BIEX mode. To better understand the underlying mechanisms that come into play, anion loading

on IX, DOC removal, treated water biostability (biodegradable DOC), metabolic activity of

41

heterotrophic biomass on the media, DBP precursor removal, specific ultraviolet absorbance at

254nm (SUVA) and NOM fractionation in influent and effluent were assessed. The present study

is the first to evaluate the advantages and limitations of implementing bicarbonate-form IX resins

operated in BIEX mode at pilot-scale.

2.2 Methods and materials

2.2.1 Resin preconditioning The resin type used in this study is Purolite® A860, a type-I strong base anion exchange resin with

polyacrylic and quaternary ammonium as backbone and functional group, respectively. The

average diameter of Purolite® A860 resin is of 0.75 mm according to the supplier. Based on

previous measurements, the ion exchange capacity is quantified as 0.68 eq/L of resin and the

virgin resin is initially loaded with chloride (Amini et al., 2018). Batch preconditioning of both

chloride-form and bicarbonate-form IX was carried out in a lab-scale 10-L beaker prior to its

transfer to pilot-scale packed-bed filters. The preconditioning of chloride-form IX consisted of

three steps. The virgin IX was first mixed with 120 g/L NaCl solution (volume ratio = 1 L of resin:3

L of brine) at 250 rpm for 30 min. Then, it was mixed with demineralized water (volume ratio =

1:3) at 250 rpm for 5 min to wash out the excess chloride on resin surface. Finally, resin beads

were soaked in demineralized water (to evite ion exchange before test) before use (within 48h).

Contrary to chloride-form IX, bicarbonate-form IX are not commercially available. Consequently,

bicarbonate-form IX was prepared using chloride-form IX which was first mixed with 80 g/L

NaHCO3 solution (volume ratio = 1:3) at 250 rpm for 30 min. Then, the IX beads were mixed in

demineralized water (volume ratio = 1:3) at 250 rpm for 5 min to wash out the excess bicarbonate

on resin surface. A conversion rate was used to evaluate the conversion efficiency from chloride-

form IX to bicarbonate-form IX, which is defined as the chloride quantity in the brine (i.e., the

NaHCO3 solution after regeneration) divided by the theoretical chloride capacity on the IX beads.

These two steps were repeated three times in order to achieve a final conversion rate of 97 %

(data not shown). After conversion, the bicarbonate-form IX beads were soaked in demineralized

water before use (within 48h).

42

2.2.2 Pilot location and source water characteristics The pilot-scale columns were set up at the Pont-Viau drinking water treatment plant (Laval,

Canada) and were directly fed with raw water from Des Prairies River which is characterised by a

moderate DOC concentration (≈7.0 mg C/L), a neutral pH (≈7.2) and a low sulfate concentration

(5.6 mg/L) (Table S2.1). The pilot-scale columns were operated over a 9-month period (from April

2019 to January 2020).

2.2.3 Pilot design and operation Two polyvinyl chloride (PVC) columns (inner diameter of 101.6 mm) containing 0.6 m (≈ 4.8 L) of

either bicarbonate-form IX or chloride-form IX were co-currently operated for 250 days without

regeneration (hereafter BIEX-B and BIEX-C filter). The filtration rate was maintained at 2.4 m/h (4

bed volumes/h) throughout the study, which corresponded to an empty bed contact time (EBCT)

of 15 min. Columns were backwashed on a weekly basis with air (2 min at 55 m/h) and then

followed by unchlorinated treated (20 m/h) water from the WTP. Filter backwash was terminated

either based on wash water turbidity (≤10 NTU) or when a total backwash volume of 50 L (≈ 10

bed volumes) was reached.

2.2.4 Analytical methods Water temperature, pH and turbidity were monitored on-site on a weekly basis. Temperature

and pH were measured with a multimeter (HACH HQ40D), and turbidity was quantified with a

turbidity meter (Hach 2100) using Standard Method 2130 (APHA, 2017). Influent and effluent

samples were also collected on a weekly basis to assess DOC (TOC meter, Sievers 5310C, GE water,

USA) and UV254 absorbance (Ultrospec 3100pro, GE Healthcare, USA). Biodegradable DOC (BDOC),

which consists of measuring the DOC reduction in water samples incubated for 30 days with an

inoculum of suspended bacteria from the raw water, was assessed in parallel according to the

method developed by Servais et al. (1989) and optimised by Markarian et al. (2010). To assess

influent and effluent NOM reactivity, trihalomethane (THM) and haloacetic acids (HAA)

precursors were assessed on a weekly basis from June to August according to the Uniform

Formation Conditions test (UFC) (i.e., by maintaining a free chlorine residual of 1.0 ± 0.5 mg Cl2/L

after a contact time of 24h at pH 8.0 and 20°C) (Summers et al., 1996). HAA5 extraction was done

43

according to Method 552.3 (USEPA, 2003). Samples were then analyzed by gas chromatography

(7890B GC system, Agilent Technologies, USA) according to the Methods 524.2 (THM) and 552.3

(HAA) (USEPA, 1992; USEPA, 2003). Samples were also collected in May, June, July, August and

December to allow NOM characterisation by liquid chromatography – organic carbon detection

(LC-OCD) analyses in both influent and effluent through a broad range of operation phases. Such

characterisation allows the quantification of five NOM sub-fractions: biopolymers (BP), humics

(HS), building blocks (BB), low molecular weight acids (LMW acids) and low molecular weight

neutrals (LMW neutrals) (Huber et al., 2011). All analyses (DOC, UV254, BDOC, THM & HAA, LC-

OCD) were performed in duplicate, and mean values are reported in this study.

Anions (Cl-, SO42-, NO3-, NO2-, Br-) were quantified on 0.45 μm filtered samples (Merck Millex®-HV)

using an ion chromatograph (ICS 5000 AS-DP DIONEX, USA) equipped with an AS18 column.

Bicarbonate concentration was evaluated from alkalinity measurements which were performed

using the acid titration method 2320B (APHA, 2017). Heterotrophic biomass was evaluated using

the 14C-labelled glucose respiration test developed by Servais et al. (1991). Briefly, colonized

resins were extracted from different depths of the column (2, 10, 30, 50 cm) and incubated in

sealed penicillin flasks at 20°C for 3h with 14C-labelled glucose solution (1 mM with a radioactivity

0.1-1 µCi/mL). Then, the samples were acidified by adding H2SO4 (10 %) prior to being bubbled

using N2 for 10 min to extract CO2. CO2 gas was trapped in 2 mL Carbo-sorbE (PerkinElmer) and 8

mL PermafluorE+ (PerkinElmer) and then was measured by a liquid scintillation analyser

(PerkinElmer Tri-Carb 4910 TR) for radioactivity. The result was obtained in nmol glucose/(mL×h),

and then it was conversed to µg C biomass/mL using the factor 1.1 µg C/nmol glucose/h (Servais

et al., 1991).

2.2.5 Calculation of BIEX loadings To elucidate the progressive exhaustion of BIEX filters, the loadings of anions and DOC were

calculated using the weekly cumulative charge balance (expressed in eq/L of resin) as described

in equation (2.1):

#(%, ') = ∑ ∑ !"!",!,$#"%&',!,$$×'×()*()*!"

+,-+,.

/,01/,. (2.1)

44

where q (i,j) is the total accumulative loading (eq/L) of solute j in BIEX filter after week i. Five

solutes were considered in this equation: DOC, chloride, bicarbonate, sulfate and nitrate. Nitrite

and bromide were neglected from this analysis given that their concentrations in raw water were

below the limit of detection throughout the study period (NO2- < 0.2 mg/L; Br- < 0.1 mg/L). Cin, i, j

and Cout, i, j respectively represent the weekly concentration of solute j at week i in the influent

and effluent. Q is the daily flow (460 L/day) and Vresin is the volume of resin in the columns (4.8 L).

DOC charge density was assumed to be 10 meq/g C, a common value for NOM at neutral pH based

on Boyer et al. (2008).

2.2.6 Statistical analysis A Shapiro-Wilk test was used to confirm the normality of the data at a significance level of p >

0.05 before conducting the hypothesis tests. Based on the outcome, the hypothesis test was

conducted using either a paired t-test or a paired Wilcoxon test for normally and non-normally

distributed data, respectively. The null hypothesis was bicarbonate-form BIEX and chloride-form

BIEX were identical for the tested metrics at a significance level of p > 0.05. All statistical analyses

were conducted in the software R (version 3.6.3).

2.3 Results and discussion

2.3.1 BIEX loading during pilot study During the operation of BIEX columns, both NOM and background anions in feed water were

exchanged onto IX resins, and the BIEX loading of these solutes evolves with the bed volumes

(BV) of operation (Figure 2.1A). Prior to 5 000 BV, bicarbonate was continuously released until it

was exhausted. Meanwhile, DOC and sulfate continuously increased on the IX resins whereas

chloride first increased on the IX resins and then rapidly decreased until exhaustion (≈ 5 000 BV).

Little nitrate was exchanged onto the IX resins (max loading: 0.03 eq/L resin) mainly due to its low

concentration in feed waters (mean concentration 0.9 mg/L). Consequently, the dominant

mechanism on BIEX-B, prior to 5 000 BV, involved DOC and sulfate exchanging for bicarbonate.

As the exchange capacity was due to pre-charged bicarbonate, this phase is termed primary IX.

From 5 000 to 15 000 BV, sulfate that was previously retained during the primary IX phase began

45

to be exchanged with DOC. Other anions (chloride, bicarbonate and nitrate) remained

insignificantly loaded on the IX resin. Given that DOC was removed by exchanging with sulfate,

we have defined this phase as secondary IX. Subsequent to 15 000 BV, the IX resin was deemed

to be fully loaded with NOM since the IX loading with respect to bicarbonate and sulfate was

considered to be exhausted. However, a continuous increase of DOC loading was observed, which

is hypothesized to be attributable to the contribution of biomass (biosorption, biodegradation,

bioregeneration). The BIEX-C filter exhibited a very similar loading pattern to BIEX-B (Figure 2.1B),

with the exception of chloride which was initially charged onto the IX resins and whose

concentration gradually decreased until exhaustion during the primary IX. By the end of the study,

DOC loading on the IX resin in BIEX-C filter was 0.73 eq/L resin as opposed to 0.68 eq/L resin in

the BIEX-B filter. This higher loading was attributed to the slightly higher NOM removal in BIEX-C

filter, which will be presented in the next section.

Figure 2.1 - Cumulative loading (eq/L of resin) in the (A) bicarbonate-form BIEX (BIEX-B) and (B)

chloride-form BIEX (BIEX-C) for four anions and dissolved organic carbon (DOC) throughout the

study period (≈ 24 000 BV or 250 days). No regeneration occurred during this period.

2.3.2 NOM removal Effluent DOC was normalized to influent DOC to account for the variation of DOC concentration

in the BIEX effluents (DOC/DOC0). The normalized DOC concentrations are presented as a function

of bed volumes in Figure 2.2A. Prior to 5 000 BV (i.e., primary IX), DOC was efficiently removed

B

0

0.2

0.4

0.6

0.8

0 5000 10000 15000 20000 25000

Anio

n lo

adin

g (e

q/L

resi

n)

Bed volumes

DOCChlorideBicarbonateSulfateNitrate

A Primary IX

Secondary IX

Exhaustion

0

0.2

0.4

0.6

0.8

0 5000 10000 15000 20000 25000

Anio

n lo

adin

g (e

q/L

resi

n)

Bed volumes

Primary IX

Secondary IX Exhaustion

IX capacity IX capacity

HCO3-Cl-

NO3-SO42-

46

by both BIEX filters, although the mean performance of BIEX-B (70 ± 15 %) proved to be slightly

lower than BIEX-C (76 ± 8 %) (paired t-test, p = 0.04). At approximately 5 000 BV, a DOC

breakthrough was observed in both BIEX filters which resulted in normalized DOC concentrations

of 0.64 (DOC = 5.13 mg C/L) and 0.70 (DOC = 5.36 mg C/L) in BIEX-B and BIEX-C effluents,

respectively. From 5 000 to 15 000 BV (i.e., secondary IX), both effluent DOC concentrations first

decreased and then stabilized after 7 000 BV. During this phase, BIEX-B achieved equivalent DOC

removal compared to the BIEX-C (55-56 %). Subsequent to 15 000 BV (i.e., exhaustion), DOC

concentrations started to rise in both BIEX effluents until they reached a steady state (≈19 000

BV). Throughout the study period, the BIEX-B achieved a marginally lower DOC removal than did

the BIEX-C filter with a median of 49 % as opposed to 53 % (paired Wilcoxon test, p < 0.001).

The NOM removal process in BIEX filters is believed to be primarily due to the chromatographic

elution of different solutes (Liu et al., 2020). To portray this, NOM in raw water could be broken

up into three conceptual fractions (Figure 2.2B): 1) NOM1 has no affinity for the resin and breaks

through right from the time when the operation started; 2) NOM2 and NOM3 have respectively

a lower and higher affinity to the resin as compared to sulfate, but both have a higher affinity

than bicarbonate (i.e., NOM3>SO42->NOM2>HCO3->NOM1 as an affinity sequence). Hence, prior

to the DOC breakthrough (i.e., primary IX), NOM3 and NOM2 were both removed through the

exchange with bicarbonate (or chloride in the case of BIEX-C filter). When bicarbonate was

exhausted, NOM2 could no longer be exchanged and was progressively released from the IX as

other solutes with higher affinities (e.g., sulfate and NOM3) were exchanged against NOM2 on

the surface of the resins. The release of NOM2 from the IX phase increased the DOC concentration

and thus leading to a DOC breakthrough in BIEX effluents. Subsequent to the release of NOM2

(i.e., secondary IX), effluent DOC concentrations reached stability as the exchange between

NOM3 and previously charged sulfate on the IX took place. During the exhaustion phase

(subsequent to 15 000 BV), DOC in both BIEX effluents rose up again due to facts that IX capacity

was completely exhausted, and NOM removed through biomass contribution (biosorption,

biodegradation and bioregeneration) decreased as temperature decreased from 18°C to 1°C.

Finally, BIEX filters achieved a stable DOC removal (5-7 %) in cold water conditions which is

47

believed to be fully due to the biomass contribution (biosorption, biodegradation and

bioregeneration).

Figure 2.2 - (A) Normalized effluent dissolved organic carbon concentrations (DOC/DOC0) of


period (≈ 24 000 BV or 250 days). (B) conceptional displacement of different solutes in the BIEX-

B filter in primary ion exchange, secondary ion exchange and exhaustion phases. NOM was broken

up into three conceptional fractions with an affinity sequence of NOM3>SO42->NOM2>HCO3-

>NOM1.

2.3.3 BDOC removal Evaluation of BDOC removal allows an appreciation of the treated water biostability of both BIEX

filters. Over the course of the study, BDOC concentrations remained stable in raw water with 0.42

0

5

10

15

20

25

30

0.0

0.2

0.4

0.6

0.8

1.0

1.2

0 5000 10000 15000 20000 25000Te

mpe

ratu

re (°

C)

DOC/

DOC 0

Bed volume

BIEX-BBIEX-C

Temp.

Primary IX Secondary IX ExhaustionA

B

HCO3-

NOM3

SO42-

NOM2

NOM3

NOM3

SO42-

Primary IX Secondary IX Exhaustion

NOM1 NOM2NOM1

SO42-NOM2NOM1

NOM3SO42-NOM2NOM1

Mean DOC0 = 7 ± 0.7 mg C/L

48

± 0.15 mg C/L (i.e., 6 ± 2 % of DOC). BDOC removals in both BIEX filters followed a similar trend

throughout the study (Figure 2.3). Briefly, during primary IX, BDOC removal in the BIEX-B filter

decreased from 80 % to 23 % following the similar trend to DOC removal in the filter, suggesting

that BDOC removal was mainly due to ion exchange during primary IX. During secondary IX (i.e.,

subsequent to 52 days of operation), BDOC removal in the BIEX-B filter was significantly improved

to 70 % as temperature increased from 13°C (5 000 BV) to 26°C (11 000 BV) whereas DOC removal

remained stable from 7 000 to 11 000 BV. The improvement of BDOC removal was attributed to

biomass contribution (biosorption, biodegradation and bioregeneration) in warm water

conditions. However, as BDOC is only a marginal fraction of influent DOC in the current study (6

% of influent DOC), such improvement had limited impact on the overall DOC removal. Finally,

during exhaustion phase, BDOC removal in the BIEX-B filter dropped to 19 % as temperature

decreased to 2°C. Overall, average BDOC removal in BIEX-B filter (50 %) constantly proved to be

higher than BIEX-C (33 %) (paired t-test, p = 0.02). The higher BDOC removal in the BIEX-B filter

might be attributed to the favorable pH conditions as the release of bicarbonate led to a more

alkaline effluent than did the chloride one (mean pH:7.3 vs. 6.8 during primary IX).

Figure 2.3 - Biodegradable dissolved organic carbon (BDOC) removal in bicarbonate-form BIEX

(BIEX-B) and chloride-form BIEX (BIEX-C) filters.

To verify if the higher BDOC removal in the BIEX-B filter was due to the higher microbial activities,

the extent of bacterial colonization was assessed. Resins were extracted at 2, 10, 30, 50 cm filter

0

5

10

15

20

25

30

0%

20%

40%

60%

80%

100%

0 5000 10000 15000 20000 25000

Tem

pera

ture

(°C)

BDO

C re

mov

al

Bed volumes

BIEX-B

BIEX-C

Temp.

Primary IX


BDOC0 = 0.42 ± 0.15 mg C/L

49

bed depth after 10 000 BV. A 14C glucose respiration test was performed on each sample to

quantify the spatial distribution of heterotrophic biomass (expressed in µg C/mL resin) in both

BIEX filters (Figure S2.1). The BIEX-B filter had a lower heterotrophic biomass than the BIEX-C filter

in the upper layers of the filter bed (2 and 10 cm deep or a corresponding EBCT of 0.5 min and

2.5 min) while no significant difference was observed for the bottom layers (30 and 50 cm deep

corresponding to 7.5 min and 12.5 min EBCT) of the two BIEX filters (p = 0.43-0.45). This result

indicates that the measurement of heterotrophic biomass by 14C glucose respiration test did not

correspond to the BDOC removal in biofilters, a conclusion which is consisted with previous

studies using other biomass assessment methods on BAC filters, such as phospholipid analysis

and adenosine triphosphate analysis (Wang et al., 1995; Urfer et al., 1997; Emelko et al., 2006;

Pharand et al., 2014). Therefore, more studies are needed to elucidate the relationship between

biomass activity and the BDOC removal.

2.3.4 THM and HAA precursors removal DBP precursor removals were assessed on a weekly basis, but solely after DOC breakthrough to

study the removal during secondary IX (5 300 - 11 000 BV) (i.e., subsequent to media

colonization). THM-UFC (mean value: 434 µg/L with 97 % chloroform and 3 %

bromodichloromethane) and HAA5-UFC (mean value: 391 µg/L with 41 % dichloroacetic acid and

59 % trichloroacetic acid) concentrations in raw water remained stable throughout the study

period. Both BIEX filters achieved similar THM-UFC or HAA5-UFC concentrations in the effluent

throughout the study period with a median of 125 THM-UFC µg/L or 110 HAA5-UFC µg/L (Figure

S2.2), suggesting that both BIEX filters had similar DBP precursors removal for the study period.

To allow a further comparison of THM or HAA precursors removal in BIEX-B and BIEX-C filters, the

NOM reactivity (DBP concentration per unit of chlorine demand) was calculated (Figure 2.4). The

THM or HAA precursor removal was equivalent for both BIEX filters as no significant differences

in THM (p = 0.19) or HAA5 reactivities (p = 0.70) in both BIEX effluents were observed. Moreover,

little differences were observed amongst THM and HAA speciation in both BIEX effluents for

chloroform (p = 0.33), bromodichloromethane (p = 0.18), dichloroacetic acid (p = 0.52) and

trichloroacetic acid (p = 0.57). These results confirmed that the mechanisms involved in the

removal of DBP precursors by both BIEX filters do not differ.

50

Figure 2.4 - Natural organic matter (NOM) reactivities (expressed as µg disinfection by-products

per mg Cl2 consumed) in raw water, bicarbonate-form BIEX (BIEX-B) and chloride-form BIEX (BIEX-

C) effluents from 5 300 to 11 000 bed volumes (i.e., secondary IX).

2.3.5 NOM hydrophobicity SUVA was calculated using available DOC and UV254 data, and effluent SUVA was normalized to

influent SUVA (SUVA/SUVA0) to assess the evolution of NOM hydrophobicity in BIEX effluents

throughout the study (Figure 2.5). Prior to 5 000 BV, the effluent of both BIEX filters increasingly

became more hydrophobic as SUVA values continuously increased through time. We hypothesize

that the removal of hydrophobic NOM (such as humics) declined during this period due to the

depletion of IX capacity on the resin surface. This is of importance given that hydrophobic NOM

has a greater DBP formation potential (Chen et al., 2008). Consequently, although effluent DOC

values remained low during primary IX phase, the DBP formation potential increased over time.

Interestingly, when DOC breakthrough occurred, the SUVA dropped from 3.5 L×m-1×mg-1 to 3.0

L×m-1×mg-1 for both BIEX filters, indicating a release of hydrophilic compounds. After

breakthrough, SUVA values continued to decline until 13 000 BV, and then increased until it

stabilized after 15 000 BV at a normalized SUVA of 0.9 for both BIEX effluents. Overall, no

significant difference in SUVA was observed between BIEX-B and BIEX-C effluents throughout this

study (paired Wilcoxon test, p = 0.44).

30

35

40

45

50

55

60

Raw

wat

er

BIEX

-B

BIEX

-C

Raw

wat

er

BIEX

-B

BIEX

-C

THM-UFC HAA5-UFC

NOM

Rea

ctiv

ity (µ

g/m

g Cl

2 de

man

d)

median

25-75 percentiles

max-min

51

Figure 2.5 - Normalized effluent specific ultraviolet absorbance at 254 nm (SUVA/SUVA0) of


period (≈ 24 000 BV or 250 days).

2.3.6 NOM fractionation LC-OCD analyses were conducted at 2 592, 5 280, 8 064, 10 656 and 22 752 BV on both BIEX

influent and effluent to compare the removal of NOM fractions during three phases: primary IX,

secondary IX and exhaustion. NOM fractions in raw water remained relatively stable throughout

this study (Table S2.2). To better understand the affinity of various NOM fractions during the

operation of BIEX filters, removals (1-C/C0) of the various NOM fractions are calculated as a

function of bed volumes (96 BV = 1 day). Globally, BIEX-B (Figure 2.6) and BIEX-C (Figure S2.3) had

similar evolution with respect to different NOM fractions throughout the study period. Briefly, in

the BIEX-B filter, HS is the preferred NOM fraction eliminated although the initial removal

efficiency (≈ 90 %) decreased to 66 % at breakthrough (5 280 BV) and then increased back to 90

% during secondary IX. In comparison, BB were also preferably removed during primary IX (78 %)

but were released from the BIEX filter at breakthrough (-62 %). Additionally, BB were no longer

removed from the BIEX filters during secondary IX or exhaustion phases. Initial LMW neutrals

removal by the BIEX-B filter (41 %) decreased to 15 % at breakthrough and recovered to 25 %

after breakthrough. Finally, BP (22 %) and LMW acids (20 %) were removed in the BIEX-B filter to

the similar extent and remained stable throughout the study.

0

0.2

0.4

0.6

0.8

1

1.2

0 5000 10000 15000 20000 25000

SUVA

/SUV

A 0

Bed volume

BIEX-B

BIEX-C

Mean SUVA0 = 3.5 ± 0.2 L/(m・mg)

Primary IX Secondary IX Exhaustion

52

Figure 2.6 - Removals of natural organic matter (NOM) fractions in bicarbonate-form BIEX (BIEX-

B).

According to the LC-OCD analysis, BB, HS and LMW neutrals all surged during the DOC

breakthrough, an evidence proving that NOM released from IX at the breakthrough (i.e., NOM2)

is mainly found within these three NOM fractions. Moreover, subsequent to DOC breakthrough,

BB were no longer removed by both BIEX filters. This may be attributable to its low charge density

as it is also hardly removed by coagulation (Huber et al., 2011). The number-averaged molecular

mass (Mn) of effluent HS (calculated from the LC-OCD) reveals a drop from 1800 Da prior to

breakthrough to 1000 Da at breakthrough. Such data demonstrate that HS in NOM2 has a lower

molecular weight (MW) than in NOM1. This observation corroborates previously reported data

where the major NOM fractions that surged during DOC breakthrough proved to possess a low

apparent molecular weight (AMW) of 0.5-1 kDa (Kim and Symons, 1991). This change in HS MW

in BIEX effluent at DOC breakthrough was accompanied with a sudden decrease of SUVA values.

This indicates that NOM2 is relatively more hydrophilic than the NOM in the BIEX effluent right

before the breakthrough. These observations lead us to conclude that NOM2 is composed of

molecules which exhibit lower charge density, lower MW and lower aromaticity than NOM3.

Nevertheless, LC-OCD and SUVA analysis does not allow a perfect discrimination among the

conceptional NOM1, NOM2 and NOM3. Given the confounding effect of biomass colonization

-100%

-75%-50%-25%

0%25%

50%75%

100%

0 5000 10000 15000 20000 25000

Rem

oval

Bed volumesBP HSBB LMW acidsLMW neutrals

Primary IX

Secondary IX

Exhaustion

53

and loading state of IX resins, more studies are needed to elucidate the NOM elution in BIEX

filters.

2.3.7 Implications for IX operation for NOM removal NOM removal in IX filters is usually correlated to the number of bed volumes, i.e., we can run an

IX filter just within the primary IX phase to maximize NOM removal. However, the IX filter will

need to be regenerated more often (≈ 4 000 BV for this source water) and thus will produce more

spent brine, which is undesirable from an operational, financial and environmental standpoint. In

order to better balance the treatment efficiency against brine production, we suggest 1) putting

BIEX filters off-line during the DOC breakthrough (4 000 BV-6 000 BV) to avoid the surge of DOC;

2) pursue filter operation until ≈ 15 000 BV before launching a regeneration to maximize IX

treatment capacity. To minimize the offline period, the throughput on the IX contactor (BV/h)

could be increased to accelerate the transition phase from primary to secondary IX. As a result,

the IX service cycle could be prolonged by 9 000 BV (considering putting the IX back online from

6 000 BV to 15 000 BV), which corresponds to a 3.25-fold increase of cycle length. Additionally,

operating IX in BIEX mode is expected to have similar DOC removal compared to conventional

operation mode (i.e., with regeneration) for the same BV of treated water. This is due to the fact

that regeneration of IX resins does not allow a complete recover of IX capacity and hence the

treatment efficiency in conventional mode would decrease as the service cycle increases

(Bassandeh et al., 2013). Further, biomass would establish itself on the IX resins due to the

prolonged service time and thus benefit the removal of BDOC and ammonia (Figure S2.4). Finally,

it should be noted that the metrics of BIEX performance are determined on a case-by-case basis,

which depends on feed water characteristics (e.g., sulfate concentration and NOM properties).

Lab-scale pretests are recommended to investigate the NOM elution profile prior to real

application.

Although BIEX can significantly improve the service time of IX resins, one still needs to regenerate

the resins when the efficiency no more satisfies the treatment objective. As bicarbonate salt has

a lower solubility in water/regeneration efficiency than chloride salt, BIEX regenerated with

bicarbonate salt may experience DOC removal efficiency decline as the operation cycle increases.

54

Our lab-scale tests demonstrated that the regeneration efficiency using bicarbonate salt (1 M/L)

is approximately 60% of that using chloride salt (1 M/L) for resins operated for approximately

13000 BV (data not shown). Therefore, the regeneration strategies using bicarbonate salt are yet

to be optimized prior to implementation for BIEX resins. To this end, BIEX filters can be thoroughly

regenerated with chloride regenerant (e.g., 12 %w/v) in an intermittent manner, and in order to

minimize brine production, such chloride regenerant could be directly reused several times

without any efficiency decline before disposal (Medina et al., 2018). Caustic can also be used to

help recovering IX capacity from exhausted resins. Additional research will be needed to identify

the optimal regeneration conditions for resins operated in BIEX mode.

2.4 Conclusions The present study is the first to assess NOM removal efficiencies of BIEX filters using bicarbonate

and chloride-form IX resins from primary ion exchange to complete exhaustion. Results

demonstrate that, as opposed to its chloride-form counterpart, bicarbonate-form BIEX achieved

a slightly lower DOC removal (median: 49 % vs. 53 %), a higher biodegradable DOC (BDOC)

removal (average: 50 % vs. 33 %) and a similar disinfection by-product precursor removal. BIEX

filters using bicarbonate-form and chloride-form IX resins offer similar NOM removal efficiencies

and ease spent brine management. BIEX filters can be regenerated with bicarbonate or chloride

solution when the efficiency no more satisfies the treatment objective. However, the

regeneration efficiency using bicarbonate salt is yet to be assessed for BIEX resins considering the

lower solubility of bicarbonate salt than chloride salt in water. Future studies need to identify the

optimal regeneration strategy for BIEX resins according to the type of counter-ion used for ion

exchange.

Acknowledgement

The authors would like to thank Yves Fontaine, Mireille Blais, Tetiana Elyart and Pont-Viau

drinking water treatment plant personnel for their support with piloting. We also acknowledge

Elise Renault, Julie Philibert, Jacinthe Mailly and Gabriel St-Jean for their assistance in sample

analysis. We also acknowledge Xiameng Feng for the review and helpful suggestions for this

paper.

55

2.5 Supplementary materials

Figure S2.1 - Average heterotrophic biomass measured by 14C glucose respiration test as a

function of media depth and empty bed contact time (EBCT) of the bicarbonate-form BIEX (BIEX-

B) and the chloride-form BIEX (BIEX-C) filters at 10 000 BV. Error bars represent the maximum and

the minimum biomass value of triplicate analysis.

Figure S2.2 - Disinfection by-product concentrations (B) in raw water, chloride-form BIEX (BIEX-

C) and bicarbonate-form BIEX (BIEX-B) effluents from 5 300 to 11 000 bed volumes (i.e., secondary

IX).

0

5

10

15

0

10

20

30

40

50

60

0 5 10 15 20

Empt

y be

d co

ntac

t tim

e (m

in)

Med

ia d

epth

(cm

)

Bacterial biomass (µg C/ml resin)

BIEX-BBIEX-C

Temperature: 25°C

BDOC: 0.25 mg C/L

Dissolved oxygen:

7.66 mg O2/L

Limit of detection:

0.5 µg C/mL resin

0

100

200

300

400

500

600

THM

-UFC

HAA

-UFC

THM

-UFC

HAA

-UFC

THM

-UFC

HAA

-UFC

Raw water BIEX-C BIEX-B

Disi

nfec

tion

by-p

rodu

ct (

µg/L

)

median

25-75 %

max-min

56

Figure S2.3 - Removals of natural organic matter (NOM) fractions in chloride-form BIEX (BIEX-C).

Figure S2.4 - Ammonia nitrogen concentration in raw water, chloride-form BIEX (BIEX-C) and

bicarbonate-form BIEX (BIEX-B) effluents from June to August (4 600 – 12 000 bed volumes).

Water temperature: 13-26 °C.

Table S2.1 - Raw water physicochemical characteristics (April. 2019 – January. 2020).

Parameters Average ± standard deviation

Units

Temperature 14.2 ± 8.0 °C

pH 7.2 ± 0.3 -

Primary IX


-100%

-75%

-50%

-25%

0%

25%

50%

75%

100%

0 5000 10000 15000 20000 25000

Rem

oval

Bed volumes

BP

HS

BB

LMW acids

LMW neutrals

0

5

10

15

20

25

30

35

Raw water BIEX-C BIEX-B

Amm

onia

nitr

ogen

(µg/

L)

median25-75 percentilesmax-min

Detection limit

57

Turbidity 13.3 ± 10.9 NTU

DOC 7.0 ± 0.7 mg C/L

UV254 0.24 ±0.03 cm-1

BDOC 0.42 ±0.15 mg C/L

Chloride 6.6 ± 3.2 mg/L

Sulfate 5.6 ± 1.5 mg/L

Nitrate 0.9 ± 0.5 mg/L

Nitrite < 0.2 mg/L

Bromide < 0.1 mg/L

Alkalinity 27.7 ± 7.4 mg CaCO3/L

Hardness 27.0 mg CaCO3/L

Ammonia 20.1 ± 8.5 µg NH3-N/L

Table S2.2 - Source water natural organic matter (NOM) fractions expressed in percentage of

dissolved organic carbon (DOC).

NOM fractions Concentration

(µg C/L) * DOC Percentage

Dissolved organic carbon

(DOC) 7001 ± 605 100%

Biopolymers

(BP) 199 ± 32 3%

Humics

(HS) 4852 ± 411 69%

Building blocks

(BB) 870 ± 41 13%

Low molecular weight acids (LMW acids) 191 ± 10 3%

Low molecular weight neutrals 511 ± 62 7%

58

(LMW neutrals)

*Average ± standard deviation.

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transformation during chlorination and chloramination of highly-polluted source water:

significance of ammonia. Water Res, 47(15), 5901-5910. doi:10.1016/j.watres.2013.07.013

USEPA. (1992). Method 524.2: Measurement of Purgeable Organic Compounds in Water by

Capillary Column Gas Chromatography/Mass Spectrometry, Revision 4.1. Cincinnati, OH.

USEPA. (2003). Method 552.3 Determination of haloacetic acids and dalapon in drinking water by

liquid-liquid microextraction, derivatization, and gas chromatography with electron capture

detection. Methods for the Determination of Organic Compounds in Drinking Water.

van den Brink, P., Satpradit, O. A., van Bentem, A., Zwijnenburg, A., Temmink, H., & van

Loosdrecht, M. (2011). Effect of temperature shocks on membrane fouling in membrane

bioreactors. Water Res, 45(15), 4491-4500. doi:10.1016/j.watres.2011.05.046







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301-310. doi:10.1016/j.watres.2018.01.052

Zhang, X. Y., Xiao, X. Z., Zheng, S. K., & Wang, W. (2019). Achieving no spent brine discharge in an

anion exchange resin-based fixed-bed process for typical DOM removal. Chemical Engineering

Research & Design, 149, 121-128. doi:10.1016/j.cherd.2019.07.008

64

Chapitre 3 – Removal of organic micropollutants from surface

waters by biological ion exchange resins

Abstract

Biological ion exchange (BIEX) refers to operating ion exchange (IX) filters with infrequent

regeneration to favor the microbial growth on resin surface and thereby contribute to the

removal of organic matter through biodegradation. However, the extent of biodegradation on

BIEX resins is still debatable due to the difficulty in discriminating between biodegradation and

IX. The objective of the present study was to evaluate the performance of BIEX resins for the

removal of organic micropollutants and thereby validate the occurrence of biodegradation. The

removals of biodegradable micropollutants (neutral: caffeine and estradiol; negative: ibuprofen

and naproxen) and nonbiodegradable micropollutants with different charges (neutral: atrazine

and thiamethoxam; negative: PFOA and PFOS) were respectively monitored during batch tests

with biotic and abiotic BIEX resins. Results demonstrated that biodegradation contributed to the

removal of caffeine, estradiol, and ibuprofen, confirming that biodegradation occurred on the

BIEX resins. Furthermore, biodegradation contributed to a lower extent to the removal of

naproxen probably due to the absence of an adapted bacterial community. The removal of

naproxen, PFOS, and PFOA were attributable to ion exchange with previously retained natural

organic matter on BIEX resins. Nonbiodegradable and neutral micropollutants (atrazine and

thiamethoxam) were minimally (6%-10%) removed during the batch tests. Overall, the present

study corroborates that biomass found on BIEX resins contribute to the removal of

micropollutants through biodegradation.

Keywords: Ion exchange resins; Biological ion exchange (BIEX); Biodegradation; Natural organic

matter (NOM); Micropollutants; PFAS

65

3.1 Introduction A large number of organic micropollutants such as pesticides, pharmaceuticals, steroidal

hormones, and industrial chemicals enter surface waters via stormwater runoffs and treated

wastewater discharge. As several of these micropollutants are persistent in the environment and

recalcitrant to drinking water treatment processes, many have been detected in treated drinking

water worldwide (Coupe and Blomquist, 2004; Wang et al., 2011; Husk et al., 2019). Although

micropollutants are often found at trace concentrations (i.e., ng/L to μg/L) in drinking water,

there has been a rising concern over the long-term effects on human health arising from the

exposure to a mixture of micropollutants. Consequently, the number of micropollutants to be

regulated in drinking water is on the rise. For example, the USEPA has recently announced its

regulatory determinations for 8 micropollutants out of the 109 contaminants found on the

Contaminant Candidate List 4 (USEPA, 2021).

Biologically active filtration (i.e., biofiltration) has been investigated as a treatment option for the

removal of organic micropollutants because it is a simple and environmentally friendly process

(Zearley and Summers, 2012; Benner et al., 2013). Granular filters loaded with adsorptive media

(e.g., granular activated carbon) or non-adsorptive media (e.g., sand and anthracite) gradually

convert into biofilters when no disinfectants are used during the operation, that is,

microorganisms will establish themselves on the surface of filtration media and develop a biofilm

(Zearley and Summers, 2012). Upon filtration, micropollutants can sorb onto the biofilm present

on the granular media surface and be subjected to biodegradation by direct catabolism or co-

metabolism (Benner et al., 2013). For instance, a wide variety of organic micropollutants proved

to be amenable to biofiltration, such as 2-methylisoborneo, geosmin, caffeine, carbamazepine,

diclofenac, and 2,4-D (Nerenberg et al., 2000; Reungoat et al., 2011; Zearley and Summers, 2012;

Nord and Bester, 2020).

In 2017, we proposed a novel biofiltration process using ion exchange (IX) resins as biomass

support for natural organic matter (NOM) removal (Schulz et al., 2017). Specifically, when

operating fixed bed IX filters without regeneration, microorganisms colonize the surface of resins

and develop a biofilm. This novel approach, referred to as biological ion exchange (BIEX), has been

66

demonstrated to achieve a higher NOM removal than conventional biological activated carbon

(BAC) filters in a lab-scale study (BIEX: 56% vs BAC: 15%) (Winter et al., 2018) and in a pilot-scale

study (BIEX: 62% vs BAC: 5%) (Amini et al., 2018). The superior performance achieved by BIEX was

later primarily attributed to ion exchange with sulfate and, to a lesser extent, biodegradation (Liu

et al., 2020; Zimmerman et al., 2021). However, given that previous studies were devoted to

investigating the performance of BIEX for NOM removal, its performance for the removal of

organic micropollutants has not yet been explored.

Mass balance studies were conducted for the BIEX filters to elucidate the mechanisms that come

into play (Amini et al., 2018; Liu et al., 2020). The authors reported that biomass contribution

accounts for a maximum of 30% of NOM removal in BIEX filters. However, the mechanism for

such contribution is still open to debate. Winter et al. (2018) reported that resins harvested from

the BIEX filters were biologically active, but the relation between biological activities and the

removal of NOM was not verified. Furthermore, DNA sequencing tests conducted with BIEX resins

suggested that heterotrophic bacteria are diverse in the BIEX filter (Edgar and Boyer, 2021), but

a direct proof of their contribution to the overall organic matter removal is yet to be made.

The overall objective of the present study was to evaluate the removal of organic micropollutants

by BIEX resins and thereby validate the biodegradation of organic micropollutants on BIEX resins.

Exhausted IX resins were harvested from a BIEX pilot and used for lab-scale batch tests where

biodegradable and nonbiodegradable micropollutants with different charges (neutral and

negative) were spiked into batch reactors to study their biotic and abiotic decay in the presence

of suspended BIEX resins. Suspect screening of transformation products was performed using

ultra-high-performance liquid chromatography coupled to high-resolution mass spectrometry

(UHPLC-HRMS). The present study reports the first investigation of the performance of BIEX resins

for organic micropollutants removal.

67

3.2 Materials and methods

3.2.1 Biological ion exchange (BIEX) resins characteristics Exhausted ion exchange resins were harvested from a pilot plant at the Pont-Viau drinking water

treatment plant (Laval, Canada) (Liu et al., 2021). The pilot plant consists of a fixed bed IX filter

loaded with polyacrylic type I macroporous strong base anion exchange resins (Purolite® A860,

Purolite, Philadelphia, USA). The IX filter was continuously fed with untreated surface water from

the Des Prairies River (Laval, Canada) and was operated without regeneration of the resins. After

5 months of operation, resins were harvested from the IX filter at the depth of 10 cm. At the

moment of IX resins harvesting, no anion release (e.g., chloride or sulfate) was observed in the

effluent at the harvesting location (data not shown), confirming that the resin capacity was fully

exhausted at the depth of extraction, and the IX resins used in the present study were mainly

loaded with NOM (Liu et al., 2021). Moreover, a 14C glucose respiration rate test conducted on

the extracted resins revealed that the resins had a heterotrophic biomass density of

approximately 15 µg C/mL resins (typical values for BAC are 9-14 µg C/mL) (Liu et al., 2021),

demonstrating that resins used in the present study were biologically active.

3.2.2 Raw water characteristics Surface water from the Des Prairies River was filtered in the lab through 0.45 μm disk membrane

(Supor® 450 PES, PALL, Port Washington, USA) and stored in the dark at 4°C before use. The water

demonstrated a moderate dissolved organic carbon (DOC) concentration (7.0 mg C/L), a neutral

pH, and low anion concentrations (chloride: 6.6 mg/L; sulfate: 5.6 mg/L). Other information on

raw water characteristics can be found in Table S3.1.

3.2.3 Target micropollutants Organic micropollutants were selected according to their biodegradability and charge state at pH

7. Biodegradable and nonbiodegradable micropollutants were selected to study the effect of

biodegradation. The biodegradability was predicted by the linear regression model and the non-

linear regression model, that is, BIOWIN and BIOWIN2 (EnviroSim Associates limited, Hamilton,

Canada). Generally, biodegradation probability ranges from 0 to 1 in the BIOWIN and BIOWIN2

68

models, with 0 suggesting the organic micropollutant is unlikely biodegradable and 1 suggesting

the organic micropollutant is easily biodegradable (Howard et al., 1992). In the present study,

micropollutants with a probability smaller than 0.5 were defined as nonbiodegradable and those

having a probability greater than 0.5 were defined as biodegradable. Furthermore, neutral and

negatively charged micropollutants were included in each group, because molecules with higher

affinity to the resin than previously retained anions can potentially be exchanged onto IX resins

even though the resin capacity is exhausted (Liu et al., 2020). Therefore, 8 organic micropollutants

including steroid hormones, pharmaceuticals, pesticides, and poly- and perfluoroalkyl substances

(PFAS) were selected (Table 3.1). Based on the selection metrics, these micropollutants were

divided into four groups: biodegradable and neutral (caffeine and estradiol), biodegradable and

negative (ibuprofen and naproxen), nonbiodegradable and neutral (atrazine and thiamethoxam),

as well as nonbiodegradable and negative (perfluorooctanoic acid (PFOA) and

perfluorooctanesulfonic (PFOS)).

Table 3.1 - Organic micropollutants selected for the present study.

Compound Structure pKaa Biowin1/B

iowin2b

Classification

(pH = 7)

Caffeine

(CAF)

-1.16

(strongest

basic)

0.6551/

0.5625

Biodegradable and

neutral

17b-Estradiol

(E2)

10.33 0.8178/

0.6452

Biodegradable and

negative Ibuprofen

(IBU) 4.85

0.8314/

0.8672

69

Naproxen

(NAP)

4.19 0.8972/

0.9611

Thiamethoxam

(THI)

0.4

(strongest

basic)

0.0789/

0.0014 Nonbiodegradable

and neutral Atrazine

(ATZ)

14.48 0.0045/

0.0000

Perfluorooctanoic

acid (PFOA) -4.2 -/0.0000

Nonbiodegradable

and negative Perfluorooctanesul

fonic acid (PFOS) -3.32 -/-

a: acid dissociation constant (pKa) data is cited from chemicalize.com (2020); b: data predicted using Estimation

Programs Interface Suite (US EPA, 2020); -: not available.

3.2.4 Batch tests Raw water was spiked with analytical grade organic micropollutants (Sigma Aldrich, St. Louis, MO)

to achieve a concentration of 10 µg/L for each selected micropollutant. Blank tests demonstrated

that micropollutant concentrations remained unchanged throughout the test (24 h), suggesting

that selected micropollutants were stable in the raw water (data not shown).

Batch tests were respectively conducted under two conditions, that is, (1) Biotic condition: 15 mL

BIEX resins were suspended in 241 mL spiked water within a 250 mL sterilized amber glass bottle;

(2) Abiotic condition: 15 ml BIEX resins were first incubated with 1.2 mL formaldehyde (37% w/v)

in a 250 ml sterilized amber glass bottle for 3h prior to the addition of 240 mL spiked water. Our

pretests demonstrated that formaldehyde inhibited microbial glucose respiration by 98%

subsequent to a 3-hour incubation using a volume ratio (media : formaldehyde) of 12.5 (Liu et al.,

70

2021). Notably, formaldehyde is a neutral substance that is not exchangeable on the BIEX resins

during the incubation.

For each condition of the batch test, 8 amber glass bottles were prepared corresponding to 6

contact times (i.e., 1, 3, 10, 30, 240, 1440 min), 1 quality control (QC) (i.e., a replicate for t = 1440

min) and 1 blank reactor (only containing spiked water). All bottles were shaken at 200 rpm on a

horizontal shaker at room temperature (22 °C). When reaching the desired contact time, samples

were immediately filtered through 0.22-μm pore size glass fiber syringe filters (Kinesis KX, Vernon

Hills, USA) and stored at 4 °C prior to micropollutant analyses. Glass fiber filters were selected

over other types of filters due to their superior recovery of the target micropollutants (Figure

S3.1).

3.2.5 Analytical methods Raw water and water samples subsequent to the batch tests were analyzed using on-line solid

phase extraction (SPE) followed by ultra-high-performance liquid chromatography equipped with

a reverse-phase column (C18, 100 mm ´ 2.1 mm, dp = 1.9 µm) coupled with high-resolution mass

spectrometry (UHPLC-HRMS) (Thermo Fisher Scientific, Waltham, MA). The results were used to

calculate micropollutant removal during batch tests and identify transformation products. More

details on the analytical method used for micropollutants analysis can be found in the

supplementary materials. All analyses were conducted in duplicate, and only mean values are

reported in this paper as data demonstrated a low variability (average relative standard deviation

< 3%).

3.2.6 Data analysis

3.2.6.1 Micropollutant removal kinetics

Batch tests data were fitted to a pseudo-first-order kinetics (commonly used to describe

biodegradation removal) and a pseudo-second-order kinetics (commonly used to describe

removal by sorption) through nonlinear regression. The normalized pseudo-first order kinetics

model can be expressed as:

71

""+= +2 = ,#3,4 (3.1)

where C0 and C are the initial micropollutant concentration (µg/L) and the micropollutant

concentration subsequent to the reaction time t (min), respectively; Cn is the normalized

micropollutant concentration; k1 is the pseudo-first-order rate constant (min-1).

Normalized sorption pseudo-second-order kinetic model can be expressed as follows.

-5. = (.#"")(""#")-/"+)(.#")-/"+)

(3.2)

where C0 and Ceq are the initial micropollutant concentration and the concentration at equilibrium

(µg/L), respectively; Cn is the normalized concentration; k2 is the pseudo-second order kinetics

rate constant (µg/(g.h)).

3.2.6.2 Suspect screening of transformation products

The identification of transformation products related to the selected organic micropollutants was

carried out following the same procedures as in our previous study (Solliec et al., 2021). Briefly,

an in-house database was first developed based on published scientific literature and includes 37

known transformation products related to the spiked micropollutants (Table S2). Then, Xcalibur

software was used to investigate the raw data spectra and identify the transformation products

based on the exact mass, the isotopic pattern, the fragmentation pattern, and the retention time.

Thereafter, the exact masses and MS/MS spectra of the potential transformation products were

compared with spectral database (i.e., METLIN and Mass Bank) or a MS-spectrum prediction

software (i.e., CFM-ID 3.0 - University of Alberta, Edmonton, Canada). This identification process

leads to a confidence level of 2 according to Schymanski et al. (2014). Finally, the concentration

of identified transformation products was estimated by semi-quantitation with the help of the

calibration curve of parent compounds. More details on the identification process are available

in the supplementary materials.

72

3.3 Results and discussion

3.3.1 Organic micropollutant concentrations in the raw water Raw water was monitored monthly to investigate the background concentrations of selected

organic micropollutants. Overall, prior to resin harvesting, caffeine, ibuprofen, atrazine, PFOA,

and PFAS were detected in raw water with a range of 4-166 ng/L whereas 17b-estradiol,

naproxen, and thiamethoxam were below detection limit (Table 3.2). The presence of some

biodegradable micropollutants in the source water likely resulted in the development of microbial

communities on BIEX resins which could degrade these micropollutants (Rolph et al., 2019). For

example, as IX resins were exposed to caffein-containing raw water during several months of

operation, caffeine-degrading bacteria may have developped on the BIEX resins. By contrast, as

naproxen was absent in the raw water, bacteria adapted to naproxen degradation were less likely

to be present on the BIEX resins.

Table 3.2 - Micropollutant concentrations in the raw water prior to resin harvesting. Values and

error bars respectively correspond to average and standard deviation of monthly measurements.

Groups Micropollutants Concentration in raw water (ng/L)

Biodegradable and neutral

Caffeine

(CAF)

166 ± 7

17b-Estradiol

(E2)

< 0.1

Biodegradable and negative

Ibuprofen

(IBU)

128 ± 4

Naproxen

(NAP)

< 0.1

Nonbiodegradable and

neutral

Thiamethoxam

(THI)

< 0.09

Atrazine 4 ± 2

73

(ATZ)

Nonbiodegradable and

negative

PFOA 4 ± 1

PFOS 50 ± 5

3.3.2 Micropollutant removal during batch tests Figure 3.1 summarized the removal of micropollutants in the biotic condition (i.e., BIEX with active

biomass) and the abiotic condition (i.e., BIEX media with inactivated biomass) during the batch

test. Concentrations at different contact times were normalized to the concentration of blank

samples (C/C0). Data were fitted to pseudo-first-order and pseudo-second-order kinetics to

determine fitted rate constants and coefficients (Table 3.3). In the following text, the analysis of

organic micropollutant removal is grouped according to the class of micropollutants.

Biodegradable and neutral micropollutants. No caffeine removal was observed under the abiotic

condition whereas a gradual decrease of caffeine was observed under the biotic condition. The

biotic removal of caffeine could be effectively modeled with a pseudo-first-order kinetics (R2 =

0.93). Given the fact that the only difference between the biotic and abiotic conditions was the

microbial activity, the results indicate that the removal of caffeine by the exhausted BIEX resins

was due to biodegradation. High caffeine removal through biodegradation has also been reported

for other types of biofilters (e.g., slow sand filtration) (Zhang et al., 2019; Hermes et al., 2019).

By contrast, estradiol was gradually removed over time under both abiotic and biotic conditions.

Estradiol removal under abiotic and biotic conditions were well described with a pseudo-second-

order kinetics (R2 = 0.68-0.69) even though estradiol removal under the abiotic condition was

better described with a pseudo-first-order kinetics (R2 = 0.92). A faster removal (k2: 0.34 vs 0.06)

was observed under the biotic condition compared to the abiotic condition, suggesting that

microbial activities contributed to the removal of estradiol through biodegradation. Abtahi et al.

(2018) reported a similar result where estradiol was removed to a higher extent under biotic

conditions than abiotic conditions for a moving bed biofilm reactor. Furthermore, complete

estradiol removal was observed under the abiotic condition at the end of the batch test (24 h),

74

revealing that the abiotic mechanism (i.e., sorption) also came to play. As estradiol is a neutral

compound, mechanisms other than IX must have contributed to its removal. Zhang and Zhou

(2004) reported that estradiol can be directly adsorbed onto the IX resin surface through Van der

Waals forces or hydrophobic interactions.

Biodegradable and negative micropollutants. Minimal removal (i.e., 14% after 24 h) was

observed for ibuprofen under the abiotic condition, even though ibuprofen is a negatively

charged compound. Meanwhile, complete removal was observed after 24 h in the biotic reactor.

The removal of ibuprofen under the biotic condition could be effectively modeled with a pseudo-

first-order kinetics (R2 = 0.99) whereas the removal under the abiotic condition could be modeled

with a pseudo-second-order kinetics (R2 =0.74). The greater extent of removal under the biotic

condition compared to the abiotic condition demonstrated that biodegradation is the main

mechanism for the removal of ibuprofen on BIEX resins whereas sorption/IX processes did not

significantly contribute to the removal of ibuprofen. This is possibly due to its low affinity for the

resins. Past studies also demonstrated that ibuprofen was removed to a greater extent in

biofilters than their abiotic counterparts (Paredes et al., 2016), suggesting that biodegradation is

a viable approach for removing ibuprofen from water.

Naproxen was removed to a similar extent under the abiotic and biotic conditions (normalized

concentration at 24 h = 0.51-0.62). The removals of naproxen under biotic and abiotic conditions

were effectively modeled with a pseudo-second-order kinetics (R2 = 0.95-0.97). The similar

removal in the biotic and abiotic conditions suggests that abiotic processes played a major role in

the removal of naproxen on BIEX resins. Given that the molecule is negatively charged, the

removal of naproxen is attributed to ion exchange with previously retained NOM on the resins.

Microbial activity had a minor contribution even though naproxen was initially defined as

biodegradable in the present study. We hypothesized that the low biodegradation of naproxen

was due to the absence of an adapted microbial community on the BIEX resins, which is consistent

with the fact that naproxen was not detected in the raw water. In summary, biodegradation

contributed to the removal of caffeine, estradiol, and ibuprofen but contributed minimally to the

removal of naproxen on the BIEX resins, revealing that the biodegradation of micropollutants on

75

the BIEX resins simultaneously depends on the characteristics of organic micropollutants and the

microbial communities colonizing the media.

Nonbiodegradable and neutral micropollutants. For both biotic and abiotic conditions, low

removals (6%-10%) were observed during batch tests for thiamethoxam and atrazine. Neither

pseudo-first-order or pseudo-second-order kinetics could be fitted to the data. The observed

removal was unlikely due to IX, as these compounds are neutral. Therefore, the minimal removal

was attributed to sorption processes. Previous studies also reported limited removal of

thiamethoxam and atrazine when using biofiltration for drinking water treatment (Hallé et al.,

2015; Gomez-Herrero et al., 2020), suggesting that alternative treatment would be needed to

remove these micropollutants.

Nonbiodegradable and negative micropollutants. Similar patterns were observed for PFOA

removal under abiotic and biotic conditions which suggests that abiotic processes were the

dominant removal mechanism. PFOA data under abiotic and biotic conditions were better

modeled with a pseudo-second-order kinetics, respectively (R2 = 0.74-0.83). Similarly, PFOS was

removed to a similar extent under the biotic and abiotic conditions. The removals of PFOS under

abiotic and biotic conditions were also better modeled with a pseudo-second-order kinetics,

respectively (R2 = 0.70-0.77). As expected, biodegradation offered a minimal contribution to the

removal of PFOA and PFOS due to their strong resistance to biodegradation (Liou et al., 2010; Li

et al., 2019). On the other hand, ion exchange is well known to be an effective method to remove

PFOA and PFOS (Dixit et al., 2019). Therefore, the removal of PFOA and PFOS is most likely

achieved through ion exchange due to their higher affinity than previously retained NOM on the

resins. In summary, batch tests with nonbiodegradable micropollutants confirmed that negative

micropollutants can be removed by BIEX resins through ion exchange if the contaminants have a

higher affinity than previously retained anions on the resin.

76

Figure 3.1 - Micropollutants removal during batch tests under biotic and abiotic conditions.

Trendlines correspond to the kinetic model (pseudo-1st or pseudo-2nd order) fitted to the data.

0.0

0.2

0.4

0.6

0.8

1.0

1.2

1 100 10000

Caffeine Estradiol

Ibuprofen Naproxen

Thiamethoxam

PFOA PFOS

0.0

0.2

0.4

0.6

0.8

1.0

1.2

1 100 10000

Atrazine

Biodegradable and neutral

Biodegradable and negative

Nonbiodegradable and negative

Nonbiodegradable and neutral

0.0

0.2

0.4

0.6

0.8

1.0

1.2

1 100 10000

C/C

0

0.0

0.2

0.4

0.6

0.8

1.0

1.2

1 100 10000

0.0

0.2

0.4

0.6

0.8

1.0

1.2

1 100 100000.0

0.2

0.4

0.6

0.8

1.0

1.2

1 100 10000

Biotic

Abiotic

Modeling(Biotic)Modeling(Abiotic)

0.0

0.2

0.4

0.6

0.8

1.0

1.2

1 100 10000

Time (min)

0.0

0.2

0.4

0.6

0.8

1.0

1.2

1 100 10000

77

Kinetic parameters of the fitted models are summarized in Table 3. Error bars were omitted due

to overlapping with data points.

Table 3.3 - Rate constants (k) and coefficients of determination (R2) for the kinetic models of

micropollutant removal during batch tests. Fitting with R2 lower than 0.25 was designated as not

available.

Micropollutants Test

Condition

Pseudo-first order Pseudo-second order

k1 (min-1) R2 k2 (µg/(g.h)) R2

Caffeine Biotic 0.0007 0.93 0.002 0.49

Abiotic Not available Not available

Estradiol Biotic 0.002 0.63 0.34 0.68

Abiotic 0.003 0.92 0.06 0.69

Ibuprofen Biotic 0.003 0.99 0.008 0.95

Abiotic Not available 4.2 0.74

Naproxen Biotic 0.0003 0.46 0.24 0.95

Abiotic Not available 0.85 0.97

Thiamethoxam Biotic Not available Not available


Atrazine Biotic Not available Not available


PFOA Biotic 0.0002 0.52 0.41 0.83

Abiotic 0.0002 0.66 0.26 0.74

PFOS Biotic 0.0008 0.58 0.32 0.70

Abiotic 0.0007 0.59 0.18 0.77

78

Micropollutant concentrations after 24-h batch tests are summarized in Figure 3.2. Estradiol was

removed to the greatest extent under either abiotic or biotic conditions due to sorption

processes. Caffeine and ibuprofen were only efficiently removed under the biotic condition

(normalized concentration: 0.3 and 0.02, respectively) thanks to the presence of acclimated

biomass on the BIEX resins. Minimal difference was observed for naproxen (p > 0.05, t-test),

PFOA, and PFOS between biotic and abiotic conditions as they were mainly removed through ion

exchange with previously retained NOM. Atrazine and thiamethoxam were not amenable to BIEX

resins as sorption processes, biodegradation, and ion exchange are all inefficient for their

removal. Overall, the order of organic micropollutant removal extent under the biotic condition

is estradiol > ibuprofen > PFOS > caffeine > naproxen > PFOA > atrazine > thiamethoxam.

Figure 3.2 – Normalized micropollutant concentrations after 24 h batch test. Error bars

correspond to the minimum and maximum values between test groups and controls. CAF:

caffeine; E2: Estradiol; IBU: ibuprofen; NAP: naproxen; THI: thiamethoxam; ATZ: atrazine; PFOA:

perfluorooctanoic acid; PFOS: perfluorooctanesulfonic acid.

3.3.3 Suspect screening of transformation products Two transformation products were identified from the in-house database using the suspect

screening approach (Figure S3.2-3.7). The concentrations of theobromine and 1-

hydroxyibuprofen during the batch test were estimated using the standard curve of the parent

0.0

0.2

0.4

0.6

0.8

1.0

1.2

CAF E2

IBU

NAP

THI

ATZ

PFOA

PFOS

C/C

0

AbioticBiotic

Biodegradable Nonbiodegradable

79

ions (i.e., caffeine and ibuprofen) (Table 3.4). Theobromine was identified as a potential

biotransformation product of caffeine as it mainly results from biodegradation processes (Suzuki

and Waller, 1984). Theobromine was not found in the raw water and the abiotic condition except

for the contact time of 1440 min (i.e., 24h), whereas it was observed throughout the batch test

conducted under the biotic condition. No clear trend was observed for the variation of

theobromine under the biotic condition, hypothetically owing to the low concentrations

measured. The contrasting results of theobromine under the abiotic and biotic conditions

corroborate that biodegradation is an active mechanism for the removal of caffeine on the BIEX

resins.

By contrast, 1-hydroxyibuprofen was observed under both biotic and abiotic conditions (Table

3.4). Previous studies reported that 1-hydroxyibuprofen was a biotransformation product of

ibuprofen (Salgado et al., 2020). In the present study, the presence of 1-hydroxyibuprofen in the

abiotic condition was due to the biodegradation of ibuprofen in the raw water (i.e., before

micropollutant spiking) according to our suspect screening results (concentration estimated: 20

ng/L). No significant difference can be observed between the abiotic and the biotic condition

(p>0.05) possibly due to the fast biodegradation of 1-hydroxyibuprofen under the biotic

condition. Rutere et al. (2020) also reported minimal concentration detected for 1-

hydroxyibuprofen in their ibuprofen biodegradation study, suggesting that 1-hydroxyibuprofen is

a biotransformation intermediate of ibuprofen.

Table 3.4 - Estimated concentrations of theobromine and 1-hydroxyibuprofen during the batch

test.

Contact time (min)

Estimated concentration (ng/L)

Transformation products Abiotic condition Biotic condition

Theobromine

1 Not found 7 ± 3

3 Not found 3 ± 2

10 Not found 8 ± 1

30 Not found 3 ± 5

80

240 Not found 1 ± 1

1440 3 ± 3 5 ± 2

1 20 ± 5 37 ± 3

3 29 ± 3 25 ± 4

10 16 ± 1 38 ± 6

1-hydroxyibuprofen 30 31 ± 8 26 ± 6

240 31 ± 20 50 ± 3

1440 36 ± 8 34 ± 24

3.3.4 Implications on the application of ion exchange resins Conventional use of ion exchange resins for NOM removal implies frequent regeneration with

high concentration of NaCl solution (8%-12%w/v). The regeneration produces a highly

concentrated brine of which the disposal can be costly and difficult (Levchuk et al., 2018).

Although regeneration allows recovering resin exchange capacity, it also removes biofilms on the

resin surface. Operating ion exchange resins with a lower regeneration frequency, i.e., in

biological ion exchange (BIEX) mode, is an alternative to the conventional strategy of operation

for NOM removal applications, given that NOM can still be removed through secondary ion

exchange (i.e., exchange with sulfate) subsequent to the exhaustion of primary ion exchange

capacity (i.e., exchange with chloride) (Liu et al., 2020; Edgar and Boyer, 2020; Zimmerman et al.,

2021). With less frequent regeneration, microbial communities can fully develop on the resin

surface. Although past studies have stated that bacteria may bind the ion exchange sites and

potentially decrease the treatment efficiency (Wachinski, 2006; Edgar and Boyer, 2020), the

present study demonstrated that biomass on resins can also contribute to the removal of

biodegradable micropollutants. Therefore, ion exchange resins are a promising alternative to

conventional media (e.g., activated carbon) as a biomass support in biofiltration processes. Future

studies will need to systematically compare the cost-efficiency of using ion exchange resins as

biomass support compared to conventional media for biofiltration. The first full-scale application

81

of BIEX has been put in operation in Middle River (BC, Canada) in 2019. Monitoring its

performance is underway to confirm the superiority of BIEX over BAC filtration for both NOM as

well as organic micropollutant removal.

3.4 Conclusions

The present study evaluated micropollutant removal using biological ion exchange (BIEX) resins.

Results indicate that biodegradation can contribute to the removal of caffeine, estradiol, and

ibuprofen, corroborating that biodegradation is a removal mechanism on BIEX resins. However,

the extent of biodegradation also depends on the presence of an adapted microbial community

on the resin surface given that we observed a lower contribution from biodegradation for the

removal of naproxen which was not present in the source water. The removals of naproxen, PFOA,

and PFOS were attributable to ion exchange with previously retained natural organic matter

(NOM) on the resins. Finally, nonbiodegradable and neutral compounds, that is thiamethoxam

and atrazine, were not effectively removed by the BIEX resins, suggesting that alternative

treatment options should be considered to target these molecules. Future study should compare

the cost-benefit of IX resins versus the other traditional media (e.g., sand, anthracite, and

activated carbon) used for biofiltration.

Acknowledgements

We would like to thank the technical support of Sung Vo Duy, Jacinthe Mailly, Julie Philibert, and

Yves Fontaine. This work was completed at part of the NSERC-Industrial Chair in Drinking Water

at Polytechnique Montréal with the financial support of its partners, namely City of Montréal,

Veolia Water Technologies Canada Inc., City of Laval, City of Longueuil and the City of Repentigny.

All experiments were completed at CREDEAU laboratories, a research infrastructure supported

by the Canadian Foundation for Innovation (CFI).


82

Table S3.1 - Raw water characteristics

Parameters Unit Average

pH - 7.2

Turbidity NTU 13.3

Dissolved organic matter (DOC) mg C/L 7.0

UV254 cm-1 0.24

Biodegradable DOC (BDOC) mg C/L 0.42

Biopolymers (BP) μg C/L 218

Humic substances (HS) μg C/L 4769

Building Blocks (BB) μg C/L 866

Low molecular weight acids (LMW acids) μg C/L 189

Low molecular weight neutrals (LMW neutrals) μg C/L 536

Chloride mg/L 6.6

Sulfate mg/L 5.6

Nitrate mg/L 0.9

Nitrite mg/L < 0.2

Bromide mg/L < 0.1

Alkalinity mg CaCO3/L 27.7

Hardness mg CaCO3/L 27.0

Ammonia μg NH3-N/L 20.1

Filter Selection. Spiked water and filtrated spiked water were analyzed in triplicate using UHPLC-

HRMS to evaluate the micropollutants recovery rates for each type of filters (Figure S3.1). Overall,

the glass fiber filter (GFF) showed to have the best recovery rates for all selected micropollutants

with an average of 98 ± 3 %.

83

Figure S3.1 - Micropollutant recovery of different types of filters.

Micropollutants analysis by UHPLC/HRMS. Quantitative analysis of micropollutants was

performed with an on-line solid-phase extraction coupled to ultra-high performance liquid

chromatography (UHPLC) and high-resolution mass spectrometry (HRMS). A Thermo Scientific

Dionex UltiMate 3000 pump (Thermo Fisher Scientific, Waltham, MA) was used for on-line SPE

coupled to a second Thermo Scientific Dionex UltiMate 3000 RS pump (Thermo Fisher Scientific,

Waltham, MA) for chromatographic separation controlled by Chromeleon 7.2 Software (Thermo

Fisher Scientific, Waltham, MA). The Thermo Scientific Dionex Ultimate 3000 Series TCC-3000RS

column compartment is equipped with a Hypersil GOLDTM C18 column (100 mm × 2.1 mm, dp =

1.9 µm) and preceded by a guard filter cartridge (5 mm × 2.1 mm, 0.2 µm porosity) (Thermo Fisher

Scientific, Waltham, MA) which was used for reversed-phase chromatographic separation of

selected compounds at 50 °C. On-line SPE was performed by a Hypersil Gold column (20 x 2mm,

dp = 12 µm). A PAL system RTC autosampler was used (Zwingen, Switzerland) for injection, and

samples were kept at 4 °C in a fresh compartment before injection.

0

20

40

60

80

100

120

140

160

CAF THIATZ

KET IBUNAP E2

PFOA

PFOS

Reco

ver r

ate

(%)

Cellulose acetate Glass fiber Nitrocellulose Nylon Polycarbonate PVDF

84

The on-line SPE pre-concentration step was done as follows. 1 mL of water samples was injected

in the injection loop with a loading speed set at 1 mL/min. Then, the column is washed with milli-

Q water (with 0.1% HCOOH) during 2 min at 1 mL/min. Following the sample loading step, the

SPE column was then back flushed with MeOH during 5 min at 1.5 mL/min (with 0.1% HCOOH)

and the elution was transferred using the analytical pump gradient directly through the analytical

column. The chromatographic separation was achieved with mobile phase A, which consisted of

Milli-Q water and mobile phase B, which consisted of MeOH (analytical grade) with 0.1 mM NH4F

(prepared daily) at a flow rate of 0.450 uL/min. The gradient of separation was starting and kept

at 40% B for 2 min, and then the MeOH was increased to 100% from 2 to 7 min and was held

constant at 100% for 1 min. Then, the mobile phases were brought back to initial conditions and

maintained for 2 min for column conditioning before the next injection. A heated electrospray

ionization interface (HESI-II) operated in positive/negative switching mode was used for the

ionization of selected and potential transformation product compounds. Standard parameters

were chosen so as not to limit the ionization of candidates during the chromatographic separation

course. The ionization spray voltage was set at +/-4000V; capillary temperature was set at 275 °C;

the vaporizer temperature was set at 300 °C; sheath gas and auxiliary gas flow were set at 50 and

30 arbitrary units, respectively. HRMS detection was performed by a quadrupole-Orbitrap (Q-

Exactive) mass spectrometer controlled by the Xcalibur 3.0 software (Thermo Fisher Scientific,

Waltham, MA). Instrument calibration in positive mode was done every 7 days using both LTQ

Velos ESI Positive and Negative Ion Calibration Solution (Pierce Biothechnology Inc. Rockford, IL)

by direct infusion to get mass accuracies for all target compounds within the 5 ppm range during

analysis. The Q-Exactive parameter values of resolving power, automatic gain control (AGC), and

ion time (IT) were chosen to improve selectivity and sensitivity during the measurement of

samples and were set respectively at 3x106 ion capacity and 50 ms filling. Each sample was

analyzed in full scan acquisition (m/z 150-600) with a resolving power of 70,000 FWHM (m/z 200).

In order to identify potential transformation products, the data-dependent acquisition (DDA)

mode was running on samples to collect fragmentation information. The scan cycle of the DDA

mode was composed of a full scan acquisition (m/z 150-600) with a resolving power of 70,000

FWHM (m/z 200) and fragmentation events with a resolving power of 17,500 FWHM (m/z 200).

85

For the fragmentation mode, the AGC was set at 2x105 ion capacity and the IT parameter were

set at 100 ms. The intensity threshold was fixed at 1x104 to trigger the fragmentation with

stepped NCE set at 20, 30 and 40 (arbitrary unit). The precursor ions were filtered by the

quadrupole at an isolation width of 0.4 FWHM at m/z 200. The top ten most abundant ions were

selected for fragmentation and a single loop count were used for multiplexing.

Table S3.2 - Potential transformation products of selected micropollutants based on the

literature.

Transformation products Associated compound

Elemental composition Ionization [M+H]+

or [M-H]-

Diethylatrazine Atrazine C6H10ClN5 Positive 188.07030

Deisopropylatrazine Atrazine C5H8ClN5 Positive 174.05465

Diaminoatrazine Atrazine C3H4ClN5 Positive 146.02335

Hydroxyatrazine Atrazine C8H15N5O Positive 198.13549

Clothianidine Thiamethoxam C6H8ClN5O2S Positive 250.01655

Paraxanthine Caffeine C7H8N4O2 Positive 181.07255

Theobromine Caffeine C7H8N4O2 Positive 181.07255

Theophylline Caffeine C7H8N4O2 Positive 181.07255

N-nitrosodimethyl amine Caffeine C2H6N2O Positive 75.05583

N-nitrosomethylethylamine Caffeine C3H8N2O Positive 89.07148

N-nitrosodiethylamine Caffeine C4H10N2O Positive 103.08713

N-nitrosodibutylamine Caffeine C8H18N2O Positive 159.14973

N-nitrosopiperidine Caffeine C5H10N2O Positive 115.08713

N-nitrosomorpholine Caffeine C4H8N2O2 Positive 117.06640

N-nitrosopyrrolidine Caffeine C4H8N2O Positive 101.07148

Estrone Estradiol C18H22O2 Negative 269.15415

2-Hydroxyestrone Estradiol C18H24O2 Negative 271.16981

4-Hydroxyestrone Estradiol C18H22O3 Negative 285.14907

4-Methoxyestrone Estradiol C19H24O3 Negative 299.16472

16α-Hydroxyestrone Estradiol C18H22O3 Negative 285.14907

17α-Epiestriol Estradiol C18H24O3 Negative 287.16472

Estriol Estradiol C18H24O3 Negative 287.16472

17-α-ethinylestradiol Estradiol C20H24O2 Negative 295.16980

2-Methoxyestrone Estradiol C19H24O3 Negative 299.16472

17ß-Estradiol-3- glucuronide Estradiol C24H32O8 Negative 447.20189

Estradiol-3-sulfate Estradiol C18H24O5S Negative 351.12662

Ibuprofen glucuronide Ibuprofen C19H26O8 Negative 381.15494

2-Hydroxyibuprofen Ibuprofen C13H18O3 Negative 221.11777


86

Carboxy-ibuprofen Ibuprofen C14H18O4 Negative 249.11268


2-Phenylpropanoic acid Ibuprofen C9H10O2 Negative 149.06025

Isobutylbenzene Ibuprofen C10H14 Negative 133.10172

Propionic acid Ibuprofen C3H6O2 Negative 73.02895

6-O-Desmethylnaproxen Naproxen C13H12O3 Negative 215.07082

Desmethylnaproxen-6-O-sulfate Naproxen C13H12O6S Negative 295.02763

7-Hydroxynaproxen Naproxen C14H14O4 Negative 245.08138

Identification and semi-quantification of theobromine as a transformation product of caffeine.

Two chromatographic peaks were recorded at the exact mass of m/z 181.07255 with mass

accuracies ≤ 5 ppm in positive mode. The feature measured at 2.97 min had a mass accuracy of

2.25 ppm for C7H9N4O2+, while the parent ion [M+H]+ acquired at 4.01 min had a mass accuracy

of -1.67 ppm for C7H9N4O2+. Thus, these features may be attributed to one of the transformation

products of caffeine (i.e., paraxanthine, theophylline, and theobromine) since they have the same

exact masses. However, theobromine has a higher Kow value compared to paraxanthine and

theophylline (Log Kow: -0.05 for theobromine versus -0.39 for paraxanthine and theophylline)

(KOWWIN v1.67 estimate), suggesting that theobromine would have a longer retention time than

paraxanthine and theophylline. Then, both the fragmentation patterns provided by the DDA

mode at 2.97 and 4.01 min were studied to further confirm the presence of theobromine in the

samples. The feature at 2.97 min did not show fragment ions of theobromine, nor paraxanthine

or theophylline in its fragmentation spectra. While fragmentation data of the feature at 4.01 min

showed fragment ions of theobromine at m/z 123.08044 and 135.11662 that match with

fragmentation mass spectra of theobromine. Other key product ions (i.e., m/z 110.07153 and

138.06687) of theobromine were measured at lower intensities in the fragmentation spectra.

Therefore, chromatographic peak at 4.01 min was attributed to theobromine.

The semi-quantification of theobromine was based on the [M+H]+ chromatographic peak of

caffeine and by the use of an external standard calibration curve. The quality control of the results

was assessed in terms of retention time reproducibility and mass accuracy. The retention time

did not vary (RSD < 1%) and the exact mass of the [M+H]+ remained ≤ 5 ppm for all the samples.

87

Figure S3.2 - Chromatographic separation at m/z 181.07255 (positive mode).

Figure S3.3 - Isotopic pattern at m/z 181.07255 (positive mode).

RT: 0.00 - 9.00 SM: 7G

0 1 2 3 4 5 6 7 8 9Time (min)

0

5

10

15

20

25

30

35

40

45

50

55

60

65

70

75

80

85

90

95

100

Rel

ativ

e A

bund

ance

4.01

2.97

NL: 1.27E5m/z= 181.07074-181.07436 F: FTMS + p ESI Full ms [50.0000-600.0000] MS PI_DDA

PI_DDA #1321 RT: 2.97 AV: 1 NL: 1.03E6T: FTMS + p ESI Full ms [50.0000-600.0000]

180.5 180.6 180.7 180.8 180.9 181.0 181.1 181.2 181.3 181.4 181.5 181.6m/z

0

5

10

15

20

25

30

35

40

45

50

55

60

65

70

75

80

85

90

95

100

Rel

ativ

e A

bund

ance

181.08617

181.12263181.06084

181.00073

88

Figure S3.4 - Fragmentation spectra at 4.01 min for m/z 181.07255 with specific fragments of

theobromine.

Identification and semi-quantification of 1-hydroxyibuprofen as a transformation product of

ibuprofen. One chromatographic pic was observed at the exact mass of m/z 221.11666 in

negative mode. The ion measured at 5.09 min had an accuracy of 7 ppm for C13H18O3 and

corresponds to 1-hydroxyibuprofen. The fragmentation pattern provided by the DDA mode at

5.09 min was studied to further confirm the presence of 1-hydroxyibuprofen in the samples. The

feature at 5.09 min showed the fragments of 1-hydroxyibuprofen at m/z 59.01258, 149.09671,

and 177.12813. Thus, the pic at 5.09 was attributed to 1-hydroxyibuprofen, and the concentration

was estimated using the external standard calibration curve of the parent ion (i.e., ibuprofen).

PI_DDA #1727 RT: 3.90 AV: 1 NL: 3.35E5F: FTMS + p ESI d Full ms2 [email protected] [50.0000-390.0000]

60 80 100 120 140 160 180 200 220 240 260 280 300 320 340 360 380m/z

0

5

10

15

20

25

30

35

40

45

50

55

60

65

70

75

80

85

90

95

100

Rel

ativ

e Ab

unda

nce

123.08044

149.02307

181.12134

163.03857105.03374

135.11662

93.07027

83.0497269.03420 316.95996 379.94791358.52509221.85918

89

Figure S3.5 - Chromatographic separation at m/z 221.11666 (negative mode).

Figure S3.6 - Isotopic pattern at m/z 221.11666 (negative mode).

RT: 0.00 - 9.01 SM: 7G

0 1 2 3 4 5 6 7 8 9

Time (min)

0

5

10

15

20

25

30

35

40

45

50

55

60

65

70

75

80

85

90

95

100

Relative A

bundance

5.09

3.843.38

4.734.11

5.343.10

5.93

2.31

NL: 1.43E6

m/z=

221.11666-

221.11888 F: FTMS

- p ESI Full ms

[50.0000-600.0000]

MS NI_DDA

NI_DDA #2102 RT: 5.09 AV: 1 NL: 2.16E6T: FTMS - p ESI Full ms [50.0000-600.0000]

220.6 220.8 221.0 221.2 221.4 221.6 221.8 222.0 222.2 222.4 222.6 222.8 223.0 223.2 223.4 223.6m/z

0

5

10

15

20

25

30

35

40

45

50

55

60

65

70

75

80

85

90

95

100

105

110

115

120

Rel

ativ

e Ab

unda

nce

221.11852

222.12181

223.13416

90

Figure S3.7 - Fragmentation spectra at 5.09 min for m/z 221.11666 with specific fragments of 1-

hydroxyibuprofen.

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exchange capable of sulphate-based secondary ion exchange during long-term DOC removal.

Water Res, 196, 117036. doi:10.1016/j.watres.2021.117036

95

Chapitre 4 – Biological ion exchange as an alternative to

biological activated carbon for drinking water treatment

Abstract

Biological ion exchange (BIEX) has proved to remove natural organic matter (NOM) better than

biological activated carbon (BAC). This raises the question if BIEX can be integrated into a full-

scale drinking water treatment plant to remove NOM and ammonia. In this study, a pilot plant

consisting of one BIEX filter, three GAC filters and one BAC filter was set up as second-stage

filtration at the Sainte-Rose drinking water treatment plant (Laval, Canada). The pilot plant was

operated for a period of nine months without regeneration of the ion exchange resins. The

influent water showed low DOC (2.5 mg/L) and high sulfate concentrations (28.2 mg/L). Except

for a short peak of DOC released at about 1 000 BV, the BIEX filter achieved a nearly constant

removal of 29 - 36 % over the whole study period. The DOC removals of GAC were similar to BIEX

at < 8000 BV but then stabilized at 13 - 24 % after 8 000 BV. Most DOC removal in the BIEX filter

was achieved at the top 30 cm layer (81 %) compared to 62 - 66 % removal in the GAC/BAC filters

in the same layer. After the rapid exhaustion of the primary ion exchange capacity (< 1 000 BV),

sulfate displaced the fraction of NOM with lower affinity than sulfate, corresponding to the initial

DOC release in the BIEX filter. The fraction of NOM with higher affinity than sulfate can still replace

sulfate, which explains the good long-term performance of the BIEX filter. BIEX released ammonia

with an average of 15 % in warm water condition, probably related to the small diameter of the

column which limited backwash effectiveness.

Keywords: biological ion exchange; biological activated carbon; natural organic matter; anion

exchange resin; nitrification

96

4.1 Introduction Removing natural organic matter (NOM) is one of the main objectives during drinking water

treatment as NOM may cause taste, odor and color issues (Edzwald 2010, Thurman 2012),

membrane and filter fouling (Gibert et al. 2013, Kennedy et al. 2008, Kim and Dempsey 2013),

formation of disinfection by-products (DBPs) (Krasner et al. 2006, Richardson et al. 2007, Brezinski

et al. 2019) as well as biofilm growth in the distribution system (Hijnen et al. 2018). Water

treatment plants (WTPs) currently employ different techniques for NOM removal, including

coagulation/flocculation (Matilainen et al. 2010), activated carbon adsorption (Velten et al. 2011),

and biofiltration (Korth et al. 2001). Among these options, NOM removal through biofiltration has

received considerable attention due to its low maintenance cost and ease of operation. Rapid-

rate biofilters can be designed using various filtration media to act as a support for the

development of biomass. Among them, granular activated carbon (GAC) has been shown to offer

better performance compared to inert medias (e.g., sand, anthracite), an advantage related to

their sorptive capacity. After the exhaustion of this capacity, the GAC transitions into the so-called

biological mode (i.e. biological activated carbon or BAC) which can also offer high NOM removal

in cold water due to their higher surface area or porosity available for biomass growth and the

potential bio-regeneration of the adsorption capacity (Basu et al. 2016). Nevertheless, NOM

removal with BAC filtration is fairly limited (5 % - 20 % DOC removal) and the kinetics is highly

impacted by temperature (Terry and Summers 2018). For this reason, biofiltration is not typically

used as the sole NOM removal process within a treatment train unless the source water naturally

offers a low NOM content.

Recently, we have reported that ion exchange (IX) resins can be used as an alternative media for

biofiltration (Schulz et al. 2017). Firstly, IX resins remove NOM by anion exchange with a counter-

ion (usually chloride). Then, similar to a GAC filter, a progressive depletion of exchange sites is

observed while the surface of the resin is colonised by biofilm which participates in NOM

reduction via biodegradation. During a lab-scale study (Schulz et al. 2017), IX filters were fed with

a 5 mg C/L surface water with low mineral anion content for a period of one year. We observed

an average DOC removal of 60 % for an IX filter and 40 % for an IX filter made abiotic through

continuous sodium azide injection. The development of biofilm was proposed as the reason for

97

this increased performance and the process was referred to as biological ion exchange or BIEX. In

a follow-up lab-scale study using the same source water, BIEX was shown to provide superior DOC

removals (56 ± 7 %) compared to a BAC filter (15 ± 5 %) (Winter et al. 2018). This performance

was confirmed in a pilot study (Amini et al. 2018), where four types of filters (IX, BIEX, GAC and

BAC) were fed for 331 days (without regeneration of BIEX) using a river water of high DOC (7 - 8

mg C/L) and low anionic mineral content (5 - 8 mg SO4-2/L). During summer conditions after five

months of operation (≈840 bed volumes), BIEX achieved 62 % DOC removal whereas the BAC

filter achieved only 7 %. Nitrification in the BIEX filter was as efficient as in the BAC filter,

confirming that the BIEX media was biologically active. IX was exhausted significantly later than

GAC (90 d vs. < 7 d, respectively). While a carbon mass balance calculated after 331 days of

operation showed that about 31 % of DOC removal in the BIEX filter was due to biodegradation,

the high performance of BIEX could not be explained by biodegradation alone. Secondary ion

displacements (e.g., NOM displacing sulfate) was proposed as an important mechanism to explain

BIEX performance after initial anion exchange process with chloride was exhausted. However,

more studies are needed to elucidate the competition of NOM with sorbed anions after the

exhaustion of the initial IX capacity.

Resin suppliers consider biomass growth on resin a nuisance as it may cause excessive fouling in

the IX filter (Flemming 1987). Hence, conventional IX are regularly regenerated using a highly

concentrated salt solution (usually 10 - 12 % w/v NaCl solution) to regain the ion exchange

capacity and alleviate biofouling phenomena. Regeneration produces a brine containing high

concentrations of sodium chloride and NOM which has to be disposed of (Levchuk et al. 2018).

Direct disposal of brine to the aquatic environment or to the sewer is prohibited in many

jurisdictions, since high concentrations of chloride can cause corrosion of plumbing, have

detrimental effects on biological wastewater treatment processes and can be toxic to the aquatic

fauna and flora (Rokicki and Boyer 2011). BIEX provides a novel ion exchange operation strategy

since it implies greatly reducing the frequency of regeneration. Operating IX in BIEX mode could

also potentially lengthen the service life of the resin, which is strongly dependent on the number

of regeneration cycles (Hochmuller 1984).

98

Up to now, we have only evaluated BIEX as a standalone process using surface waters with low

mineral content as we targeted applications in small rural systems given the simplicity of

operation. The high performance achieved so far has raised the question as whether BIEX could

also be an attractive option for large-scale WTPs. For such application, it is anticipated that BIEX

filters would be located after coagulation and settling. Therefore, a lower influent DOC would be

expected compared to the conditions tested so far. If alum is used as coagulant, high sulfate

concentrations in the feed water are expected which may be in competition with NOM for IX sites.

Thus, the general objective of this study was to investigate the potential of a BIEX filter to replace

a second-stage BAC filter as a polishing process for simultaneous NOM removal and nitrification.

Specifically, we compared BIEX with four alternative GAC (three fresh and one exhausted). NOM,

DBP precursors, ammonia removals as well as head loss rates served as metrics to compare their

performances.

4.2 Materials and Methods

4.2.1 Pilot location and source water characteristics This study was conducted for a period of nine months at the Sainte-Rose drinking water treatment

plant (Laval, Canada). With a daily production capacity of 110 000 m3, the plant is fed by the Mille-

Iles River and employs coagulation (alum), flocculation, sedimentation, sand-anthracite filtration,

ozonation, BAC filtration and chlorination. We set up pilot filters after ozonation, in parallel to the

full-scale BAC filtration. The influent to the pilot showed low turbidity (» 0.5 NTU) and low DOC

(» 2.5 mg C/L) (Table 4.1). The concentration of sulfate (» 28.2 mg/L) was high due to the use of

alum as a coagulant (average dosage » 40 mg /L). This condition of operation corresponds to a

sulfate/DOC ratio of 11.2 (28.2 mg/L : 2.5 mg/L). These characteristics contrasted strongly with

our previous assay (Amini et al. 2018) where a BIEX filter was fed with raw waters with a DOC of

7 - 8 mg C/L and 5 - 8 mg/L of sulfate which translates into a sulfate/DOC ratio roughly ten times

lower (0.6 - 1.1).

99

Table 4.1 - Influent water characteristics (average ± standard deviation) during the pilot study

period (April 11, 2018 to January 07, 2019)

Parameters Unit Value*

Temperature °C 14.1 ± 9.2

Turbidity NTU ~ 0.5

DOC mg C/L 2.5 ± 0.2

UVA254 cm-1 0.019 ±

0.007

SUVA L/mg×m ~ 0.76

pH - 6.5 ± 0.4

Chloride mg/L 7.7 ± 4.4

Sulfate mg/L 28.2 ± 4.5

Alkalinity CaCO3 mg/L 12.1 ± 7.7

Ammonia µg N/L 78.0 ± 26.0 *average ± standard deviation

4.2.2 Pilot plant design and operation The pilot plant consisted of five parallel columns (CPVC, 5.08 cm diameter), each being filled with

a 30 cm sand sublayer (0.6 L) and a 150 cm top layer (3.0 L) of either GAC, BAC or IX (fig. 4.1). The

five top layer media investigated in this study (Table 4.2) were: 1) virgin, type-I macroporous

strong base anion exchange resin with an acrylic quaternary amine backbone in chloride form

(Puroliteâ A860, IX capacity: 0.8 eq/L), 2) three virgin wood-based activated carbons (Calgon

Acticarboneâ BGX, Nucharâ WV-B30, Jacobi PICABIOL-HP120) and 3) exhausted activated carbon

(Jacobi PICABIOL-2). The latter was recovered from a full-scale BAC filter at the Chomedey

drinking water treatment plant (Laval, Canada), where it had been in operation for 2 years. Only

wood-based or lignite GAC media were tested during this study as past research conducting at

this location had shown their superiority over mineral-based GAC with respect to nitrification.

Four sampling nozzles were located on each of the columns at different depths (10 cm, 30 cm, 60

cm, 150 cm) to profile the column performance (fig. 4.1). The columns were operated under

100

pressure and were equipped with manometers to monitor head loss accumulation. Given that IX

resins have low resistance to ozone and that the columns were fed with post-ozonated waters,

we added a pre-filter containing 5 cm of fresh GAC (Jacobi PICABIOL-HP120) with an empty bed

contact time (EBCT) of 10 seconds in order to destroy any residual ozone that may be present in

the influent. The interference on influent water characteristics was tested and found to be

negligible (average DOC removal during study ≤ 3 %, average ammonia removal during study ≤ 1

%, data not shown).

The filtration rate was kept constant at 10 m/h (10.8 min EBCT) which is a typical filtration design

rate for a second stage filter. The columns (including the GAC pre-filter) were backwashed every

two weeks using first air injection (3 min at 10 m/h) and then unchlorinated BAC filtered water

from the plant. Backwash water flowrate was adjusted to yield a 50 % bed expansion. The

duration of the backwash was media specific as it was conducted until the backwash waters

turbidity was lower than 10 NTU. Typically, backwash lasted 40 min for the BIEX filter and 20 min

for the GAC and BAC filters.

Figure 4.1 - Schematic overview of the pilot plant consisting of five parallel columns: BIEX, three

GAC filters and one BAC filter. The pilot is fed with coagulated/settled/filtered/ozonated surface

waters from the Sainte-Rose drinking water treatment plant.

PumpPump

BIEX GAC1 GAC2 GAC3 BAC

O3destructor

Ion exchange resin Activated carbonSandWater level

Legend

Sampling point

10 cm30 cm

60 cm

150 cm

180 cm

M M M M M

101

Table 4.2 - Media characteristics for IX, three different GAC and BAC used in this study

Abbreviations Media Diameter

(mm)

D10

(mm) Cu

Surface area

(m2/g)

BIEX Puroliteâ A860 0.3-1.2 0.5 1.7 2.19

GAC1 Calgon Acticarboneâ BGX 0.5-0.7 0.6 1.8 1550-1650

GAC2 Nucharâ WV-B30 0.8-1.1 1.0 1.8 1400-1600

GAC3 Jacobi PICABIOL-HP120 1.2-1.4 1.3 1.4 1400

BAC Jacobi PICABIOL-2 1.2-1.4 1.3 1.4 1600

Cu: Uniformity coefficient = d60/d10; D10: measured with a MASTERSIZER 3000 (Malvern, UK);

Other information was provided by suppliers.

4.2.3 Analytical methods Influent and effluent water characteristics were sampled weekly for the study period.

Temperature, pH, and dissolved oxygen were measured on-site with a multimeter (HACH HQ40D).

DOC and UVA254 samples were first prefiltered with a syringe filter (0.45 µm Merck Millex®-HV)

prior to analysis on a TOC-meter (Sievers 5310C, GE water, USA) or a spectrophotometer

(Ultrospec 3100pro, GE Healthcare, USA). Anions (nitrate, chloride, sulfate) were measured with

an ion chromatograph (ICS 5000 AS-DP DIONEX, USA) equipped with an AS18 column. Bicarbonate

concentration was evaluated from alkalinity measurements which were performed using the acid

titration method 5.10 (USEPA 2012). THM and HAA5 precursors were measured after chlorination

according to the Uniform Formation Conditions technique (UFC) (Summers et al. 1996), i.e., by

maintaining a free chlorine residual of 1.0 ± 0.5 mg Cl2/L after a contact time of 24h at pH 8.0 and

20 °C. They were analyzed by gas chromatography (7890B GC system from Agilent Technologies,

USA) according to methods 524.2 (THM) and 552.3 (HAA) (USEPA 2003). Ammonia concentrations

(in µg N/L) were quantified in triplicate using the indophenol colorimetric method NF T90-015

(AFNOR 2000) with a 5 cm spectrophotometric cell in order to achieve a detection limit of 5 µg

N/L.

102

4.3 Results

4.3.1 Head loss accumulation Head losses were recorded after (i) 30 min of operation following a backwash (D1), (ii) seven days

of operation (D7) and (iii) fourteen days of operation (D14) (fig. 4.2). The BIEX filter was composed

of a finer media which explains why it generated the highest head loss compared to GAC and BAC

filters. The median BIEX head losses were 1.48 m for D1, 3.44 m for D7 and 3.60 m for D14, which

translates into a head loss accumulation of 2.12 m after fourteen days. As a comparison, GAC1,

which had higher head loss than the other GAC, exhibited a median head loss of only 0.72, 0.73

and 0.81 m for D1, D7 and D14 conditions, respectively (accumulation of 0.09 m after fourteen

days). The other filters GAC2, GAC3 and BAC had lower head losses with values between 0.36 -

0.59 m for all the monitored conditions with an accumulation of only 0.11 - 0.16 m after fourteen

days. Overall, the head loss accumulation after 14 days was low and very similar for all GAC/BAC

media tested. On the other hand, the head loss of BIEX filter after fourteen days of operation

fluctuated between 2.19 m (minimum) and 5.63 m (maximum) during the study. Such head losses

are high compared to the typical maximum allowable head loss (2.4 m) in a gravity-fed granular

media filter according to (Kawamura 2000). This also explains why we decided to operate the

filters under pressure rather than by gravity. Finally, we observed that most of the head loss

accumulation in the BIEX filter occurred during the first week of operation, a phenomenon which

differed from the other GAC/BAC filters. We suspect that the BIEX media was not fully compacted

after 30 min of operation following the backwash.

103

Figure 4.2 - Head loss accumulation for the BIEX, three GAC and the BAC filter after (i) 30 min of

operation following a backwash (D1), (ii) seven days of operation (D7) and (iii) fourteen days of

operation (D14).

4.3.2 NOM removal

During the study period, the average DOC concentration and UVA254 of the influent were 2.5 ±

0.2 mg C/L and 0.019 ± 0.007 cm-1 (average ± standard deviation), respectively. We illustrated the

normalized DOC concentration (Ceffluent/Cinfluent) and the normalized UVA254 absorbance

(UVeffluent/UVinfluent) as a function of the number of bed volumes (BV) of filtered water (fig. 4.3 &

S4.1). Each day of operation was equivalent to about 133 BV. The performance of the filters was

divided into three phases. (1) NOM release from BIEX and initial GAC exhaustion (0 to 8 000 BV):

BIEX first showed a 50 % DOC release compared to the influent after one week of operation (» 1

000 BV). After this initial DOC breakthrough, BIEX removed DOC at an average of 29 % from 1000

to 8000 BV. During this period, all GAC filters achieved higher performances (GAC1: 48 %, GAC2:

39 %, GAC3: 32 %, on average) while the BAC filter (i.e., exhausted GAC) achieved 10 % DOC

removal. (2) From 8 000 to 25 000 BV (mid-June to mid-October): During this period, the water

temperature rose progressively from 18°C to 26°C in July-August before it dropped to 13°C in

October. The BIEX filter achieved a higher DOC removal than GAC/BAC filters during this period

with an average of 36 % DOC removal. All GAC filters exhibited a similar performance with average

0

1

2

3

4

5

6

D1 D7 D14 D1 D7 D1

4 D1 D7 D14 D1 D7 D1

4 D1 D7 D14

Headloss(mofwater)

Operation days

BIEX GAC1 GAC2 GAC3 BAC

Min-Max

25 %-75 %

Median

104

DOC removals of 24 %, 22 %, 21% for GAC1, GAC2 and GAC3, respectively. These performances

were slightly superior to the BAC filter which achieved 17 % removal during this period, an

indication that the GAC filters still exhibited a small adsorption capacity. (3) From 25 000 to 37

000 BV (October-January): During this period, the temperature declined from 13 to 1°C. The BIEX

filter maintained an average DOC removal of 30 %. In contrast, the average DOC removals

declined to 16 %, 13 % and 15 % for GAC1, GAC2 and GAC3, respectively, while the BAC filter

achieved 13 % removal. In summary, apart from the initial DOC release observed at about 1 000

BV, BIEX offered equivalent performance to GAC (< 8 000 BV) or superior (> 8 000 BV) to GAC or

BAC filtration. Among the three tested GAC, GAC1 showed higher performance than GAC2 and

GAC3 (p < 0.05) while GAC2 and GAC3 showed no significant difference with regards to DOC

removal (p > 0.05).

Figure 4.3 - Normalized DOC concentrations (Ceffluent/Cinfluent) in the effluents of the BIEX filter,

three GAC filters and one BAC filter for the study period from April 11, 2018 to January 07, 2019.

With respect to UV254 absorbance (fig. S4.1), removals fluctuated as they were directly impacted

by the ozonation conditions preceding the pilot filters. The ozone residual concentration in the

influent varied from a minimum of 0.07 mg O3/L to a maximum of 0.6 mg O3/L (according to the

0

5

10

15

20

25

30

0.0

0.2

0.4

0.6

0.8

1.0

1.2

1.4

1.6

0 5000 10000 15000 20000 25000 30000 35000 40000

Tem

pera

ture

(°C

)

Nor

mal

ized

con

cent

ratio

n of

DO

C (C

efflu

ent/C

influ

ent)

Bed Volume

BIEX GAC1

GAC2 GAC3

BAC Temp

Days of operation

Cinfluent=2.20 ±0.18 mg/L

Cinfluent=2.44±0.14 mg/L

Cinfluent=2.77 ±0.12 mg/L

0 50 100 150 200 250 300

105

on-line monitor of the plant). Nevertheless, the trend in UV254 removal in the BIEX filter exhibited

a similar pattern to the one of DOC removal. For example, BIEX surpassed the other filters at

about 8000 BV and became the most efficient media until the end of the study. For the entire

study period, BIEX provided the best UVA254 removal with an average of 48 % compared to GAC1

(31 %), GAC2 (28 %), GAC3 (25 %) and BAC (15 %). The fact that the UVA254 in BIEX effluent was

sustained for a very long period, independent of water temperature, suggests that removal of

aromatic NOM through ion exchange was an active mechanism throughout the study.

We realized a profile study to monitor DOC removal across the filter beds at about 21 000 BV =

154 days of operation (fig. 4.4). In the BIEX filter, the upper 30 cm layer (1.8 min EBCT) made up

for 81 % of the DOC removal while the other 150 cm (9 min EBCT) accounted for the remaining

19 %. However, in the case of the activated carbon filters, all filters showed similar patterns where

62 - 66 % of the total DOC removal was achieved in the top 30 cm while the remainder (34 - 38

%) was removed in the lower 150 cm layer. The rapid removal kinetic of BIEX is more consistent

with ion exchange than biodegradation. This finding also indicates that the BIEX filter would be a

more robust process than BAC filtration given that NOM is removed more effectively inside the

filter.

Figure 4.4 - Distribution of total DOC removal (0.87, 0.76, 0.73, 0.71, 0.52 mg C/L) in the upper

(30 cm) vs. lower layer (150 cm) for BIEX, GAC1, GAC2, GAC3 and BAC, realized at about 21 000

BV.

0%

20%

40%

60%

80%

100%

BIEX GAC1 GAC2 GAC3 BAC0-30 cm 30-180 cm

106

4.3.3 Removal of THM and HAA precursors From August 22, 2018 to January 7, 2019 (18 000 BV to 37 000 BV), we monitored THM and HAA5

precursors biweekly in the influent and effluents (fig. 4.5). The average influent THM-UFC and

HAA5-UFC concentrations were 60 µg/L and 68 µg/L, respectively. BIEX achieved the highest THM

and HAA precursors reductions with an average of 48 % and 66 %, respectively. Meanwhile, all

GAC filters had similar performance with average reductions of 28 - 33 % for THM-UFC and 45 -

48 % for HAA5-UFC. Removal of THM and HAA precursors averaged 28 % and 44 %, respectively

in the BAC. The precursors removal performance was consistent with the observed DOC and

UVA254 removals for the five filters, confirming the higher NOM removal performance in the BIEX

filter.

(A) (B)

Figure 4.5 - Distribution of (A) THM-UFC and (B) HAA5-UFC precursors in the influent and effluents

of the filter media under investigation. Period: 18 000 to 32 000 BV. Numbers indicate the average

concentrations.

4.3.4 Impact of BIEX on inorganic anions The major inorganic anions concentrations were monitored weekly in the influent and BIEX

effluent. The average influent concentrations (± standard deviation) of sulfate, chloride,

bicarbonate and nitrate for the study period were: 28.2 ± 4.5 mg/L, 7.7 ± 4.4 mg/L, 14.8 ± 9.4

mg/L and 1.1 ± 0.6 mg/L, respectively. Given that ion exchange resin was initially charged with

chloride, the average concentration of chloride in the BIEX effluent was higher than in the influent

(9.1 ± 11.7 mg/L). In order to understand the dynamics of ion exchange occurring on the resins,

20

30

40

50

60

70

80

Influent BIEX GAC1 GAC2 GAC3 BAC

THM-UFC

(µg/L)

min-max

25 %-75 %average

60μg

/L31

μg/L 38

μg/

L

41 μg

/L

40 μ

g/L

42 μg

/L

10

20

30

40

50

60

70

80

90

Influent BIEX GAC1 GAC2 GAC3 BAC

HAA5-UFC

(µg/L)

68 μg

/L23

μg/

L 34 μ

g/L

36 μg

/L

34 μg

/L

36 μg

/L

min-max

25 %-75 %average

107

the loading of anions on the resin was calculated using a weekly cumulative charge balance

(expressed as eq/L) as described by equation (4.1):

#(%, ') = ∑ ∑ !"!",!,$#"%&',!,$$×'×()*()*!"

+,0+,.

/,09/,. (4.1)

where q(i,j) is the total cumulative loading of the anion sorbed on the resin after week i (expressed

as eq/L of resin). The index j describes the three solutes considered in the mass balance (DOC,

sulfate and chloride). Cin,i,j and Cout,i,j describe the concentrations of the solute j during week i in

the influent and effluent, respectively. Q is the flow (486 L/d) while V is the volume of resin in the

column (3 L). Bicarbonate and nitrate were neglected from the calculation as they were found not

to be significantly removed in the BIEX filter. DOC charge density was assumed to be 10 meq/g C,

a value representative of a low SUVA (0.7 L/mg×m) NOM at pH 6.5 based on Boyer et al. (2008).

Figure 4.6 presents the result of the cumulative loading of the three studied solutes on the resin

during the study. The initial capacity of the fresh resin was 0.80 eq/L present under the form of

chloride. During the first week of operation, this chloride was replaced by sulfate which reached

0.76 eq/L. In other words, the high concentration of sulfate in the influent almost “regenerated”

the IX resin from chloride-form into a sulfate-form within 1 000 BV. At this point, the resin was

therefore fully exhausted with respect to chloride release, a period (≈ 1 000 BV) which

corresponded to the significant DOC release (3.31 mg C/L) measured in the effluent. We postulate

that the incoming sulfate displaced the fraction of NOM on the resin with a lower affinity than

sulfate. Interestingly, the sulfate concentration on the resin started to decline after 1 000 BV

whereas the DOC and chloride concentrations on the resin started to rise. At the end of the study,

the DOC and chloride concentrations on the resin were about 0.33 and 0.10 eq/L, respectively

while the sulfate concentration had decreased to about 0.19 eq/L resin. The total loading of the

three studied solutes equals to 0.62 eq/L resin (lower than IX capacity 0.68 eq/L resin), indicating

that other anions excluded from this study also occupied a portion of IX position on the resin

phase. The variation in sulfate concentration was not monotonous and most likely reflects the

impact of variations in influent water characteristics. Finally, we do not know if the DOC would

have eventually broken through once all the sulfate had been released from the resin. However,

figure 4.6 suggests that the resin would have been fully loaded with NOM and chloride at approx.

108

40 000 BV. Considering that IX filters are typically regenerated after 48 - 72 h of operation (500-1

000 BV), the BIEX mode of operation would translate into very low regeneration costs.

Figure 4.6 - Evolution of the anion concentrations on the resin during 34 000 BV.

4.3.5 Removal of ammonia

During the study period, the average concentration (± standard deviation) of ammonia in the

influent was 78 ± 26 µg N/L. We monitored the effluent ammonia concentration of the five filters

(fig. 4.7). Before 11 weeks (» 10 000 BV), BIEX, GAC1, GAC2 and GAC3 showed negligible ammonia

removals with averages of 0 %, 9 %, 13 % and 10 %, respectively. Meanwhile, the BAC filter

released ammonia with an average of 19 % during this period. We postulate that this may have

been due to a period of the nitrifying biomass reacclimating to its new environment given that it

was collected from a first stage biofilter. As water temperature continued to rise, GAC1 and GAC2

were the first filters to nitrify (11 weeks » 10 000 BV) while GAC3 and BAC followed the same

trend one week later (12 weeks » 11 000 BV). From 15 to 25 weeks (14 000 - 24 000 BV), all GAC

and BAC filters provided excellent ammonia removals (98 - 99 %) under warm water condition

(17 - 26°C). On the other hand, the BIEX filter released an average 15 % ammonia from weeks 11

to 25, i.e., during summer conditions. After 25 weeks (» 24 000 BV), as temperature declined

below 15°C, ammonia removal first declined in the BAC filter which eventually completely lost

treatment capacity after 32 weeks or 4 weeks of operation below 10°C. The GAC1, 2 and 3 filters

also suffered from the decline of temperature but were still nitrifying 14%, 49% and 32%

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

0 10000 20000 30000 40000

Ani

on c

once

ntra

tion

on th

e re

sin

(eq/

L re

sin)

Bed Volume

DOCChlorideSulfate

0 50 100 150 200 250 300Days of operation

Initial chloride release

IEX capacity = 0.8 eq/L

109

respectively after 35 weeks (or 7 weeks below 10°C). Meanwhile, the BIEX continued to show no

ammonia removal during this period. Overall, for the entire study period, GAC2 had the best

performance with an average removal of 68 % while GAC1 and GAC3 showed an equivalent (p >

0.05) performance, with an average of 59 % removal.

Figure 4.7 - Normalized ammonia concentration in the effluents of the BIEX filter, three GAC filters

and one BAC filter for the study period from April 11, 2018 to January 7, 2019.

4.4 Discussion

4.4.1 Head loss in the BIEX filter The BIEX filter had the highest head loss compared to GAC and BAC filters mainly due to the

smaller resin bead size. Operation in a pressurized vessel is thus advisable. Alternatively, the use

of larger IX bead size could be an option to mitigate head loss, although using larger beads may

extend the mass transfer zone inside the resin particles (Ball and Harries 1989). Future research

will also need to address the design conditions required to properly backwash BIEX filters. The

development of a crust in the upper layer is a phenomenon that will need to be managed with

proper air/water backwash and probably more frequent backwash than what was done during

this study (i.e., once every 2 weeks) to reduce the head loss accumulation in BIEX filters.

0

5

10

15

20

25

30

0.0

0.2

0.4

0.6

0.8

1.0

1.2

1.4

1.6

0 5000 10000 15000 20000 25000 30000 35000 40000

Tem

pera

ture

(°C

)

Nor

mol

ized

con

cent

ratio

n of

am

mon

ia(C

efflu

ent/C

influ

ent)

Bed Volume

BIEX GAC1GAC2 GAC3BAC Temp

110

4.4.2 NOM removal mechanisms in the BIEX filter In our previous study (Amini et al. 2018), we have evaluated that the BIEX filter operated under

summer conditions removed as much as 62 % of NOM, using a high DOC/low inorganic anions

surface water. The observed performance was much higher compared to a BAC filter (7 %). In this

study, DOC removal in the BIEX was once again higher than what was achieved in the GAC/BAC

filters investigated, despite being fed with a low DOC/high sulfate pre-treated water. The sulfate

concentration in the influent is an important factor for IX processes, as (i) high sulfate

concentration can lead to a reduction in DOC removal (Ates and Incetan 2013, Dixit et al. 2018),

and (ii) they can reverse the NOM preference from low molecular weight (MW) species to high

MW ones (Tan and Kilduff 2007). However, Verdickt et al. (2012) found that the DOC removal

efficiency did not reduce significantly using an IX regenerated with sulfate (33 %) compared to a

conventional IX with chloride as counter-ion (42 %). Also, using IX in sulfate form can avoid

unnecessary anions exchange, since anions having lower affinities towards IX compared to sulfate

(chloride, bicarbonate, nitrate), cannot be further exchanged. This finding indicates that IX can

still be used for NOM removal even with a high concentration of sulfate in the influent due to the

exchange between sulfate and the fraction of NOM with higher affinity than sulfate.

Figure 4.8 presents a simplified schematic overview of the ion exchange dynamics occurring

during the long-term operation of the BIEX media (assuming that only NOM, sulfate and chloride

are the dominant solutes to consider). The affinity of each anionic species towards IX depends on

two factors. On the one hand, the intrinsic property of the anion molecule will affect its affinity

for the anion exchange resin. For example, several authors found that charge density (Boyer et

al. 2008, Finkbeiner et al. 2019), hydrophobicity (Li and SenGupta 1998) and molecular size (Bazri

and Mohseni 2016) can affect the affinity of NOM molecules. On the other hand, the

concentration of anions can also affect their competitiveness, since the more molecules are

present, the higher are the chances that the species would occupy the ion exchange site. Given

that NOM is an assemblage of molecules with variable charge, molecular weight and

hydrophobicity, we divided NOM into three different groups based on their affinities for IX. The

first fraction NOM1 has no affinity for the resin (e.g., the unexchangeable fraction) and is

therefore expected to breakthrough at time 0. The remaining NOM fractions are referred to as

111

NOM2 and NOM3 which respectively describes the NOM fractions with either a lower or higher

affinity than sulfate. Initially, the resin is loaded with chloride ions (fig. 4.8a) which are exchanged

for NOM3, sulfate and NOM2 (fig. 4.8b). Once the chloride capacity is exhausted (» 1 000 BV in

this study), NOM2 is displaced by sulfate (fig. 4.8c). Meanwhile, NOM3 continues to displace

sulfate which explains the long-term performance of the resin for NOM removal (fig. 4.8d). The

unexchangeable NOM (NOM1) was 0.48 mg C/L (22 % of the NOM in the influent). The removal

efficiency stabilized at about 29 % (NOM3) after the DOC release peak, which means that NOM2

accounted for about 49 % of the overall NOM in the influent. Even though NOM1 and NOM2 were

no longer removable after the DOC breakthrough, BIEX filter can still yield 29 - 36 % DOC removal

until the end of the study. Such phenomenon had already been reported by Fu and Symons (1990)

who observed that an anion exchange filter was still removing DOC after breakthrough due to

displacement of sulfate. Typically, IX filters are regenerated near the sulfate breakthrough to

avoid organic matter leakage (Kim and Symons 1991), an operation strategy requiring frequent

regeneration and thus producing a concentrated brine. In this study, the IX filter achieved a

prolonged NOM removal after the sulfate breakthrough mainly due to the exchange between

sulfate and the NOM fraction with higher affinity than sulfate. This performance was similar or

better than those of GAC/BAC filters for a period of at least 35 000 BV. Hence, we suggest 1)

setting IX filters off-line during a short period (e.g., 800 – 900 BV) to avoid the NOM leakage; and

2) operating IX filters beyond the sulfate breakthrough to remove NOM by secondary ion

exchange as well as biodegradation to avoid frequent regeneration (i.e., BIEX mode).

112

Figure 4.8 - Displacement of NOM fractions in the BIEX filter a) virgin IEX; b) NOM3, sulfate and

NOM2 replace chloride while NOM1 is nonexchangeable; c) NOM3 and sulfate replace NOM2

leading to the DOC release in the BIEX filter; d) NOM3 replaces sulfate, which explains the long-

term performance of NOM removal in the BIEX filter. N.B. The anion on each band presents the

dominant species but not the only one.

A mass balance was performed at 34 000 BV to confirm the proposed mechanism of NOM removal

according to equation (4.2).

LNOM = DOC × PNOM3 × Vwater × CD (4.2)

where LNOM is the cumulative loading of NOM on the resin (eq/L resin) due to the IX mechanism.

DOC is the average DOC concentration in the influent during the study (2.5 mg C/L). PNOM3 is the

percentage of NOM3 in the influent NOM (about 29 %), a fraction that can continually accumulate

on the resin. Vwater is the volume of water considered in the mass balance (34 000 BV). CD is the

charge density of NOM3 (9.0 - 10.6 meq/g C), a range estimated based on Boyer et al. (2008) at

pH 6.5 (transphilic acids excluded). Consequently, the cumulative loading of NOM due to the IX

mechanism at 34 000 BV is estimated to be 0.22 - 0.26 eq/L resin. A similar calculation was

performed for sulfate (28.2 mg/L), which showed that the breakthrough of sulfate is expected at

1 362 BV, similar to the observed sulfate breakthrough during this study (approx. 1 000 BV).

However, we obtained a higher loading value of NOM (0.33 eq/L resin) according to equation (1),

Cl-

Cl-

SO42-

NOM2

NOM3

SO42-

NOM2

NOM3

SO42-

NOM3

Cl-NOM1

NOM1NOM2

SO42-NOM1NOM2

a) b) c) d)

113

a calculation that includes all mechanisms of NOM removal in the BIEX filter. Therefore, NOM

removal in addition to IX mechanism is estimated to be 0.07 - 0.11 eq/L resin (i.e., 21 - 33 % of

the total NOM removal), which is speculated to be due to the biomass contribution (e.g.,

biosorption, biodegradation, bioregeneration). However, more studies are still needed to

elucidate the contribution of biodegradation during the operation of IX filters in BIEX mode.

From our previous study (Amini et al. 2018), NOM1, NOM2 and NOM3 are respectively estimated

to be 20 %, 3 % and 77 % of the overall influent NOM, a different breakdown compared to the

current study (22 %, 49 %, 29 %) which is mainly due to the sulfate/DOC concentration in the

influents. In this study, we had a higher sulfate concentration in the influent compared to Amini

et al. (2018) (28.2 mg/L vs. 5 - 8 mg/L), which leads to a higher proportion of NOM2 and lower

proportion of NOM3. This higher proportion of NOM2 favored a net DOC release whereas the

lower proportion of NOM3 results in a lower NOM removal performance after DOC breakthrough

compared to our previous study (29 % vs. 62 %). To conclude, higher sulfate concentrations in the

influent can increase the NOM2 proportion and decrease NOM3 proportion to the overall influent

NOM, which then translates into a higher DOC breakthrough and lower NOM removal efficiency

after DOC breakthrough. Although we hypothesize that NOM2/NOM3 ratios are mostly

determined by the competition (i.e., sulfate/DOC ratio), the NOM characteristics can also play an

important role. For example, Kim and Symons (1991) found that the NOM fraction with < 0.5 K

MW was badly removed throughout the column test (NOM1) while 0.5 - 1 K and 1 - 5 K MW

fractions were the major organic fractions that surged during DOC breakthrough (NOM2).

However, more studies are needed to elucidate to what extent other NOM characteristics (charge

density, hydrophobicity) also play a role in the sulfate/NOM competition.

Evaluating the coupling of ozonation with BIEX was of interest in our study. Firstly, ozonation

fractionates higher MW organic matter with a corresponding increase in lower MW organic

matter (Amy et al. 1988). This shall facilitate the diffusion of organic matter into the resin pores

and reduce the size-exclusion phenomenon. In addition, ozonation also enhances NOM

biodegradability (Hozalski et al. 1999), and thus should enhance biodegradation on the BIEX

media. However, IX resins are incompatible with dissolved ozone. In this study, we employed a

GAC filter with a short EBCT as a preventive measure to protect BIEX. In a full-scale application, it

114

would be more cost-effective to quench the ozone residual chemically or plan for a longer

retention time before entering into a BIEX filter in order to let the ozone residual decay naturally.

4.4.3 Ammonia release in the BIEX filter In our previous study (Amini et al. 2018), we found that BIEX had a similar ammonia removal

efficiency compared to BAC while in this study we obtained results where the BIEX filter released

15 % ammonia in warm waters. Our results also demonstrate that this ammonia release

originated from the top 10 cm layer where most NOM was removed (fig. S4.2), and a thick

biomass layer was observed (fig. S4.3). Also, during the biweekly backwash of the BIEX filter, air

injection could barely break down the solid crust of the thick biomass layer. The crust broke into

smaller IX resin agglomerates, a phenomenon mainly due to the insufficient air injection rate and

small column diameter. Further, the influent water characteristics in the former study (untreated

surface waters) may also explain the successful nitrification obtained in 2018, since pH, alkalinity,

micronutrients (P, Cu etc.) were more favorable to nitrification than in our present study.

4.5 Conclusion

This was the first pilot study evaluating the application of BIEX as a second-stage filter within a

drinking water treatment plant fed with low DOC/high sulfate pretreated surface water. We

highlight the following findings:

• Despite the 50 % DOC release observed at about 1 000 BV, BIEX achieved similar (< 8 000

BV) or higher performance (> 8 000 BV) compared to GAC/BAC filters with 29 - 36 % DOC

removal.

• In the BIEX filter, after the rapid exhaustion of primary ion exchange capacity (< 1 000 BV),

sulfate replaced the fraction of NOM with lower affinity than sulfate (NOM2) leading to

the DOC release.

• In the BIEX filter, the exchange between the fraction of NOM with higher affinity than

sulfate (NOM3) and sulfate is the dominant mechanism to explain the long-term

performance of NOM removal.

115

• Within the top 30 cm layer (1.8 min EBCT), BIEX filter achieved 81 % of the total DOC

removal whereas GAC/BAC filters realized only 62 - 66 % at the same depth.

• BIEX released ammonia in warm water conditions with an average of 15 %, a phenomenon

may due to the small diameter of the column and the improper backwash.

Future studies will need to address the design conditions as well as backwash strategies to

properly wash a BIEX filter.

Acknowledgement

The authors would like to thank Yves Fontaine and Mireille Blais for their support in pilot plant

installation. We also acknowledge Julie Philibert, Jacinthe Mailly, Gabriel St-Jean for the

assistance of chemical analysis. We appreciate Sainte-Rose drinking water treatment plant for

their site support to the pilot plant. Finally, we acknowledge the CREATE Program in

environmental decontamination technologies and integrated water and wastewater

management (TEDGIEER) for the Ph.D. Scholarship awarded to Zhen Liu.


Figure S4.1 - Normalized UVA254 (UVAout/UVAin) in the effluents of the BIEX filter, three GAC filters

and one BAC filter for the study period from April 11, 2018 to January 7, 2019.

0.0

0.2

0.4

0.6

0.8

1.0

1.2

1.4

1.6

0 5000 10000 15000 20000 25000 30000 35000 40000

Nor

mal

ized

abs

orba

nce

of U

VA25

4(U

VAef

fluen

t/UVA

Inff

uent

)

Bed Volume

BIEX GAC1

GAC2 GAC3

BAC

116

Ammonia removal profile study. Ammonia removal profiles (fig. S4.2) were also performed in

warm waters (at week 21 » 20 000 BV). In the BIEX filter, we identified that ammonia release

originated from the upper 10 cm (0.6 min EBCT). In this section of the media, a thick layer of

biofilm was detectable visually (fig. S4.3). This layer proved to be difficult to breakdown during

backwash due to the small diameters of the column which limited backwash effectiveness. For

the other filters, nitrification was completed within the first 60 cm (3.6 min EBCT) in GAC2, GAC3

and BAC and even faster in GAC1 where ammonia was fully removed within the upper 30 cm (1.8

min EBCT).

Figure S4.2 - Ammonia removal as function of media depth (or EBCT), profile realized in warm

waters (week 21, » 20 000 BV, T = 21°C) for BIEX filter, three GAC filters and one BAC filter.

0 2 4 6 8 10

0

0.2

0.4

0.6

0.8

1

1.2

1.4

0 30 60 90 120 150 180

EBCT (min)

Nor

mal

ized

con

cent

ratio

n of

am

mon

ia

(Cpr

ofil/C

influ

ent)

Media Depth (cm)

BIEXGAC1GAC2GAC3BAC

SandBIEX, BAC or GAC

117

Figure S4.3 - A thick layer of biomass formed on the top of BIEX media, which proved to be difficult

to breakdown during the backwash.

Ion exchange resin shrinking. IX resin bead size was visually smaller after nine months of

operation. We previously measured the Purolite A860 resin median diameter to be varying from

610 µm for virgin resin down to 519 µm after one year of operation without regeneration (Amini

et al. 2018), which confirms our observation in this study. IX resin shrinks first depending on the

ionic strength in the water as in our case where the influent water contained a high concentration

of sulfate due to the counter-osmotic pressure (Lodi et al. 2017, Sun et al. 2016). Moreover, we

speculate that multivalent anions sorbed onto IX also play a role with regards to IX resin shrinking

and ammonia release (fig. S4.4). Initially charged with monovalent anions (chloride), virgin IX first

exhibited a uniform and loosely-packed morphology (fig. S4.4a). However, multivalent anions,

such as sulfate and some NOM molecules, would be exchanged onto IX due to their higher

affinities. The multivalent anions could then potentially form inter-chain or intra-chain bridges

within or between the IX polymers (fig. S4.4b). Ion bridging, which acts as the attraction force

between IX resin polymers, gradually leads to the IX resin shrinking and the formation of

agglomerates. Areas inside the agglomerates without flowing water (i.e., water channeling

phenomenon) would then turn into anoxic/anaerobic zones where ammonia is produced through

the ammonification of organic nitrogen while nitrification is restrained under anoxic condition.

118

However, the phenomenon of ion bridging induced by multivalent ions was only reported in

polyelectrolyte brush systems (Mei et al. 2006, Yu et al. 2017), a similar ion exchange system to

IX. Therefore, more studies are needed to validate this hypothesis.

Figure S4.4 - Schematic presentation of a) virgin IX with loosely-packed and uniform morphology

at the beginning of test; 2) inter-chain or intra-chain ion bridging induced by multivalent anions

after long period of operation;

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123

Chapitre 5 – Alleviating the burden of ion exchange brine in

water treatment: from operational strategies to brine

management

Abstract

Ion exchange (IX) using synthetic resins is a cost-efficient technology to cope with a wide range

of contaminants in water treatment. However, implementing IX processes is constrained by the

regeneration of IX resins that generates a highly concentrated brine (i.e., IX brine), the disposal

of which is costly and detrimental to ecosystems. In an effort to make the application of IX resins

more sustainable in water treatment, substantial research has been conducted on the

optimization of IX resins operation and the management of IX brine. The present review critically

evaluates the literature surrounding IX operational strategies and IX brine management which

can be used to limit the negative impacts arising from IX brine. To this end, we first analyzed the

physicochemical characteristics of brines from the regeneration of IX resins. Then, we critically

evaluated IX operational strategies that facilitate brine management, including resin selection,

contactor selection, operational modes, and regeneration strategies. Furthermore, we analyzed

IX brine management strategies, including brine reuse and brine disposal (without or with

treatment). Finally, a novel workflow for the IX water treatment plant design that integrates IX

operational strategies and IX brine management is proposed, thereby highlighting the areas that

make IX technology more sustainable for water treatment.

Keywords: Ion exchange resins; Water treatment; Waste management; Brine reuse; Brine

disposal; Sustainability

5.1 Introduction

Contaminants are removed by IX resins via two main mechanisms which are driven by the

functional groups of contaminants. Charged groups can be exchanged onto the IX functional

groups through electrostatic interactions whereas uncharged groups can be adsorbed onto resin

124

surface through Van der Waal force or hydrophobic interactions. For example, polyvalent metals,

such as calcium, magnesium, copper are mainly removed through ion exchange using cation

exchange (CIX) resins (Edzwald, 2010). While natural organic matter (NOM) and organic

micropollutants (e.g., polyperfluoroalkyl substances and microcystins) are removed through both

ion exchange and adsorption using anion exchange (AIX) resins (Liu et al., 2020; Dixit et al., 2021).

The use of IX resins for water treatment has numerous advantages over other treatment

technologies. First, the operation of IX resins is flexible in terms of application scale, allowing the

implantation of IX resins for both small-scale (e.g., household point-of-use) and large-scale (e.g.,

water treatment plants (WTP)) applications. Second, IX resins are simple to operate; they are

usually operated in fixed-bed contactors with a short empty bed contact time (EBCT) (e.g., 2-8

min). Third, some IX resins are reusable and can be regenerated on-site using typically a highly

concentrated NaCl brine (e.g., 8%-12% w/v) to recover the IX capacity (Edzwald, 2010). However,

as the majority of the regenerant ends up in the brine, the ion exchange brine (i.e., IX brine) is

composed of a high concentration of NaCl. The disposal of such brine is costly, and its direct

discharge to the environment adversely impacts ecosystems (Rokicki and Boyer, 2011; Ness and

Boyer, 2017). Consequently, the management of IX brine has become a bottleneck in the

application of IX resins for water treatment, with some WTP abandoning IX processes due to the

lack of a proper IX brine management scheme (Levchuk et al., 2018).

In contrast to other separation processes (e.g., membrane and distillation), where residual

management has been reviewed in detail in the last decade (Mezher et al., 2011; Pramanik et al.,

2017; Panagopoulos et al., 2019), no similar effort has been made to review IX processes, even

though a large number of studies have been devoted to overcoming barriers in IX brine

management. Therefore, the main objective of this review is to critically evaluate IX operational

strategies and IX brine management that can be used to limit the negative impacts imposed by IX

brine. To this end, first, we summarize the physicochemical characteristics and chemical

composition of IX brine generated from both CIX and AIX processes. Second, we assess IX

operational strategies and IX brine management reported in scientific literature. Finally, we

propose a novel workflow for an IX water treatment plant design that integrates IX operational

strategies and IX brine management.

125

5.2 Ion exchange brine characteristics

Table 5.1 summarizes the reported physicochemical characteristics and chemical composition of

IX brine following a single regeneration (i.e., without multiple brine reuse). Generally, IX brine

following a single regeneration has a neutral or slightly basic pH of 7-8 and a high electrical

conductivity (~100 mS/cm) because of the use of concentrated fresh brine (e.g., 35-150 g NaCl/L).

The chemical composition of IX brine varies widely owing to the diversity of factors coming into

play. First, the chemical composition of IX brine depends on the target contaminants removed

during the treatment cycle as well as the IX operational conditions (e.g., IX filter run time, type of

fresh brine, duration of rinsing cycle after regeneration). A longer run time accumulates more

contaminants, leading to a higher concentration in the IX brine. On the other hand, a longer

rinsing cycle post-regeneration can dilute the IX brine. Second, the physicochemical

characteristics of the feed water, as well as the upstream pretreatment, are expected to impact

the chemical composition of the brine. During the treatment cycle, both the target contaminants

and competing ions were removed. Accordingly, the IX brine is a mixture of brine, target

contaminants, competing ions, and other background ions present in the make-up water. Overall,

given the confounding factors contributing to the chemical composition, IX brine characteristics

should be studied on a case-by-case basis.

In addition to IX brine from a single regeneration, a number of studies have documented IX brine

characteristics following multiple brine reuses. As expected, contaminants and non-target ions

(e.g., sulfate and bicarbonate) accumulated as the number of reuse cycles increased. For example,

sulfate, bicarbonate, and nitrate were found to accumulate in the IX brine while using anion

exchange resins to remove nitrate (Bea et al., 2002) and arsenic (Clifford et al., 2003). However,

the concentrations of non-target ions tended to stabilize if no treatment was conducted during

reuse. Lehman et al. (2008) noted that sulfate and bicarbonate accumulated in the brine of AIX

(used to remove nitrate and perchlorate) until they reached stable concentrations after 13-15

cycles. Similarly, Duan et al. (2020) found that the loading of bicarbonate on the AIX resin (used

to remove nitrate) reached a stable condition after 10 cycles. Although Duan et al. (2020)

reported that no significant impact on the IX performance was observed for sulfate accumulation

in the brine, one should note that their conclusion was based on a short operation period (96 BV).

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The accumulation of non-target anions, such as sulfate, can deteriorate the IX treatment

efficiency (especially for long-term operation) or shorten the service time of IX resins as

mentioned by Liu and Clifford (1996). Therefore, it is often recommended to apply treatment

strategies for IX brine prior to multiple reuses, a topic that will be discussed later in the present

review.

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Table 5.1 - Physicochemical characteristics of ion exchange brine following a single regeneration

Targets Fresh brine

Composition of brines

References Ca2+ Mg2+ Na+ Cr

Total As SO42- Cl- NO3

- HCO3- ClO4

- TOC pH Cond.

g/L g/L g/L mg/L mg/L g/L g/L g/L g/L mg/L mg/L - mS/cm

Hardness 69 g/L KCl 0.91 0.20 - - - - - - - - - - Birnhack et al. (2019)

Hardness 150 g/L NaCl 15.3 0.30 11 - - 0.2 132 - - - - - - Cob et al. (2014)

Hardness - 51.2 7.70 24 - - 0.0 149 - - - - - - Cob et al. (2014)

Hardness - 2.03 1.38 2 - - 0.1 35 - - - - - - Gryta et al. (2005)

Cr 117 g/L NaCl - - - 34 0.20 18.0 49 - - - - - - Plummer et al. (2018)

Cr 117 g/L NaCl - - - 71 0.20 - 52 0.4 - - 289 - - Seidel et al. (2014)

Cr 117 g/L NaCl - - - 62 - - 52 - - - - - - Seidel et al. (2014)

Cr 117 g/L NaCl - - - 80 - - 47 - - - - - - Seidel et al. (2014)

Cr 117 g/L NaCl - - 41 29 0.12 16.3 50 2.3 2.9 - - 8.3 123 Korak et al. (2018)



Cr - - - 23 320 3.60 35.0 15 0.7 2.9 - - 9.2 98 Arias-Paic and Korak (2020)

Cr - - - - 10-100 - 4.8-

48 - - - - - - - Pakzadeh and Batista (2011a)

Cr - - - 44 - 0.01 27.9 49 0.7 1.2 - - - - Homan et al. (2018)

Cr NaCl - - 29 17 0.07 11.3 37 1.5 1.9 - - - - Homan et al. (2018)



Cr NaHCO3 - - 23 12 0.01 0.2 5 0.7 54.1 - - - - Homan et al. (2018)

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Table 5.1 - Physicochemical characteristics of ion exchange brine following a single regeneration (continued)

Targets Fresh brine

Composition of brines References

Ca2+ Mg2+ Na+ Cr Total As SO4

2- Cl- NO3- HCO3

- ClO4- TOC pH Cond.

mg/L mg/L g/L mg/L mg/L g/L g/L g/L g/L mg/L mg/L - mS/cm

Cr Na2SO4 - - 41 11 0.07 83.1 1 0.4 1.5 - - - - Homan et al. (2018)

Cr 117 g/L NaCl - - 48 364 - - 71 4.6 - - - - - Siegel and Clifford (1988)

Cr 59 g/L NaCl - - 25 100 - - 36 4.6 - - - - - Siegel and Clifford (1988)

As - - - 16-47 - 5-

120 4.8-48 24-73 - - - - 8-

10 - Pakzadeh and Batista (2011b)

As - - - - 300 0.6 24 - 0.3 - - - - An et al. (2005)

As - - - 24 - 300 0.6 36 - 0.3 - - - - An et al. (2011)

NO3- - - - 29 - - 1.7-2.1

8.4-11.2

2.7-3.7

5.8-9.1 - -

8.0-9.2

- Clifford and Liu (1993)

NO3- - 39 19 - - 2.6 12 2.1 0.1 - - 6.8 45 McAdam and Judd (2008a)

NO3- - 11 3 45 - - 1.2 70 6.5 1.8 - - - - Bergquist et al. (2016)

NO3- - - - - - 1.4 12 2.9 - - - - - Trogl et al. (2011)

NO3- - 512 367 46 - - 2.4 75 9.4 - - - 7.0 - Hutchison et Zilles (2018)

NO3- 60 g/L NaCl 27 21 17 - - 6.0 25 1.7 - - 24 7.8 - Yang et al. (2013)

NO3- NaHCO3 - - 23 - - 0.3 0 1.0 61.0 - - - - Paidar et al. (2004)

NO3- NaHCO3 - - - - - 1.3 0 0.9 - - 8.5 - Paidar et al. (2004)

NO3- - 172 - 42 - - 4.4 51 8.8 0.5 - - 7.4 108 McAdam and Judd (2008b)

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Table 5.1 - Physicochemical characteristics of ion exchange brine following a single regeneration (continued)

Targets Fresh brine

Composition of brines

References Ca2+ Mg2+ Na+ Cr Total As SO4

2- Cl- NO3- HCO3

- ClO4- TOC pH Cond.

mg/L mg/L g/L mg/L mg/L g/L g/L g/L g/L mg/L mg/L - mS/cm

NO3- - - - 40 - - 9.0 46 10.0 11.0 - - 7.7 113 Arias-Paic and Korak (2020)

ClO4- 70 g/L NaCl 40 1 - - - 4.7 32 2.4 8.5 2.0 - 7.4 - Liu et al. (2013)

ClO4- - 22 3 - - - 1.6 1.6 - 3.5 - - 92 Hiremath et al. (2006)

ClO4- - - - 53-100 - - 0.6-6 0.4-

4.0 11.0 4.3 5-10 Hiremath et al. (2006)

COD 35 g/L NaCl - - 15 - - 7.5 13 0.3 7.1 - - - - Vaudevire et al. (2019)

Colorant - - - 20 - - 30 - - - 5000 - - Wadley et al. (1995)

NO3-/Hard. - 52 - - - - 5.7 43 15.1 1.2 - 7.4 125 McAdam and Judd (2008b)

*Cond.: Conductivity (mS/cm); TOC: total organic carbon; -: data not available.

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5.3 Ion exchange operational strategies to facilitate brine management

5.3.1 Resin selection strategies

While the selection of resins first depends on the target contaminants and feed water

characteristics (e.g., pH and background ions), the properties of selected resins can impact brine

management for a given application. For example, strong basic anion exchange (SBA) resins are

usually regenerated with concentrated sodium chloride solution whereas weak basic anion

exchange (WBA) resins are mostly regenerated with caustic soda solution. The brine management

strategies for these two resins can be significantly different given the different brine corrosivity.

The resin backbone materials can also impact the regeneration strategy. For example, as

polystyrene resins are more hydrophobic than polyacrylic ones (Levchuk et al., 2018), they may

favor the uptake of a given contaminant but negatively impact regeneration efficacy. Therefore,

resin properties should be taken into account during the selection of resins to anticipate the best

brine management strategy. Generally, bench- or pilot-scale tests are often conducted to identify

the most efficient resin for a given application. Some commercial software can offer the

comparison of different resins, such as LewaPlusÒ (Lanxess), WAVE (Dupont), Resinex Design

Software (Jacobi) and PRSM™ (Purolite). The cost associated with IX brine management can be

significantly reduced by maximizing treatment efficiency and reducing regeneration frequency.

Therefore, selecting the most efficient IX resin for a given treatment objective can be the most

effective solution to reduce the management cost of IX brine (Plummer et al., 2018).

Furthermore, modifying the resin structure to reduce brine production is also of high interest and

has received attention from IX resin suppliers and researchers. For example, the Shallow Shell™

Technology (SST®), patented by Purolite, is an alternative CIX for conventional IX resins used in

water softening and demineralization. Briefly, unlike conventional strong acidic cation exchange

(SAC) resins, the SST® resin is produced with an inert center eliminating the IX sites that are slower

and more difficult to reach, more difficult to regenerate, and more likely to suffer from pore

plugging (Purolite, 2014). By doing so, the supplier states that regeneration of SST® resins requires

15% less salt and 50% less water for rinse and dilution (Purolite 2014). The city of Chilton

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(Wisconsin, USA) renewed its softeners with SST® resin and found that the salt usage decreased

by approximately 32% and that the water consumption used for a regeneration cycle decreased

by 16% compared to conventional resins (WTP production capacity: 2600 m3/day) (Olsen, 2015).

Nonetheless, Olsen (2015) also noted that regeneration became more frequent because of the

reduced resin capacity, with an average of 2.9 times per day using SST® resins vs. 2.5 times per

day using conventional resins (an increase of 16%). More frequent regeneration may have an

impact on equipment costs as more softeners must be provisioned if they are more often off-

duty. In 2016, Arar (2016) reported that the performance was similar between conventional IX

resins and SST® resins for the removal of copper, but the SST® resins demonstrated a higher

regeneration efficiency than conventional resins (100% vs. 80% using 2 M H2SO4). Site-specific

validation is still required to compare the overall costs of operating SST® resins and conventional

resins for a given application.

The application of single-use resins is on the rise in the water treatment sector (especially for

PFAS and perchlorate removal) thanks to the simplified operation and maintenance and the

current limitations regarding the disposal of contaminated brine (Choe et al., 2013; Dixit et al.,

2020). The exhausted single-use resins are usually incinerated off-site, which eliminates the

generation of IX brine. A life cycle assessment study comparing the environmental impact of

regenerable IX vs single-use IX for perchlorate removal shows that the single-use resin is more

environmentally sustainable and more economic than regenerable IX (Choe et al., 2013).

Additionally, the authors also demonstrated that, for regenerable IX systems, the dominant

contribution to environmental impact comes from the salt used for regeneration (> 80%), and

strategies to treat and recycle the brine significantly reduce the environmental impact of

regenerable IX systems (Choe et al., 2013; Choe et al., 2015). Furthermore, given the fact that

resin manufacturing has been identified as an important contributor to the overall environmental

impact of IX systems according to life cycle assessment (Choe et al., 2013; Amini et al., 2015),

resin reuse should always be encouraged as long as the brine management scheme is more

sustainable than resin replacement. A thorough life cycle assessment needs to be conducted on

a case-by-case basis to compare the environmental impact of single-use resins and regenerable

resins given that the outcome is impacted by several input factors.

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5.3.2 Resin contactor configuration

Fixed-bed contactors are by far the most common configuration for IX resins due to the simple

operation. Two variants of implementation are used in the industry: 1) Standard dual train

configuration where one IX contactor is on-line while a second contactor is standby or undergoing

regeneration; 2) Lead-lag configuration in which feed water passes through the first contactor

(i.e., lead contactor) and then the second contactor (i.e., lag contactor). The lead contactor is

taken offline for regeneration when the breakthrough occurs in the effluent. Following

regeneration, the lead-lag reactor configuration is inverted: the regenerated former lead

contactor returns to the service at the lag position and the former lag contactor becomes the new

lead contactor (Hosea et al., 1988). The lead-lag configuration can not only offer a higher removal

than standard dual train configuration but also maximize the IX operation capacity prior to

regeneration, an advantage that can benefit brine management by reducing regeneration

frequency.

Moving-bed contactors, on the other hand, offers a configuration where the resin is either

fluidized or moving inside the reactor. Such configurations have attracted attention over the last

decade due to their ability to remove NOM from turbid raw waters. Examples of these include

magnetic ion exchange (MIEX) (Slunjski et al., 2000), fluidized ion exchange (FIX) (Cornelissen et

al., 2009) and suspended ion exchange (SIX) (Metcalfe et al., 2015). A life cycle assessment study

comparing fixed-bed IX contactors and the MIEX process suggested that fixed-bed IX contactors

consume more salt and produce more brine than the MIEX process (Amini et al., 2015).

Nonetheless, the authors attributed this observation to the improper maintenance of resins for

the fixed-bed contactors. Further investigations are still required to systematically compare fixed-

bed contactors and moving-bed contactors in terms of brine management.

5.3.3 Novel ion exchange operational mode

A novel operational mode for AIX resins for NOM removal was first proposed by Schulz et al.

(2017). This strategy, referred to as biological ion exchange (BIEX), involves operating AIX resins

with infrequent regeneration to promote the microbial colonization of resins, thereby prolonging

IX service time through biodegradation (Schulz et al., 2017). In a pilot-scale study where the BIEX

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filter had not been regenerated for 330 days, the BIEX filer achieved slightly lower dissolved

organic carbon (DOC) removal compared to its weekly regenerated counterpart (62% vs. 80%)

(Amini et al., 2018). NOM removal in the BIEX filter was attributed to a combination of ion

exchange, especially secondary ion exchange, where NOM exchanges with pre-retained sulfate

and biodegradation (Liu et al., 2020). Mass balance studies revealed that at most, 30% of NOM

removal was attributed to the biodegradation (Amini et al., 2018; Liu et al., 2020). Although

running IX resins in BIEX mode could significantly reduce brine production, one should note that

NOM removal is lower than the conventional mode (i.e., with regular regeneration). Therefore,

BIEX is only applicable when the reduced NOM removal performance still meets the treatment

objective. The regeneration frequency for a BIEX filter is site-specific and depends on the feed

water characteristics. For example, in a study by Amini et al. (2018), water treated with BIEX was

able to comply with the Québec trihalomethane (THM) standard (80 µg/L) for one month whereas

another pilot test carried out in the Middle River (BC, Canada) was able to operate with BIEX for

1 year without regeneration while meeting the THM standard (Zimmermann et al., 2020). Overall,

source waters with low sulfate concentrations (e.g., 3-5 mg/L) are ideal for the BIEX process.

5.3.4 Novel ion exchange regeneration strategies

5.3.4.1 Segmented regeneration

Segmented regeneration, patented by Ionex SG, is an alternative to conventional regeneration

aimed at eluting the contaminant loaded on IX resins in the smallest possible volume of IX brine.

This technique has been mainly applied to AIX resins for the removal of chromium and nitrate

(Plummer et al., 2018). Briefly, the segmented regeneration consists of 1) using a regenerant of

low concentration (e.g., 0.2 M NaCl) to first elute secondary ions (e.g., sulfate and bicarbonate)

and 2) using a regenerant at high concentrations (e.g., 2 M NaCl) to elute chromium or nitrate

from the IX resin (Waite, 2018). The low-concentration IX brine, that is, the brine containing non-

target ions, can be reused or even recycled to the head of the treatment train (Korak et al., 2017).

Regarding the high-concentration regeneration, a chromatographic peak of the contaminant is

eluted and therefore concentrated in a smaller volume of brine. Meanwhile, high-concentration

IX brine that contains little contaminant can be reused for the next cycles along with the rinsing

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water (Plummer al., 2018). This strategy has already been applied to full-scale AIX systems in

California, and the volume of IX brine to be discharged was reduced by > 80% compared to a

conventional regeneration method (Plummer et al., 2018). However, prior to full-scale

application, this approach requires laboratory pretests on a case-by-case basis to investigate the

elution profile achieved during regeneration.

5.3.4.2 Alternative regenerants to NaCl

NaCl is by far the most common solution for regenerating IX resins because of the affordable and

chemically inert features of NaCl. However, given the stricter jurisdictions on wastewater

discharge and concerns over ecosystem impairment, many studies have pursued greener

alternatives to NaCl. It is worth noting that the trade-off between IX operation performance and

IX regeneration performance should be fully considered when selecting alternative regenerant. A

more selective regenerant (compared to NaCl) could offer a better contaminant elution during IX

regeneration, however, it could also lower the IX operation performance in the following cycles

due to the higher affinity to resins. Relative resin selectivity for ions can be found in de Dardel

and Arden (2008).

5.3.4.2.1 Alternative regenerants for cation exchange resins

Potassium salts (e.g., KCl) have been suggested for the regeneration of softeners as an alternative

to NaCl to regenerate CIX resins because of their 1) lower harmful impact on the environment; 2)

favorable increase of potassium in the human diet as opposed to sodium; 3) value for irrigation

and soil properties; and 4) excellent solubility in water (254 g KCl/L at 20 °C) (Birnack et al. 2019).

In 2014, Maul et al. (2014) evaluated the regeneration efficiency and the environmental impacts

of IX brine for four types of salt (NaCl, KCl, NaHCO3, KHCO3). They concluded that potassium salts

(KCl and KHCO3) have a higher efficiency than sodium salts for CIX regeneration, and the use of

KCl brine for irrigation could offset the demand for potassium-based fertilizers. However, the use

of KCl salt may increase the environmental burden compared to NaCl salt based on the study of

the life cycle assessment (Maul et al., 2014). In addition, the higher cost of KCl salts (3-5 times

more expensive than NaCl) still hinders its application, especially for large water utilities (Birnack

et al., 2019).

135

In 2016, Li et al. (2016a) proposed a softening process using CIX resins in aluminum (Al3+)-form

rather than sodium (Na+). Given that the precipitation of aluminum hydroxide occurs quickly in

water, using aluminum-form IX resins simultaneously removes hardness and alkalinity. The

mechanisms are expressed by equations (5.1-5.3).

2(# − %&!")!()!#*********************** + 3-.$# → 3(# − %&!")$-.$#*********************** + 2()!# (5.1)

2()!# + 61$& → 2()(&1)!(2) ↓ +61# (5.2)

61# + 61-&!" → 61$& + 6-&$(4) ↑ (5.3)

Although the use of aluminum-form resins for softening is counterintuitive because of the

trivalence of aluminum, Li et al. (2016) demonstrated that aluminum-form resins removed

hardness to a 20% lower effluent concentration than the resin in sodium form. In addition, unlike

regeneration with NaCl, where a high concentration must be used, a lower concentration (3% w/v

AlCl3) is sufficient for regeneration (quantity of Al required for regenerant: hardness to be

removed ≈ 1:1), demonstrating its advantage in IX brine management. Nevertheless, aluminum

hydroxide precipitation was observed on the surface of the resins both after exhaustion and after

regeneration with AlCl3, a phenomenon that may lower the resin performance for subsequent

cycles. Therefore, the viability of this solution for long-term operation has yet to be

demonstrated.

5.3.4.2.2 Alternative regenerants for anion exchange resins

The most studied alternative for AIX resins is sodium bicarbonate (NaHCO3) for two reasons: 1)

bicarbonate has a similar affinity to chloride on AIX resins (Rokicki and Boyer, 2011); and 2)

bicarbonate is less environmentally detrimental than chloride and can be discharged to the

environment or into the sewer (Jelinek et al., 2004). The performance of bicarbonate-form AIX

resins has been investigated for the removal of NOM in a lab-scale study (Walker and Boyer,

2011). The authors demonstrated that bicarbonate-form AIX had the same performance as

chloride-form AIX for NOM removal at the beginning of the test. However, because of the lower

regeneration efficiency of the bicarbonate solution, DOC removal by bicarbonate-form resins was

7%-18% lower than their chloride-form counterparts after 21 service cycles. Moreover,

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bicarbonate-form AIX resins have also been evaluated for the removal of chromium (VI). Matosic

et al. (2000) reported that bicarbonate-form resins achieved a slightly lower removal of chromium

(VI) than their chloride-form counterparts, whereas Li et al. (2016b) reported a similar

performance between bicarbonate- and chloride-form resins for the same application propose.

Overall, the lower solubility of NaHCO3 in water (approximately 100 g/L vs. 360 g/L for NaCl at 25

°C) would limit the maximum concentration of regenerant that can be used, and the cost of

NaHCO3 salt is approximately 3 times more expensive than NaCl salt (based on the 2020 prices in

Montreal, Canada for an NSF-certified product). Therefore, the application of bicarbonate salt as

an alternative regenerant for AIX resins remains limited.

Two alternative regenerants have been proposed for the recovery of perchlorate from AIX resins.

Given that the high affinity of perchlorate for AIX resins makes it difficult to use NaCl as a

regenerant, Gu et al. (2001) proposed using tetrachloroferrate (FeCl4-) as an alternative and

achieved approximately 100% recovery with 5 L regenerant/L resin. Furthermore, Gutiettre et al.

(2008) suggested using iodide as an alternative to recover perchlorate from exhausted resins.

However, the management of such brine is difficult, not to mention the high cost of iodide salts

and the absence of NSF-certified products.

5.3.4.3 No-chemical-addition regeneration

Discovering the chemical-free regeneration technique has been a goal of IX research since the

initial discovery of this technology. Chemical-free regeneration eliminates the transport of

chemicals (i.e., salts). These methods can be divided into three groups based on their

mechanisms: biological, electrochemical, and thermal regeneration, which will be reviewed in

detail in the following sections.

5.3.4.3.1 Biological regeneration

With a biological approach, the regeneration of exhausted resins and the destruction of the

contaminant loaded on the resins occur simultaneously by directly contacting the spent resins

with an active biomass. To date, this method has been mostly evaluated for perchlorate- and

nitrate-laden IX resins, given that the target contaminant must be biodegradable. Wang et al.

(2008) were the first to present this idea with the aim of regenerating a resin loaded with

137

perchlorate and repeatedly regenerated resins loaded with perchlorate or nitrate. The

mechanism of biological regeneration consists of desorbing the perchlorate or nitrate from the

resin prior to biodegrading perchlorate to chloride and nitrate to N2 with an acclimated biomass

(Xiao et al., 2010; Ebrahimi and Roberts, 2013; Ebrahimi and Roberts, 2015; Sharbatmaleki et al.,

2015). Although biological regeneration is technically viable, the long regeneration time and the

need for washing procedures to remove biomass from the resin surface limit its full-scale

application (Venkatesen et al., 2010).

5.3.4.3.2 Electrochemical regeneration

Some studies have resorted to using electrodialysis with bipolar membranes (EDBM) to produce

acid or base solutions for the on-site regeneration of IX resins. This method was successfully

tested for CIX and AIX resins in full-scale applications. For example, Bolton (1992) investigated the

on-site generation of acids and bases via EDBM and then successfully applied the generated acid

and base into full-scale CIX and AIX contactors with efficiency similar to that of conventional

chemical regeneration. Similarly, Chen et al. (2016) used EDBM to generate HCl and NaOH from

a NaCl solution (100 mM), and then used the acid to regenerate weak acid cation (WAC) softeners

and the base was used to promote the crystallization of CaCO3 and Mg(OH)2. Consequently, Chen

et al., (2016) reported that the energy required to produce the acid and the base was an order of

magnitude lower than that required to purchase the chemical. Although chemicals (e.g., 100 mM

NaCl) are still needed in the process of EDBM, their usage can be significantly reduced compared

to conventional IX regeneration with a high concentration of NaCl. In addition, the on-site

production of regenerants could offer an advantage for remote communities, which would

otherwise depend on long-distance transportation of chemicals (Bolton, 1992). However, an

economic analysis comparing this method to conventional chemical regeneration is still needed

as the capital expenditure for implementing EDBM may prohibit the viability of this strategy.

Furthermore, Xing et al. (2007) evaluated electrodeionization (EDI) for in situ regeneration of

resins loaded with chromium (VI) and recovered it as chromic acid. Briefly, AIX resins first adsorb

chromium (VI) during the treatment cycle and after the saturation of chromium (VI) on the resin

phase, the electricity supply is triggered in the IX chamber to regenerate resins through water

138

electrolysis (with OH-). In their work, the regeneration cycle was maintained for 24 h using a

current of 0.25 A. Approximately 93% of the IX capacity was recovered, an efficiency similar to

chemical regeneration. Later, Su et al. (2013, 2014) developed a novel process based on the

principle of EDI but without the use of ion exchange membranes and referred to it as membrane-

free electrodeionization. This process produces high-purity water and consists of (i) a service cycle

where mixed-bed resins (i.e., CIX and AIX) remove impurities from water and (ii) a regeneration

cycle where resins are regenerated by H+ and OH- generated through water electrolysis. However,

membrane-free electrodeionization has mainly been studied for the production of high purity

water and has been only tested with synthetic water of high quality (10-70 μS/cm in conductivity,

e.g., osmosis permeate) (Shen et al., 2014; Hu et al., 2015; Hu et al., 2016). Therefore, the

feasibility and stability of this process for municipal and industrial feed remain uncertain.

5.3.4.3.3 Thermal regeneration

With thermal methods, resins are regenerated at high temperatures using hot water

(approximately 80 °C) (Bolto et al., 1978). This process was first applied to a mixed bed of weak

base anion exchange (WBA) and weak acid cation (WAC) exchange resins (Battaerd et al., 1973).

Later, a novel resin material that simultaneously incorporated the functional groups of WBA and

WAC resins (i.e., Sirotherm resin) was developed (Bolto and Jackson, 1983). The Sirotherm resin

is thermally regenerable, which means that the resin is capable of eliminating ions in an aqueous

solution at room temperature and then desorbing ions at high temperatures due to unfavorable

thermodynamics (Bolto and Jackson, 1983). However, this process failed during trials at the

industrial scale and research on this subject was suspended (CSIROpedia, 2014). In 2009, a new

thermally regenerable resin crosslinked with polyacrylic acid and ethoxylated polyethyleneimine

was synthesized (Chanda et al., 2009). This resin was successfully tested for 10 cycles for the

desalination of brackish water (Chanda et al., 2010). However, the results showed that the IX

capacity decreased by a factor of 8 when the temperature increased from 30 °C to 80 °C (Chanda

et al., 2010), suggesting that the resin can only be partially regenerated. To improve the efficiency

of thermal regeneration, Chandrasekara and Pashley (2015) used ammonium bicarbonate to

regenerate the Sirotherm resin and demonstrated the feasibility of thermally recycling

ammonium bicarbonate during regeneration. Briefly, exhausted resins were soaked in

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ammonium bicarbonate solution (2 M) at 20 °C for 15 h, and then the resins were heated to 80

°C for 1 h to decompose ammonium bicarbonate while returning the resins to their initial form

(i.e., tertiary amine and carboxylic acid). Carbon dioxide and ammonia gases were recovered and

used for subsequent regeneration. However, this process is still in the development phase, and a

full-scale demonstration is needed to prove its feasibility for carbon dioxide and ammonia gas

recovery. Overall, thermal regeneration requires further development.

5.4 Ion exchange brine management

5.4.1 Ion exchange brine reuse

IX brine can potentially be reused as a regenerant because of its high salt content. Reusing brine

for IX regeneration allows for the direct reduction of salt usage and water consumption (Verdickt

et al., 2011; Flodman and Dvorak, 2012). Therefore, given the great economic and environmental

benefits of reuse, one should always consider IX brine reuse when designing the IX brine

management scheme. However, it is also worth noting that a loss of treatment performance may

occur if the recycled regenerant is of low quality. A reduction in treatment efficiency may lead to

an increase in the regeneration frequency, which could completely or partially offset the benefits

of reuse. Salt makeup and/or pH adjustment may be required before the direct reuse of the IX

brine. In cases where direct reuse is not feasible, pretreatment strategies may have to be

considered prior to brine reuse. In this section, we summarize past studies on direct IX brine reuse

(with or without NaCl addition) and treatment strategies for the reuse of IX brine.

5.4.1.1 Direct reuse

The IX brine can sometimes be reused directly without any makeup with fresh salts. Clifford et al.

(2003) investigated the direct reuse of NaCl brine (117 g/L) for the regeneration of arsenic-laden

AIX resins, and the IX brine was successfully reused for six times without any efficiency decline.

Similarly, Medina et al. (2018) evaluated the direct reuse of IX brine for the regeneration of NOM-

laden AIX resins with a saturated regenerant (360 g/L NaCl), and successfully reused the brine for

three times with little impact on IX performance. Duan et al. (2020) reused the IX brine to

regenerate nitrate-laden AIX resins for 23 times, however they reported that the nitrate removal

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efficiency gradually decreased from 100% to 38%. Overall, direct reuse of IX brine without any

chemical makeup is feasible, but the number of reuses needs to be defined on a case-by-case

basis.

Make-up with salt (e.g., NaCl) and/or pH adjustment (e.g., with NaOH) are common practices to

improve regeneration efficiency during the direct reuse of brine. Clifford et al. (2003) successfully

reused IX brine to regenerate arsenic-laden AIX resins for 14 cycles with salt compensation of 58.5

g NaCl/L before each reuse. For the same application purpose, An et al. (2005) achieved 7 cycles

of reuse with stable arsenic recovery (> 95 %) with pH adjustment to 9.2-10 before each reuse.

Kim and Symons (1991) reported that the total organic carbon (TOC) in the treated effluent

remained stable for nine operation cycles when reusing IX brine with readjustment each time to

2 M NaCl and 1.5 M NaOH. However, during direct brine reuse to regenerate AIX resins charged

with chromium, Plummer et al. (2018) noted that the resins were not fully regenerated in the

second reuse cycle, even though they compensated the brine to 117 g NaCl/L. Foldman and

Dvorak (2012) investigated the reuse of IX brine for softener regeneration with compensation of

saturated NaCl solution (360 g/L) and concluded that the benefits of reuse came at the cost of

reduced performance for the system. Accordingly, the feasibility of reuse must be investigated

on a case-by-case basis, as its success depends on initial regenerant concentration, operational

mode, raw water characteristics, and target contaminants.

5.4.1.2 Treatment strategies for reuse

In cases where direct reuse of IX brine is not feasible, some studies have sought to verify whether

treatment strategies would allow the reuse of IX brine. The objectives of brine treatment are to:

1) remove impurities (e.g., contaminants or non-target ions) from IX brine to improve brine reuse

efficacy; or 2) recover valuable resources (e.g., NaCl, CaCl2 and MgCl2) from IX brine for reuse or

resell. Substantial research has been devoted to the development of eco-efficient treatment

strategies for IX brine. These strategies are grouped and critically reviewed in the following

subsections based on the target contaminants.

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5.4.1.2.1 Treatment strategies for hardness-laden brine

Taking advantage of reverse osmosis (RO) brine. To remove hardness from CIX brine, Cob et al.

(2014) combined RO brine (loaded with sodium and bicarbonate) with hardness-laden IX brine.

The authors demonstrated that more than 99% of CaCO3 was precipitated with a pH adjustment

to 11, and the remaining IX brine (loaded with sodium chloride) could be reused for the

regeneration of CIX resins. Furthermore, Vanoppen et al. (2016) demonstrated the possibility of

directly using an RO brine to regenerate a CIX filter with minor addition of NaCl (2.5 g/L). However,

despite being economical, this method is only suitable for situations where CIX and RO systems

simultaneously exist in the treatment process, such as at a seawater desalination plant where CIX

is implemented prior to RO to remove hardness.

Resource recovery by hybrid processes. Numerous treatment strategies have sought to recover

valuable resources from hardness-laden IX brine. Birnhack et al. (2019) successfully applied a

hybrid process consisting of nanofiltration (NF), dia-NF, and RO to recover potassium salt from a

KCl brine. Gryta et al. (2005) investigated the use of membrane distillation (MD) to recover NaCl

from brine, however they found that MD performance was deteriorated due to fouling (e.g., Ca,

Mg, and Si) and the separation of NaCl from the salt mixture was a challenge. Micari et al., (2019a)

successfully extracted a pure NaCl solution from the softener brine using a treatment train of NF,

crystallization, and multi-effect distillation (MED). A techno-economic assessment was conducted

for the treatment train (Micari et al., 2019a), as well as the MED process (Micari et al., 2019b).

The authors reported that: 1) the lowest treatment cost of 4.9 USD/m3 brine could be achieved

with an NF recovery rate of 25%; 2) the evaporator is the most important contributor to the

process operating cost (30% of total cost); and 3) plane multi-effect distillation (MED without the

use of a thermo-compressor) supplied with recycled heat at a pressure of 1 bar was the most

economical way to operate the evaporator. Furthermore, eutectic crystallization has been

proposed as an alternative to evaporative crystallization for salt recovery from IX brine because

it consumes 6-7 times less energy than evaporative crystallization (Van der Ham et al., 1998, Lewis

et al., 2010, Fernandez-Torres et al., 2012). In a laboratory-scale study, Cob et al. (2014)

investigated the use of eutectic crystallization to treat softener brine. The authors were able to

obtain ice and NaCl crystals by lowering the temperature to -29.4 °C. Additionally, the authors

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concluded that a powerful scraping system was needed to avoid scaling caused by ice. More

recently, a feasibility study was conducted by Ahmad et al. (2017) to investigate the potential of

eutectic crystallization for the treatment of softener brine. The authors reported that the

presence of impurities (e.g., Ca and Mg) lowered the eutectic temperature for NaCl crystallization

but reduced scaling in the reactor, which improved the efficiency of energy transfer. From an

economic point of view, they concluded that eutectic crystallization led to a similar cost compared

to evaporative crystallization due to its high capitalization costs.

5.4.1.2.2 Treatment strategies for arsenic-laden brine

Coagulation and adsorption. Coagulation is a viable approach for removing arsenic from IX brine.

Clifford et al. (2003) successfully used FeCl3 to remove arsenic (As) from IX brine, and the authors

reported that arsenic removal depends on the initial arsenic concentration, pH of the brine, and

Fe/As ratio applied. An et al. (2005) also used FeCl3 to remove arsenic from IX brine, and the

treated brine was reused for regeneration without any decline in the performance of IX. Pakzadeh

and Batista (2011b) simulated the arsenic removal process in IX brine using a surface

complexation model and found that alkalinity and ionic strength are expected to affect the

efficiency of FeCl3. In addition to coagulation, An et al. (2011) developed starch-bridged magnetite

nanoparticles to adsorb arsenic from the brine. The authors concluded that arsenic was fully

eliminated under optimal conditions (Fe/As ratio = 7.6, contact time = 1 h). Nevertheless, the

method was only demonstrated with a synthetic brine (300 mg/L As with 6% NaCl w/w); its

performance using a real IX brine and the cost of application at an industrial scale remains

unknown.

5.4.1.2.3 Treatment strategies for chrome-laden brine

Reduction-coagulation-filtration. Regarding the treatment of chromium (VI) loaded with IX brine,

a reduction-coagulation-filtration process is widely used with many reactants that have been

evaluated for such applications, including sulfite in acidic conditions (Siegel and Clifford, 1988),

ferrous iron (Siegel and Clifford, 1988; Li et al., 2016b; Homan et al., 2018; Plummer et al., 2018),

hydrazine (Siegel and Clifford, 1988), polysulfides (Pakzadeh and Batista, 2011a; Plummer et al.,

2018), bisulfite, and stannous sulfide (Plummer et al., 2018). Among them, ferrous iron has been

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shown to be the most cost-effective (Siegel and Clifford, 1988; Plummer et al., 2018), though it

does not work in a bicarbonate brine (Homan et al., 2018).

5.4.1.2.4 Treatment strategies for NOM-laden brine

Coagulation and adsorption. Coagulation has been proven to be an efficient method for

removing NOM from IX brine. For example, Verdickt et al. (2011) used FeCl3 and post-filtration to

purify the brine before reuse. Subsequent to the adjustment of conductivity and pH (i.e., adding

NaCl/NaOH), the recycled brine was successfully reused 14 times with little impact on the IX

performance. Similarly, Medina et al. (2018) investigated the use of alum as a coagulant for NOM

removal in the brine. Although the treatment enabled three reuse cycles, the authors observed

that the regeneration efficiency was lower than that of brine direct reuse because of the

introduction of sulfate ions (from alum), which are strong competitors to NOM on IX resins.

Instead of using coagulants, Zhang et al. (2019) proposed using powdered activated carbon (PAC,

5 g/L) to adsorb dissolved organic matter (DOM) in IX brine after every 6-7 service cycles. The

treatment allowed a DOM removal of approximately 80% in the brine, and the brine was

successfully reused for 50 service cycles. However, the PAC must be filtered out from the solution

prior to reusing the treated brine, which makes this option more complex. Overall, the practice

of coagulation or PAC adsorption is feasible for IX brine laden with NOM, but the requirement of

post-filtration, the concern over other inorganic ion accumulation in the brine (e.g., sulfate), and

the disposal of sludge (or saturated media) still need careful consideration.

Resource recovery by thermal and/or electrodialysis processes. Numerous studies have sought

to recover NOM and/or salt from IX brine for reuse. Vaudevire et al. (2013) evaluated the dynamic

vapor recompression (DVR) process to recover salts from NF-concentrated IX brine. The authors

reported the co-precipitation of NaCl, Na2CO3, and Na2SO4 when the concentration factor

exceeded 10. However, the authors also noted that it was impossible to separate NaCl from the

salt mixture. Kabsch-Korbutowicz et al. (2011) first demonstrated the feasibility of using

electrodialysis (ED) to recover NaCl from a synthetic IX brine, although a fraction of NOM was

shown to pass through the IX membrane and contaminate the recycled NaCl solution. Vaudevire

et al. (2019) demonstrated a two-step ED process to recover NOM and NaCl solutions: the first

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stage consisted of recovering NaCl using a selective monovalent AIX membrane, while the second

stage consisted of separating multivalent anions from the solution loaded with NOM by a

standard IX membrane. Haddad et al. (2019) compared the performance of conventional and

selective monovalent IX ED membranes. The results showed that selective monovalent IX

membranes could recover NaCl of high purity with an energy requirement of approximately 2

kWh/kg NaCl. In addition, the authors also tested the use of a pulsed electric field to intensify the

separation and reduce membrane fouling, but the benefit was only noted for the conventional IX

membrane. More recently, Haddad et al. (2021) combined ED with membrane distillation (MD),

and they demonstrated that MD was a viable option to concentrate the purified NaCl produced

by ED. Overall, ED is a promising technique for NOM-laden brine treatment, but its application is

currently hindered by the capitalization costs of the equipment and the fact that the treated brine

still requires a post-concentration step to raise the concentration of NaCl to the required level for

the regeneration (i.e., 8%-12%w/w).

Resource recovery by membranes processes. Numerous studies have sought to use membrane

processes to recover resources from NOM-laden IX brine. Wadley et al. (1995) successfully

applied NF to recover NaCl from an IX brine generated from a decolorization process. Kabsch-

Korbutowicz et al. (2011) evaluated six ultrafiltration membranes (UF) and two NF membranes to

recover NaCl from a NOM-laden brine, and the results demonstrated that tight UF (5 kDa) and NF

were capable of providing a permeate (i.e., NaCl solution) of satisfactory quality; however, NOM

fouling was an important issue because the high salinity imposed an extremely high operating

pressure (120 bar). More recently, Caltran et al. (2020) demonstrated that ceramic NF yielded

less fouling than polymeric NF during the separation of NOM from IX brine. However, the poor

rejection of sulfate, especially at high ionic strength, is problematic, given that sulfate may lower

the performance of AIX during the following service cycles. Overall, using membrane processes

to recover salt and NOM from IX brine is a promising process and deserves further study.

5.4.1.2.5 Treatment strategies for trace organic pollutants-laden brine

Electrochemical processes. Few studies have investigated the treatment strategies for PFAS in IX

brine. Schaefer et al. (2020) reported that PFAS-laden IX brine is amenable to electrochemical

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treatment using boron-doped diamond anodes, but the treatment performance depends on the

fresh regenerant type. Chloride regenerant inhibits the electrochemical treatment of PFAS, as

chloride will also be oxidized at the anodes to perchlorate. Singh et al. (2020) successfully

degraded PFAS-precursors and long- and short-chain PFASs from IX brine using a sequential

plasma reactor, indicating that plasma treatment is promising for PFAS treatment from IX brine.

Resource recovery. Organic solvents, such as methanol and ethanol, can be used to regenerate

PFAS-laden resins with or without the addition of inorganic salts (Dixit et al., 2021). The organic

solvent can be recovered by distillation and reused for the next regeneration cycle (Singh et al.,

2020). One should note that such a strategy is only applicable for industrial applications where

the presence of a solvent residue in treated water is not causing any health concern. Furthermore,

the feasibility and sustainability of this strategy also remain to be evaluated at full-scale.

5.4.1.2.6 Treatment strategies for nitrate- and perchlorate-laden brine

The treatment strategies for nitrate and perchlorate are similar, that is, the contaminants are

chemically or biologically reduced to reduced forms so that the brine can be reused for the

following cycles.

Catalytic reduction. Catalytic reduction is a technique that has received considerable attention

for this subject. Pintar et al. (2001) first developed a catalyst (Pd-Cu/γ-Al2O3) that successfully

treated a nitrate-loaded brine using H2 as an electron donor. In their subsequent study (Pintar

and Batista, 2006), they modified the process into a two-stage treatment where the first stage

was dedicated to the reduction of nitrate to nitrite with the help of Pd-Cu/γ-Al2O3 and the second

stage reduced nitrite to nitrogen gas using the monometallic catalyst Pd/γ-Al2O3. This process

reduced the production of ammonium compared to the prior designs. Similarly, Liu et al. (2013)

proposed a two-step process to treat an IX brine loaded with nitrate and perchlorate; nitrate was

reduced to nitrogen gas in the first stage, while perchlorate was reduced to chloride using a Re-

Pd/C catalyst. More recently, another catalyst has been successfully developed by synthesizing

palladium and indium on activated carbon (Pd-In/C) to reduce nitrate in IX brine (Choe et al.,

2015). Meanwhile, numerous reactor configurations have been proposed to increase the stability

and the performance of catalytic reduction, such as batch reactors (Choe et al., 2015), fixed-bed

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column reactors (Choe et al., 2015; Bergquist et al., 2016), and trickle bed reactors (Bergquist et

al., 2017). In addition, Yang et al. (2013) investigated the feasibility of photocatalytic reduction

using TiO2 as a catalyst and formic acid as an electron donor for the treatment of nitrate in IX

brine. They observed that the presence of sulfate in the brine is the dominant factor preventing

the treatment, that is, there exists a competition between sulfate and nitrate on TiO2 sites.

Overall, despite being feasible, catalytic reduction is usually complicated to operate and a long-

term techno-economic analysis of these processes is still lacking.

Zero-valent iron nanoparticles. Zero-valent iron nanoparticles (Fe0) are capable of reducing

nitrate (Choe et al., 2000) and perchlorate (Moore et al., 2003) to their reduced forms. However,

the small size of the nanoparticles (1-100 nm) promoted their agglomeration and thus reduced

their reactivity. To resolve this problem, He and Zhao (2005) and He et al. (2007) proposed

modifying the nanoparticles using soluble starch or carboxymethyl sodium cellulose (CMC) as

stabilizers. These stabilizers can increase the specific surface area, physical stability, and the

reaction kinetics of Fe0 nanoparticles. In subsequent studies, researchers demonstrated the

feasibility of using stabilized Fe0 to degrade nitrate (Xiong et al., 2007) and perchlorate (Xiong et

al., 2009) in an IX brine, even though the reaction rates were limited when the NaCl concentration

exceeded 60 g/L. Further studies are needed to resolve the complex issue of the recovery and the

disposal of nanoparticles before real application.

Electrolysis. Some studies have assessed electrolysis as an alternative to remove nitrate from IX

brine. Paidar et al. (2004) developed an electrolysis reactor where a simulated bicarbonate brine

was circulated in inert media with titanium (Ti) as the anode and copper (Cu) as the cathode.

Subsequent to the reduction of nitrate to ammonium and oxygen at the cathode (Cu), the treated

brine was successfully reused for the regeneration of IX resins. In another study, Dortsiou et al.

(2009) compared tin (Sn) and bismuth (Bi) as cathodes for the electrolysis treatment of a

simulated bicarbonate brine while using platinum anode in both cases. The authors reported that

the type of cathode electrode had an impact on the final reduction products for nitrate. The final

reduction products of nitrate were mostly nitrogen gas (47%) and nitrous oxide (41%) when Sn

was used as the cathode. While, only 28% of nitrogen gas and 39 % ammonium was produced

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when Bi was used as the cathode. More recently, Duan et al. (2020) used an electrolysis system

to treat chloride brine using iron (Fe) as cathode and IrO2-RuO2/Ti as anode. They found that

ammonium generated by the reduction of nitrate was oxidized to nitrogen gas due to the

presence of chlorine (i.e., breakpoint chlorination), a by-product generated during the electrolysis

of chloride at the anode. Overall, electrolysis is an appealing technology for the treatment of

nitrate-laden IX brine. Future investigations should assess the scale-up cost and the sustainability

of this method for brine treatment.

Biodegradation. Biodegradation allows the removal of nitrate (van der Hoek and Klapwijk, 1989)

and perchlorate (Logan et al., 2001) from IX brine. The greatest challenge of this method is related

to the high salinity of the brine, which inhibits bacterial growth. To solve this problem, halophilic

bacterial communities have been isolated from various high-salinity environments, such as

seawater, marine sediments, salt marshes, and biofilm/sludge from a seawater filter (Logan et al.,

2001; Okeke et al., 2002; Cang et al., 2004). These strains were then enriched in the laboratory

and acclimated to the brine environment. Given that biodegradation is a redox process, some

reactions may benefit from the addition of an electron donor, such as methanol (Clifford and Liu,

1993), ethanol (McAdam et al., 2010), acetate (Lehman et al., 2008), lactate, and glycerol (Peyton

et al., 2001), of which the treatment efficiency depends on the pH and salinity. The addition of

these electron donors complicates the reuse of brine for drinking water applications due to carry-

over concerns. Several reactor configurations have been tested with varying degrees of success

for the treatment of nitrate, perchlorate, or the simultaneous removal of these two substances,

such as an upflow sludge blanket (Van der Hoek and Klapwijk, 1989), a sequential batch reactor

(Clifford and Liu, 1993; Lehman et al., 2008), a fluidized bed reactor (Patel et al., 2008; Xiao et al.,

2010), a membrane bioreactor (McAdam and Judd, 2008; McAdam et al., 2010), and a membrane

biofilm reactor with H2 as the electron donor (Chung et al., 2007; Sahu et al., 2009). To optimize

biological treatment, Lin et al. (2007) reported that the addition of divalent cations (Ca2+, Mg2+)

stabilizes the bacterial culture and promotes perchlorate biodegradation. Li et al. (2015)

suggested using a mixture of bicarbonate and chloride regenerants to increase the alkalinity of

the brine, which would stabilize the denitrification process. Two groups of investigators also

demonstrated the possibility of encapsulating bacteria (Trogl et al., 2011) or enzymes (Hutchison

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et al., 2013) to protect them from high salinity while allowing the degradation of nitrate and

perchlorate in the brine. So far, these methods have only been demonstrated at the lab scale.

Overall, biodegradation is a promising process for the treatment of nitrate and/or perchlorate-

laden IX brine.

Hybrid processes. Finally, hybrid processes were tested to remove nitrate and other impurities

to improve the regeneration efficiency of the recycled brine. Bae et al. (2002) tested a treatment

system consisting of biological denitrification, biological reduction of sulfate, sand filtration, and

granular activated carbon (GAC) filtration; the brine was reused for 24 times with a performance

comparable to that of the fresh brine. Later, the authors suggested using precipitation and

coagulation (with BaCl2 and FeCl3) to replace the biological sulfate reduction process (Bae et al.,

2004), and the brine was reused for 11 times with success. Klas et al. (2015) applied a sequential

batch denitrification bioreactor followed by ozonation for suspended solids and DOC removal.

The system operated reliably for over a year and achieved steady performance.

Resource recovery. Few studies have investigated the feasibility of resource recovery from

nitrate- and perchlorate-laden IX brine. Vaudevire et al. (2013) investigated the use of dynamic

vapor recompression (DVR) to recover salts from denitrified IX brine but failed to separate NaCl

from the salt mixture. Huo et al. (2020) successfully used a hybrid hydrogenation/MD process to

treat an IX brine containing a high concentration of nitrate. Briefly, nitrate was first reduced to

ammonia in the presence of a low-cost catalyst (Ru/C), and the ammonia was then recovered by

MD as (NH4)2SO4, which is a valuable commercial fertilizer.

5.4.2 Ion exchange brine disposal

Although brine reuse is beneficial for IX brine management, brine disposal is still a common

management solution, especially for the cases where: 1) IX brine reuse is neither possible or

economically viable (e.g., small-scale communities) or 2) recycled brine reaches its end of life and

the disposal of such brine is inevitable. The choice of disposal strategy is site-specific, depending

on costs, volume of brine, level of treatment required before release, physicochemical

characteristics of brine, local hydrogeological conditions, public acceptability, and local

regulations (Jensen and Darby, 2016).

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5.4.2.1 Direct disposal

Direct disposal to a nearby surface water body is a simple and economical measure. However,

this option can only be considered when no harmful contaminants are present in the brine (e.g.,

arsenic-laden brine). Mitigation measures should be considered before disposal, such as the use

of diffusers, mixing systems, or mixing zones to dilute the brine, so that the discharge of brine

may conform to local discharge criteria (e.g., for chloride).

On-site evaporation ponds are a suitable alternative for the disposal of IX brine. Briefly, when the

water evaporates due to solar heat and the wind effect in the pond, the solutes in the brine

precipitate at the bottom of the pond and this precipitate (i.e., saline sludge) must be periodically

recovered and transported elsewhere (e.g., landfill). An improved design, namely the salinity

gradient solar pond, has been proposed (Tinos and Culligan, 2012) for the disposal of desalination

brine; however, the feasibility of IX brine still needs investigation. In all cases, building a pond is

affordable (when the land is available), and the operation requires little maintenance. Jensen and

Darby (2016) demonstrated that the on-site evaporation ponds are the cheapest solution in

California among the four strategies studied when the system capacity is greater than 8 m3/h (35

GPM). However, it should be noted that this technique is only suitable for arid or semi-arid regions

where local climatic conditions are favorable (e.g., low humidity and high rate of evaporation).

A popular alternative is to discharge the IX brine into sewerage systems. This is particularly

feasible for small water treatment facilities. However, brine dilution and pH adjustment are

usually necessary prior to discharge to minimize the harmful impacts on water pipes, as well as

on the biological processes used by wastewater treatment plants (Rokicki and Boyer, 2011; Liu et

al., 2013). For remote areas where sewer systems are not available, transportation of IX brine to

the wastewater treatment plant can be an economical option. Jensen and Darby (2016) reported

that transportation to a coastal wastewater treatment plant is the cheapest option when the

water treatment plant has a design capacity lower than 8 m3/h (35 GPM).

Injecting IX brine into deep aquifers (500-1500 m) isolated from freshwater aquifers is suitable

for inland sites. However, the use of this measure strongly depends on the local geological

conditions (e.g., the existence of a confined aquifer, the transmissivity of the soil), brine

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characteristics, and public perception (Jensen and Darby, 2016), as well as having high costs and

potential risk of groundwater contamination. Therefore, this option is hardly realistic for small

systems, and this measure is an alternative for large systems when other solutions are not

available. Finally, land application in agricultural fields, parks, or golf courses is a potential option

to dispose of IX brine, especially when the brine contains beneficial nutrients, such as potassium,

NOM, and nitrate. However, dilution with fresh water is certainly necessary prior to land

application to mitigate the detrimental impact arising from high salt concentrations, making this

solution less attractive compared to other methods.

Overall, although direct disposal methods are simple to put in place, the adverse impacts of the

high salinity and harmful contaminants on the aquatic ecosystem should be fully considered

before disposal. Treatment strategies and/or dilution strategies should be considered before

disposal to lower the potential negative impacts on ecosystems.

5.4.2.2 Treatment strategies for disposal

Fewer treatment strategies have been tested for direct disposal compared to reuse. Most studies

have been devoted to reducing the volume of IX brine to favor the transportation. To begin with,

NF processes have been sought to reduce the brine volume, and numerous studies have

demonstrated that NF can reduce the brine volume by up to 85%-90% prior to its disposal

depending on the operating conditions (Schippers et al., 2004; Cartier et al., 1997; Salehi et al.,

2011; Korak et al., 2018). However, membrane fouling remains an important challenge for NF

processes. To alleviate this issue, Leong et al. (2016) successfully coupled NF with a vibratory

shear system to reduce the deposition of solutes on the surface of the membrane. Furthermore,

Ghasemipanah (2013) successfully used RO to recover water from IX brine for irrigation purposes

while reducing the brine volume. Arias-Paic and Korak (2020) proposed using forward osmosis

(FO) as an alternative to RO and NF to concentrate the brine prior to disposal. Briefly, as the

presence of a salt saturator in IX plants offers a source of chemical potential for FO, IX brine can

be concentrated by the saturated salt solution. Consequently, the authors reported that FO could

reduce the brine volume by 65%-85%. In addition to membrane processes, evaporation

crystallization can be used to further concentrate the brine, demonstrating the possibility of zero-

liquid-discharge for IX brine management (Vaudevire et al., 2013).

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5.5 Discussion and Conclusions Ion exchange (IX) with synthetic resins is a flexible, simple-to-operate, and cost-efficient

technology that can cope with a wide range of contaminants. However, IX brine arising from the

regeneration of IX resins is still a limiting factor, and the management of such by-products has

become a real bottleneck for the application of IX resins in the water treatment sector. Currently,

IX plant designs mainly consist of operational parameter calculations (e.g., flow rate, bed volume,

and running time), and a proper brine management strategy is often overlooked during the IX

plant design phase. Consequently, it is of paramount importance to integrate IX operational

strategies as well as brine management schemes in IX plant design to make the process more

sustainable.

Although it is easy to design an IX system based on the source water characteristics, defining an

optimal IX brine strategy requires treatability or pilot studies, a cost that is too often difficult to

justify for small projects. In our opinion, it would be desirable for legislators to force designers to

address the issue of IX brine management. Many solutions reviewed (e.g., switching from NaCl to

NaHCO3) involve higher operating costs but had much lower potential environmental impacts,

but it is doubtful that the industry will adopt greener alternatives unless it is required by

regulation or otherwise promoted.

Based on the IX operational strategies and IX brine management reviewed in the present paper,

a novel design workflow for IX plants is proposed as follows (Figure 5.1). First, IX operational

strategies that facilitate the management of the IX brine should be considered. When choosing

the resin during the design phase, one should select a resin that works as efficiently as possible

for the intended purpose. This requires stakeholders to test different resins from different

manufacturers for the same application and to choose the most efficient one. Second, when

designing the cycle length, if the IX process is used for NOM removal, one could consider

operating IX resins in biological mode (BIEX), which can fully exploit the IX capacity for NOM

removal and reduce the regeneration frequency. Third, during the regeneration phases,

segmented regeneration has been shown to largely reduce the brine volume, and its investigation

is encouraged, especially for chromium and nitrate removal. On the other hand, despite being

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promising, chemical-free regeneration (i.e., biological, electrolytic, and thermal methods) is still

under development, and more studies are needed to test the feasibility of full-scale applications.

Finally, alternative regenerants, such as bicarbonate salts for AIX resins and potassium salts for

CIX resins, are also encouraged to reduce the environmental impact.

Subsequent to determining the regeneration strategy, a regeneration pretest in the lab or at a

pilot-scale is strongly suggested to characterize the IX brine. This practice would benefit the IX

brine management strategy design in the following way: 1) direct reuse of IX brine could be

evaluated; 2) treatment methods could be designed if brine direct reuse is not feasible. It is worth

noting that direct reuse of IX brine is the most economically beneficial approach due to its savings

on water/salt usage as well as operation and maintenance (O&M) costs as long as regeneration

efficiency and treatment performance are not affected. IX brine treatment prior to reuse certainly

increases capital costs, but the reduction in salt and water consumption lowers the O&M costs

(Lehman et al., 2008). Nonetheless, non-monetary factors should also be considered when

choosing the best treatment method, such as process maturity, operational complexity, process

footprint, and contamination of IX vessels (Plummer et al., 2018). Overall, although IX brine

treatment followed by reuse for IX resin regeneration is potentially viable, it is still necessary to

design the treatment scheme on a case-by-case basis considering the local regulations and

conditions.

IX brine disposal is still the main option for cases where: 1) brine reuse strategies are not

economically feasible or 2) the reused brine has reached its end of life. The brine disposal

methods depend on cost, brine volume, brine characteristics, local hydrogeological conditions,

climate conditions, public acceptability, and local regulations. Treatment methods (e.g.,

membrane or evaporation) are available to reduce the brine volume or even achieve zero liquid

discharge. Overall, this review provides a vast list of options to mitigate the adverse impacts of IX

brine on the environment. Engineers and legislators are encouraged to proactively consider these

novel ideas, which will help reduce the environmental footprint of IX systems.

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Figure 5.1 - Ion exchange (IX) plant design workflow integrated with IX operational strategies and

IX brine management.

Acknowledgements

The authors would like to acknowledge Mr. Donald Ellis from the Quebec Ministry of Environment

for his valuable suggestions for this paper, and we acknowledge the NSERC CREATE program in

environmental decontamination technologies and integrated water and wastewater

management (TEDGIEER) for the Ph.D. scholarship awarded to Mr. Zhen Liu.

Determination of Targeted Contaminants

IX Resin Selection and IX Contactor Selection

Design of Cycle Length, Flow Rate and Bed Volume

Design of Regeneration Strategy

IX Brine Characterization through Pilot Test or Simulation

Reuse Feasibility Test

Determination of Reuse Times and Make-up Strategies

Evaluation of IX Brine Treatment Methods (e.g., feasibility, cost)

IX Brine Treatment and Reuse until the End of Life

IX Brine Disposal

Feasible Not Feasible

Not Viable Viable

IX Brine Reuse until the End of Life

Characterization of Raw WaterIX Operation

IX Brine Management

154

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Chapitre 6 – Conclusions et perspectives

6.1 Conclusions Dans le cadre de cette thèse, des essais de pilote et de laboratoire ont été effectués afin de

comprendre et évaluer l’application du BIEX pour l’enlèvement de la MON. En plus, une revue de

littérature a été effectuée afin de résumer les stratégies qui peuvent alléger la gestion de la

saumure engendrée par la régénération de résine. Les conclusions principales de ces travaux sont

présentées ci-dessous.

Dans le chapitre 2, les résines en forme chlorure et bicarbonate ont été opérées en parallèle sans

régénération pour 9 mois (i.e., BIEX mode). Les résultats démontrent que l’opération du filtre

BIEX pouvait être divisée en trois phases : 1) Échange d’ions préliminaire : la MON et les anions

inorganiques (tel que sulfate) échange avec des anions préchargées (i.e., chlorure et bicarbonate);

2) Échange d’ions secondaire : la MON échange avec des anions préretenus (i.e., sulfate); 3)

Épuisement : la capacité d’échange est épuisée concernant le chlorure et le sulfate, et la MON

est enlevée par la biomasse (i.e., biosorption, biodégradation et biorégénération). Trois fractions

conceptuelles de MON ont été proposées afin de comprendre l’enlèvement de la MON dans le

filtre BIEX, i.e., la MON1 a une affinité plus base que le chlorure sur la résine, et la MON2 et la

MON3 ont respectivement une plus grande et une plus base affinité par rapport au sulfate sur la

résine (séquence d’affinité : MON3>sulfate>MON2>chlorure>MON1). Par conséquent, pendant

l’échange d’ions préliminaire, la MON2 et la MON3 ont été retenues par la résine alors que la

MON1 a relargué du filtre BIEX. Cependant, à la fin de l’échange d’ions préliminaire, la MON3 et

le sulfate échange avec la MON2, ce qui entraîne une percée de COD dans l’affluent du filtre.

Pendant l’échange d’ions secondaire, la MON 3 échange avec le sulfate préretenu jusqu’à

l’épuisement de la résine. Finalement, la MON peut seulement être enlevée par la biomasse

présente sur la résine pendant la phase de l’épuisement. La caractérisation de la MON dans

l’effluent du filtre BIEX démontre que la MON2 consiste principalement en substances humiques

et blocs de construction avec une faible densité de charge, basse masse moléculaire (<1000 Da)

et faible aromaticité.

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Dans un deuxième temps, le chapitre 2 démontre que la performance du filtre BIEX en forme

chlorure est similaire à celui en forme bicarbonate. Pendant la durée de l’étude, le BIEX en forme

chlorure a réalisé un enlèvement de COD légèrement plus elevé par rapport au BIEX en forme

bicarbonate (médiane : 53% vs 49%), alors que le BIEX en forme bicarbonate a réalisé un

enlèvement de CODB plus haut par rapport au BIEX en forme chlorure (médiane : 50% vs 33%).

Par ailleurs, les filtres BIEX en forme chlorure et bicarbonate ont eu une efficacité similaire en

termes d’enlèvement de précurseurs de SPD. Ces résultats permettent de tirer la conclusion que

les performances du filtre BIEX en forme bicarbonate ne différencie pas significativement du BIEX

en forme chlorure en ce qui concerne l’enlèvement de la MON.

Dans le chapitre 3, les résines colonisées ont été prélevées du pilote en fonction à l’usine de

traitement pour le test en batch où les composés avec des charges et potentiels de

biodégradabilités différents ont été mise en contact avec les résines biotiques et abiotiques. Les

résultats ont démontré que les composés biodégradables (tels que caféine et ibuprofène)

pouvaient être biodégradés par la biomasse présente sur le BIEX. Cependant, la biodégradation

dépend aussi de la communauté microbienne présente sur la résine étant donné que naproxène,

un composé biodégradable qui n’est pas présenté dans la source d’eau, ne pouvait pas être

éliminé par la biodégradation sur le BIEX. En plus, les composés négativement chargés, tel que

naproxène, PFOA et PFOS peuvent être enlevés par l’échange avec la MON préretenue.

Finalement, les composés neutres et non-biodégradables (Thiaméthoxame et atrazine), ne

peuvent pas être enlevés par le BIEX. En bref, l’enlèvement de la matière organique par le BIEX

dépend non seulement du caractère des composés (par exemple, charge et biodégradabilité)

mais aussi de la communauté microbienne présente sur la résine.

Dans le chapitre 4, un pilote qui consiste en BIEX, CAG et CAB a été opéré en filtration secondaire

dans l’usine de production d’eau potable. Cette étude a pour but d’évaluer l’application du BIEX

dans l’usine de production d’eau potable. Les résultats ont démontré que le filtre BIEX avait

réalisé un enlèvement de COD plus élevé par rapport aux filtres CAG et CAB (29%-36% vs 13%-

24%) pour la durée d’étude (9 mois). D’ailleurs, le filtre BIEX a aussi réalisé un enlèvement de

précurseurs de SPD par rapport aux filtre CAG et CAB. Ces résultats ont indiqué que le BIEX est

plus performant par rapport aux filtre CAG et CAB en filtration secondaire en ce qui concerne

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l’enlèvement de la MON. Cependant, la percée de COD entraînée par l’élution de différentes

fractions de MON peut rehausser temporairement le COD dans l’affluent du filtre. Ce problème

pourrait être résolu en mettant le filtre BIEX hors service pendant la percée de COD. Par ailleurs,

pendant l’opération en été, le filtre BIEX a relargué de l’azote ammoniacal (15%) alors que les

filtres CAG et CAB ont pu enlever de l’azote ammoniacal. Le relarguage de l’azote ammoniacal est

indésirable étant donné qu’il peut donner un goût désagréable à l’eau potable. Ce relarguage de

l’azote ammoniacal du filtre BIEX est principalement dû au petit diamètre de la colonne et la faible

vitesse de l’air pendant le rétrolavage. Dans le chapitre 2, un enlèvement de l’azote ammoniacal

a été observé dans le filtre BIEX où le diamètre de colonne est plus grand et la vitesse de l’air est

plus forte par rapport au pilote du chapitre 4. Donc, un propre rétrolavage est nécessaire pour

que l’azote ammoniacal soit éliminé dans le BIEX. Finalement, puisque la bille de résine est

généralement plus petite par rapport au charbon actif en grain (CAG), la perte de charge du filtre

BIEX est souvent plus grande que les filtres CAG et CAB. Cela indique que le contacteur du BIEX

serait plus profond que celui du CAG et CAB ou le BIEX devrait être opéré sous pression. Un

contacteur plus profond augmentera le coût capital alors que l’opération sous pression

compliquera la maintenance du filtre et le rétrolavage (dû au contacteur fermé). Le rapport

efficacité-prix du BIEX pour l’enlèvement de la MON demeure à investiguer.

La comparaison entre le chapitre 2 et le chapitre 4 permet d’évaluer la position du filtre BIEX dans

une chaîne de traitement de l’eau potable. Dans le chapitre 2, le filtre BIEX a été alimenté

directement par l’eau brute. Le filtre BIEX a réalisé un enlèvement moyen de COD d’environ 50%

pour une durée de 9 mois, ce qui suggère que le BIEX pourrait être la stratégie principale pour

l’enlèvement de la MON (i.e., un alternatif à la coagulation). Cependant, étant donné que le filtre

BIEX n’enlève pas de turbidité, des filtres MF ou UF devraient être placés avant le BIEX afin

d’enlever la turbidité et diminuer le colmatage du filtre. Par ailleurs, le filtre BIEX peut aussi être

mis en filtration secondaire en tant que processus de polissage. Cependant, l’enlèvement de COD

serait moins efficace (environ 32%) par rapport à celui alimenté par l’eau brute (50%). C’est parce

que le caractère de MON a été changé après le traitement conventionnel (plus hydrophile après

la coagulation) et le sulfate provenant de l’utilisation de l’alun peut rivaliser avec la MON sur la

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résine. Plus d’études sont nécessaires afin de comparer le coût d’opérer le filtre BIEX aux

différentes positions dans une chaîne de traitement d’eau potable.

Dans le chapitre 5, une revue de littérature a été effectuée afin de résumer les stratégies qui

peuvent alléger la gestion de la saumure engendrée par la régénération de résines. D’ailleurs, un

nouveau schéma de conception a été proposé pour le processus d’échange d’ions. Au niveau de

l’opération de résine, on devrait choisir la résine la plus performante pour l’application destinée.

La résine SST (Shallow shell technology) et la résine non-régénérable peuvent aussi être

considérées afin de diminuer ou même éliminer la production de la saumure. L’application de

configuration « lead-lag » est encouragée parce qu’elle peut maximiser la capacité de la résine et

ainsi diminuer la fréquence de régénération. Les régénérations biologique, électrochimique et

thermale sont encore en développement et leurs applications, surtout à pleine échelle,

nécessitent encore des études. Au niveau de la gestion de la saumure, la réutilisation de la

saumure est toujours encouragée pourvue que la réutilisation n’impacte pas la performance du

cycle subséquent. Dans le cas où la réutilisation est non faisable, les technologies du traitement

de la saumure sont disponibles. Cependant, il faut noter que le traitement de la saumure

augmente le coût capital et opérationnel, et son rapport efficacité-prix devrait soigneusement

évalué par rapport à l’utilisation de la saumure fraîche. Finalement, le rejet de la saumure est

inévitable au cours de l’opération du processus d’échange d’ion, et la salinité et les substances

nocives présentes dans la saumure devraient être bien évaluées avant le rejet.

6.2 Perspectives Bien que les fractions conceptuelles de MON (i.e., MON1, MON2 et MON3) permettent d’élucider

l’échange de la MON contre la résine, les techniques de caractérisation de la MON ne pouvait pas

bien identifier les trois fractions dans cette étude. D’autre techniques de caractérisation devraient

être appliquées afin de bien distinguer ces trois fractions, tel que la mesure de densité de charge

par zetasizer. Par ailleurs, la fraction de MON dépend aussi du taux d’utilisation de la résine. Les

résines utilisées dans cette étude étaient toujours neuves. Donc, l’étude au futur devrait aussi

caractériser les fractions conceptuelles de MON avec les résines usées.

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Dans le cadre de cette étude, les résines en forme chlorure et bicarbonate ont été évaluées pour

l’application de BIEX. Cependant, l’évaluation de la résine en forme sulfate serait également

importante. Étant donné que la résine en forme sulfate peut éviter l’enlèvement non nécessaire

d’ anions inorganiques, telles que bicarbonate, chlorure et sulfate, ceci permet d’éviter la percée

de COD pendant l’opération du BIEX. Néanmoins, puisque le sulfate a une affinité plus grande par

rapport au chlorure et bicarbonate, l’enlèvement de la MON serait moins efficace en utilisant la

résine en forme sulfate. En somme, la performance du BIEX en forme sulfate pour l’enlèvement

de la MON nécessite plus d’études.

Le filtre BIEX s’avère efficace tant comme processus principal que comme processus de polissage

pour l’enlèvement de la MON. Cependant, son impact sur les autres processus de traitement

demeure à investiguer. Par exemple, quels sont les impacts sur l’ozonation si l’eau est prétraitée

par le BIEX? Quels sont les impacts sur la désinfection si l’eau est prétraitée par un processus de

polissage de BIEX? Par ailleurs, le coût du BIEX dans une filière de traitement de l’eau potable

devait être évaluée systématiquement, parce que l’intégration du BIEX aurait des impacts sur les

processus en aval (par exemple, la réduction de la consommation de produits chimiques). En bref,

les positions du BIEX dans la filière de traitement de l’eau potable demeurent à comparer en

tenant compte de la performance et des coûts.

Bien que le BIEX soit une alternative prometteuse à la coagulation, l’adsorption et aux

membranes, une étude d’analyse du cycle de vie qui compare ces technologies est encore

manquante. Une telle étude permettrait d’évaluer les impacts environnementaux de ces

technologies et ainsi fournir un nouveau point de vue lors de la sélection du processus pour

l’enlèvement de la MON.

Le problématique des microplastiques a reçu des intérêts énormes récemment dû à l’impact

néfaste potentiel sur la faune. La résine échangeuse d’ions est elle-même un type de

microplastique (<5 mm). Ainsi, il serait important d’évaluer le relarguage involontaire de résines

pendant l’opération du processus d’échange d’ions. La résine pourrait se dégrader dû au

changement de pression dans le contacteur et ainsi relarguer dans l’effluent du filtre. D’ailleurs,

la résine pourrait aussi relarguer à l’environnement par le rejet de l’eau de rétrolavage. En bref,

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la quantité de résines relargués et ses impacts potentiels sur l’environnement devraient être

évalués soigneusement.

La revue de littérature a permis d’identifier de nombreux champs d’études qui peuvent rendre le

processus d’échange d’ions plus vert. La régénération sans ajout de produits chimiques (i.e.,

régénération biologique, électrochimique et thermale) a démontré des résultats prometteurs au

laboratoire, mais leurs applications à l’échelle réelle nécessitent encore des études. De

nombreuses technologies sont disponibles en ce qui concerne le traitement de la saumure avant

la réutilisation. Cependant, la plupart de ces technologies sont en développement et leurs

applications à l’échelle réelle sont encore hasardeuses. En bref, les technologies qui peuvent

rendre le processus d’échange d’ions plus vert sont prometteuses mais immatures, et plus

d’efforts sont nécessaires afin de rendre l’échange d’ions plus durable en tant que technologie de

traitement de l’eau.

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Annexe A – Biological ion exchange as an alternative to

biological activated carbon for natural organic matter

removal: impact of temperature and empty bed contact time

(EBCT)

Abstract Biofiltration is a widely used process in drinking water treatment plants to remove natural organic

matter (NOM). A novel biofiltration process using ion exchange resins as supporting media (i.e.,

biological ion exchange or BIEX) has been demonstrated to provide a superior performance

compared to conventional biological activated carbon (BAC). In order to optimize the

performance of BIEX filters, the impact of temperature and empty bed contact time (EBCT) on

NOM removal was systematically studied. In the present study, bench-scale BIEX filters were set

up in parallel with BAC filters and operated at different temperatures (i.e., 4 °C, 10 °C and 20 °C)

and EBCTs (i.e., 7.5 min, 15 min and 30 min). Higher average dissolved organic carbon (DOC)

removal was achieved in BIEX filters (73 ± 6 %) than BAC filters (22 ± 9 %) at the steady state with

an EBCT of 30 min. Higher temperatures improved NOM removal in both BAC and BIEX filters,

with the impact being greater at lower EBCTs (i.e., 7.5 min and 15 min). Higher EBCTs could also

improve NOM removal, with the impact being greater at lower temperatures (i.e., 4 °C and 10

°C). DOC removal for BIEX and BAC filters can be modeled with a first-order kinetic model (R2 =

0.93-0.99). BAC had a higher temperature activity coefficient than BIEX (1.0675 vs. 1.0429),

indicating that temperature has a greater impact on BAC filtration than BIEX filtration. Overall,

temperature and EBCT must be considered simultaneously for biofilters to efficiently remove

NOM.

Keywords: Biological ion exchange (BIEX); biological activated carbon (BAC); natural organic

matter (NOM); kinetics; temperature; Empty bed contact time (EBCT)

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A.1 Introduction Removing natural organic matter (NOM) is often one of the major objectives of drinking water

treatment, as NOM can cause aesthetic problems for finished water (i.e., colors, tastes and odor)

(Edzwald, 2010), disrupt water treatment processes (e.g., membrane fouling) (Kennedy et al.,

2008), contribute to biofilm regrowth in distribution systems (Hijnen et al., 2018) and react with

disinfectants forming disinfection by-products (DBPs) (Krasner et al., 2006). NOM is typically

removed using technologies such as coagulation/flocculation (Matilainen et al., 2010), membrane

filtration (Lamsal et al., 2012), activated carbon adsorption (Velten et al., 2011), ion exchange (IX)

(Bolto et al., 2002) and biological filtration (i.e., biofiltration) (Urfer et al., 1997). Biofiltration has

been extensively studied as an eco-efficient process for NOM removal. Briefly, granular media

filters naturally convert into biological filters (i.e., biofilters) when no disinfectant is used during

the operation, allowing microorganisms to develop a biofilm on the surface of media and

contribute to NOM removal (Zeraley and Summers, 2012). Generally, greater NOM removal has

been reported for biofilters using activated carbon as supporting media (i.e., biological activated

carbon or BAC) compared to sand, anthracite, or their combinations (Wang, 1995; Urfer et al.,

1997; Liu, 2001; Basu et al., 2016). Because of its greater porosity, BAC filters can support greater

biomass density than other biofilters (Emelko et al., 2006). Given BAC filtration is easy-to-operate,

economically feasible and environmentally friendly, it is widely used in water treatment plants

for NOM removal (Korotta-Gamage and Sathasivan, 2017).

In 2017, we proposed using IX resins as supporting media for biofiltration, a process referred to

as biological ion exchange (BIEX). BIEX can be achieved by operating fixed bed IEX filters without

regeneration to promote the growth of microorganisms on the resins (Schulz et al., 2017). Our

lab-scale investigations revealed that BIEX filters could achieve significantly higher NOM removal

(60 %) than BAC filters (15 %) (Winter et al., 2018). More recently, pilot-scale investigations that

compared NOM removal with BAC, BIEX, granular activated carbon (GAC) and conventional IX

filtration (i.e., with weekly regeneration) demonstrated that despite being less efficient than a

conventional IX filter, the BIEX filter achieved a higher NOM removal (62 %) than did the BAC filter

(7 %) (Amini et al., 2018). The superior performance for BIEX filtration compared to BAC filtration

was primarily attributed to the ion exchange process, i.e., NOM firstly exchanged with pre-

178

charged chloride on the resin (i.e., primary ion exchange) and then exchanged with pre-retained

sulfate (i.e., secondary ion exchange) (Liu et al., 2020; Edgar et al., 2021). Mass balance studies

indicated that a maximum of 30 % NOM removal in BIEX filters was due to biodegradation (Amini

et al., 2018; Liu et al., 2020).

The performance of biofilters is highly impacted by water temperature given that temperature

directly impacts the rates associated with microbial metabolism as well as mass transfer in

biofilters (Seger and Rothman, 1996; Urfer, 1997; Terry and Summers, 2018; Moona et al., 2019).

Generally, greater NOM removal has been reported for biofilters operated at higher

temperatures than those operated at lower temperatures. For instance, Moll et al. (1999)

reported that biofilters achieved 42 % higher NOM removal at high temperatures (20 °C-35 °C)

compared to those operated at 5 °C. The impact of temperature was also linked to the empty bed

contact time (EBCT) in the biofilters. A longer EBCT leads to longer exposure of water to microbial

colonies and thus increases the amount of NOM to be removed (Moona et al., 2021). For instance,

Hozalski et al. (1999) observed that NOM removal at low temperature could be improved by

applying a longer EBCT, indicating that applying a longer EBCT could offset the negative impact

brought by temperature decline. In order to improve the NOM removal efficiency in biofilters, it

is of importance to understand how temperature and EBCT impact the performance of biofilters

and thereby determine the best operation strategies for biofiltration.

While the impact of temperature and EBCT on conventional biofilters with activated carbon,

anthracite or sand as supporting media has been extensively investigated, the impact of

temperature and EBCT on BIEX filters still remains unclear. Amini et al. (2018) reported on the

impact of temperature on pilot-scale BAC and BIEX filters, but their study was performed under

seasonally dynamic temperature conditions. In addition, their study did not consider the

combined effect of temperature and EBCT on NOM removal.

Considering that NOM removal is achieved through a combination of ion exchange and

biodegradation, the impact of temperature and EBCT on BIEX filters is expected to be different

from that on BAC filters. Hence, the objectives of the present study were to 1) evaluate the

performance of BIEX filters for NOM removal, and 2) evaluate the impact of temperature and

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EBCT on BIEX performance for NOM removal. The study considered BAC filters as a baseline

against which the performance of BIEX filters could be compared. Bench-scale BIEX filters

operated at different conditions were set up in parallel with BAC filters and their performance for

removing dissolved organic matter (DOC) was monitored for 150 days. The comparison was

assessed based on the extent and rate of NOM removal while the impact of temperature was

quantified using temperature activity coefficients.

A.2 Materials and Methods

A.2.1 Feed water

Jericho Pond water (Vancouver, Canada) was collected and transported to the University of British

Columbia on a monthly basis from December 2017 to April 2018 and then stored in dark at 4 °C

until use. Jericho Pond water was selected for the present study as it contains a broad range of

NOM fractions (biopolymers, humic substances and low molecular weight acids) based on the

analysis using size exclusion chromatography coupled with organic carbon detection (LC-OCD)

(Schulz et al., 2017; Winter et al., 2018). The pond water was filtered through 1.5 μm pore size

glass fiber filters (VWR Glass Microfiber 691) immediately after collection. Prior to use as feed

water, the filtered pond water was diluted with dechlorinated tap water to achieve a targeted

total organic carbon (TOC) of approximately 4.5 mg C/L. The actual DOC concentration of the feed

water ranged from 3.2 to 4.8 mg C/L during the study period (Table A1). The feed water had a

neutral pH and low concentrations of anions (i.e., chloride and sulfate), and NOM present in the

feed water possessed a high biodegradable fraction (approximately 23 % of the feed water DOC).

Table A1 - Feed water physicochemical characteristics. Values and confidence interval

respectively correspond to average and standard deviation of measurements during the study

period.

Parameters Unit Value

pH - 7.0 ± 0.1

Turbidity NTU 0.22 ± 0.02

TOC mg C/L 4.5 ± 0.9

180

BDOC mg C/L 0.9 ± 0.1

DOC mg C/L 4.0 ± 0.8

SUVA L/(mg C∙m) 3.9 ± 0.4

Chloride mg/L 5.0 ± 3.1

Sulfate mg/L 5.7 ± 2.1

A.2.2 Filtration media

Purolite® A860, a strongly basic macroporous polyacrylic anion exchange resin, was used for all

BIEX filters. The IEX capacity of the resin was measured to be 0.68 eq/L with an initial form of

chloride (Amini et al., 2018), and the average IEX bead diameter of 0.75 mm. Wood-based

Picabiol® GAC, with an effective diameter of 1.04 mm, was extracted from an exhausted pilot-

scale GAC filter (> 5 years of operation) and used for all BAC filters.

A.2.3 Bench-scale systems

Three sets of bench-scale systems were used with each set operated at temperature of 4 °C, 10

°C or 20 °C. The systems operated at 4 °C and 10 °C were housed in temperature-controlled rooms

while the systems operated at 20 °C were in an open laboratory at ambient temperature. Each

set of systems consists of a feed water tank (19 L, Pyrex Tank), feed pumps (Masterflex), two BIEX

filters, two BAC filters and a filtrate tank. The components of the systems operated at 20 °C were

covered in aluminum foil to minimize potential algal growth whereas the systems operated at 4

°C and 10 °C were in dark rooms and therefore algal growth was not a concern. The filters had an

internal diameter of 1.25 cm and media depth of approximately 20.0 cm, corresponding to a bed

volume (BV) of 24.5 mL. Each filter had three effluent ports (Port 1, Port 2 and Port 3)

corresponding to an EBCT of 7.5 min (192 BV/day), 15 min (96 BV/day) and 30 min (48 BV/day),

respectively (Figure A1). The filters were operated at a filtration rate of 0.4 m/h (0.82 mL/min).

The filters were backwashed when the height of the water column above the surface of the media

reached 1 m. To backwash the filters, feedwater was connected to the Port 3 and the pump was

operated at maximum flow to fluidize the media and dislodge excess biomass from the media.

The systems were operated for a total period of 150 days, during which the filters operated at 20

181

°C were backwashed approximately every 6 to 8 weeks. The filters operated at 4 °C and 10 °C

never required backwash.

During the operation of the systems, the effluent typically flowed through Port 3 (with Port 1 and

2 being closed), yielding an EBCT of 30 min. Periodically (approximately once per week), the

effluent flow was diverted to either Port 1 or 2 (with other Ports being closed) for sampling. When

effluent flowed through Port 1 or 2, the effluent EBCT was 7.5 min or 15 min, respectively. The

use of Ports 1, 2 and 3 enabled different EBCT to be considered in a single bench-scale system. To

ensure that diverting the effluent flow to Port 1 or 2 did not impact the overall system EBCT, the

diversions were limited to once per week for a maximum of 2 hours. In addition, diversions to

Port 1 or 2 were never within 2 days of each other.

Figure A1 - Bench-scale biofiltration system (single filter illustrated).

A.2.4 Analytical Methods

TOC and DOC were quantified using a Phoenix 8000 TOC analyzer (Dohrmann, US). Ultraviolet

absorbance at 254 nm absorbance (UV254) was measured using a UV300 UV-vis spectrometer

(Spectronic Unicam, USA). DOC and UV254 samples were filtered through 0.45 μm pore size disk

filters (Supor®, Pall) prior to analysis. Biological DOC (BDOC) was measured according to Servais

et al. (1989) and Markarian et al. (2010). Briefly, DOC reduction was measured in water samples

subsequent to an incubation of 30 days with an inoculum of suspended bacteria from the raw

water. Anion concentrations (i.e., chloride and sulfate) were quantified using an ion

chromatography system (DIONEX ICS-1100). Samples were filtered through 0.45 μm PVDF syringe

FeedTank

Media levelPort 1, 5 cm

Port 2, 10 cm

Port 3, 20 cm

Overflow,1 m

182

filters (Millex®-HV) prior to analysis for anions. One blank and multiple standards were analyzed

along with each series of sample analyses. All analyses were performed in duplicate, and average

of the measurements are reported.

A.2.5 Data analysis

To account for slight differences in feed water characteristics, effluent concentrations were

normalized with respect to those in the feed (C/C0).

The DOC removal kinetics for the biofilters were quantified by fitting the data at different EBCT

to the linearized form of a first-order model (presented in equation A1) using linear regression.

First-order relationships are commonly used to model the DOC removal kinetics in biofilters (Urfer

et al., 1997; Black and Bérubé, 2014; Terry and Summers, 2018). The model considered that a

fraction of the DOC cannot be removed (i.e., residual DOC).

%!%"= %#

%"+ %""%#

%"8"&' (A1)

where Ct (mg C/L) is the DOC concentration after an EBCT of t (min); C0 is the initial DOC

concentration in the feed water (mg C/L); k is the DOC removal rate constant (min-1), and CR is the

residual DOC concentration in feed water (mg C/L). CR was selected to minimize the residual sum

of squares.

The temperature activity coefficients (θ) were estimated by fitting the estimated DOC removal

rate constants to the linearized form of the commonly used power law relationship, presented in

equation A2, which is derived from the Arrhenius equation (Peleg et al., 2012) using linear

regression.

9( = 9$):(("$)) (A2)

where kT is the DOC removal rate constant at a given temperature T (°C). k20 is the NOM removal

rate constant at 20 °C, and q is the temperature activity coefficient.

183

The errors associated with the estimated parameters and normalized concentrations, as well as

the significance associated with all comparisons (e.g., ANOVA test), are based on p = 0.05.

A.3 Results and discussion

A.3.1 DOC removal

Typical normalized DOC concentrations (DOC/DOC0) for the effluents from Port 3 are presented

in Figure A2 as a function of bed volumes treated. Similar results were observed for all conditions

investigated (data not shown). During the first few weeks of operation, the normalized DOC

concentrations in the BIEX and BAC filter effluents gradually increased. Following the first few

weeks of acclimatization period, the normalized DOC concentrations in the BIEX and BAC filter

effluents remained relatively constant. However, during this latter steady state period, the

magnitude of the normalized DOC concentrations was significantly different for BIEX and BAC

filter effluents. Unless stated otherwise, in the sections that follow, all analyses are based on

average normalized DOC concentrations measured during the acclimatization or steady state

periods.

For the BIEX filters, approximately 60 % of the influent DOC (60 ± 7 %) was removed during the

acclimatization period (i.e., normalized DOC of 0.4). During this initial period (before

approximately 40 days), a microbial community was likely not yet fully established in the filters

and therefore the removal of DOC was attributed mainly to ion exchange. At steady state (after

approximately 40 days), over 70 % of the influent DOC (73 ± 6 %) was removed (i.e., normalized

DOC of 0.27). Previous studies indicated that resins became biologically active after an operation

of 40 days based on the measurement of adenosine triphosphate (ATP) (Amini et al., 2018),

suggesting that the resins in the present study were already biologically active at the steady state.

DOC removal at steady state was attributed to a combination of ion exchange and biodegradation

(Amini et al., 2018; Liu et al., 2020). Note that although the primary ion exchange capacity of the

filters defined based on chloride release was exhausted after approximately 4000 BV (data not

shown), secondary ion exchange based on the release of pre-exchanged ions (e.g., sulfate) was

not exhausted (data not shown) (Liu et al., 2020). As a consequence, ion exchange was expected

184

to have been the dominant mechanism for DOC removal throughout the 150 days study period

(Amini et al., 2018; Liu et al., 2020). A higher DOC removal for the BIEX filter was achieved at

steady state in the present study compared to that reported in previous investigations (Winter et

al., 2018; Amini et al., 2018; Liu et al., 2020). This was likely due to the different characteristics of

the feed water used in these studies (e.g., the feed water in the present study had a higher BDOC

concentration and/or a lower sulfate concentration than those in the previous studies by Amini

et al. (2018), Winter et al. (2018), and Liu et al. (2020)).

For the BAC filters, more than 45 % of the influent DOC (47 ± 12 %) was removed during the

acclimatization period (i.e., normalized DOC of 0.53). At steady state, approximately 20 % of the

influent DOC (22 ± 9 %) was removed (i.e., normalized DOC of 0.78). The greater removal of DOC

during the acclimatization period than the steady state period was attributed to adsorption onto

GAC used or the BAC filters. It is likely that the manipulation of the harvested GAC for BAC filters

revealed adsorption sites that had previously been unavailable. The removal of DOC at steady

state was attributed to biodegradation. The magnitude of DOC removal in the present study

within the BAC filters is consistent with those reported in a previous study conducted at 20 °C

using the same feed water (Black and Bérubé, 2014) as well as those reported in previous studies

by others over the range of temperatures considered (Terry and Summers, 2018).

Figure A2 - Typical normalized DOC in the BIEX and BAC filter effluents for the different

temperature considered. Results presented for Port 3 corresponding to an EBCT of 30 min.

0

0.2

0.4

0.6

0.8

1

1.2

0 2000 4000 6000 8000

DOC/DOC 0

Bed volumes

BIEX 4°C

BIEX 10°C

BIEX 20°C

BAC 4°C

BAC 10°C

BAC 20°C

BAC

BIEX

150 days

185

A.3.2 Impact of temperature and EBCT on DOC removal

The average DOC removal achieved for the different conditions investigated are presented in

Figure A3. BIEX filters consistently achieved a greater DOC removal (47 %-76 %) than did the BAC

filters (2 %-27 %). In general, an increasing temperature increased DOC removal with the

beneficial impact of temperature being greater at lower EBCTs. Similarly, an increase in EBCT

increased DOC removal with the beneficial impact of EBCT being greater at lower temperatures.

An ANOVA analysis revealed that for the BIEX filters, the increase of temperature could improve

the DOC removal in BIEX filters at an EBCT of 7.5 and 15 min (p < 0.05) whereas temperature had

little impact on the performance at an EBCT of 30 min (p > 0.05). Further, greater EBCTs can

significantly increase the DOC removal for BIEX filters operated at 4 °C and 10 °C, whereas the

DOC removal was improved to a lesser extent for BIEX filters operated at 20 °C (p < 0.05).

For the BAC filters, the increase of temperature constantly improved the DOC removal in BAC

filters for all investigated EBCTs (p < 0.05) even though the improvement on DOC removal was

more significant at low EBCTs (i.e., 7.5 min and 15 min). The increase of EBCT could improve the

DOC removal for BAC filters operated at all temperature investigated, but a more significant

improvement was observed for BAC filters operated at lower temperatures (i.e., 4 °C and 10 °C).

These results are consistent with those from previous studies where temperature was reported

to have a more significant impact on the performance of biofilters at shorter EBCT (Persson et al.,

2006; van der Aa et al., 2011). The outcomes also highlighted that the impact of temperature and

EBCT cannot be investigated separately, as the performance of biofilters for NOM removal

depends on both temperature and EBCT.

186

Figure A3 - Average DOC removal in BIEX and BAC filters during steady state (after about 40 days

of operation) for the different conditions investigated. Error bars correspond the standard error

of averages during the steady state period.

A.3.3 Impact of temperature on DOC removal kinetics

As previously discussed, the DOC removal kinetics for the BIEX and BAC filters were quantified by

fitting a first-order model (equation A1) to the normalized DOC concentrations at the different

EBCT. The model fitted to the normalized DOC vs. EBCT for the different temperature conditions

investigated is presented in Figure A4. For all conditions investigated, the first-order model could

adequately model the normalized DOC concentrations at EBCTs considered (R2 =0.93-0.99).

The normalized DOC removal rate constant and the residual normalized DOC concentration

estimated from the first-order model fitted to the data for the different conditions investigated

are listed in Table A2. Overall, the normalized DOC removal rate constants for the filters increased

as the operating temperature increased even though the increase was not consistently significant.

No consistent trend was observed between temperature and the residual normalized DOC

concentration. For the BIEX filters, the DOC removal rate constant increased from 0.12 ± 0.031

47

60

71

28

15

52

6571

814

23

62

75 76

11

19

27

0

10

20

30

40

50

60

70

80

90

BIEX7.5min

BIEX15min

BIEX30min

BAC7.5min

BAC15min

BAC30min

DO

C re

mov

al (%

)

4°C 10°C 20°C

187

min-1 to 0.24 ± 0.016 min-1 as the temperature increased from 4 °C to 20 °C. All BIEX filters had a

similar residual normalized DOC concentration (CR/C0 = 0.23-0.27). This was expected because

although temperature was expected to impact the DOC removal mechanisms in terms of kinetics,

it was not expected to impact the fraction of non-removable DOC in the feed water (i.e., CR). For

the BAC filters, the DOC removal rate constant increased from 0.02 ± 0.021 min-1 to 0.06 ± 0.003

min-1 as the temperature increased from 4 °C to 20 °C. The DOC removal rate constant at 20 °C

was similar to those reported in a previous study conducted using the same feed water (Black and

Bérubé, 2014). The DOC removal rate constants for all temperature considered were also

consistent with typical first-order DOC removal rate constant (0.02-0.18 min-1) for BAC filters

reported by others (Son et al., 2015; Terry and Summers, 2018). Similar normalized residual DOC

concentrations were observed for all the BAC filters (0.64-0.68), indicating that temperature does

not impact DOC removal mechanisms in terms of capacity.

Overall, the rate and extent of DOC removal were greater for BIEX filters than for BAC filters for

all conditions investigated. This is likely because that DOC was removed through a combination

of ion exchange and biodegradation in the BIEX filters, whereas DOC was only removed through

biodegradation in the BAC filters (Amini et al., 2018; Liu et al., 2020). Therefore, even with higher

EBCTs, BAC filters cannot remove DOC to the same extent as BIEX filters.

0

0.2

0.4

0.6

0.8

1

1.2

1.4

0 10 20 30 40

DOCC

/DO

CC0

Empty bed contact time (min)

BIEX 4°C BAC 4°CBIEX 10°C BAC 10°CBIEX 20°C BAC 20°C

BAC

BIEX

188

Figure A4 - Normalized DOC vs. EBCT at different temperatures. Note that the normalized DOC

data is the same as in Figure A3 but presented with equation A1 fitted to the data. Confidence

intervals were omitted because of overlapping with data symbols.

Table A2 - Normalized DOC removal rate constants, normalized residual DOC concentrations and

temperature activity coefficients for BIEX and BAC filters. The confidence interval as well as values

in the paratheses corresponds to the standard error of the estimated parameters.

Biofilter

type

Temperature

(°C)

Rate

constant (kT)

(min-1)

CR/C0

(-)

Temperature activity

coefficients (θ)

(Present study)

Temperature activity

coefficients (θ)

(Previous study)

Biological

ion

exchange

(BIEX)

4 0.12 ± 0.031 0.27 1.0429

(1.0429-1.0430) 1.04* 10 0.16 ± 0.011 0.28

20 0.24 ± 0.016 0.23

Biological

activated

carbon

(BAC)

4 0.02 ± 0.021 0.65 1.0675

(1.0660-1.0690) 1.04-1.06* 10 0.03 ± 0.004 0.64

20 0.06 ± 0.003 0.68

*Estimated based on the data from Amini et al. (2018) and Terry and Summers, (2018).

The temperature activity coefficients (θ), derived from the Arrhenius equation, can be used to

relate rate constants estimated at different temperatures and to report rate constants estimated

at multiple temperatures using a single rate constant at a reference temperature (e.g., 20 °C). A

temperature activity coefficient greater than 1 indicates that rate constants, and therefore a

reaction, are positively impacted by temperature. The temperature activity coefficient for the

biofilters was quantified by fitting equation 2 to the removal rate constants estimated at different

temperatures. Equation 2 fitted to the DOC removal rate constants vs. temperature for BIEX and

BAC filters is presented in Figure A5. For both BIEX and BAC filters, equation 2 could adequately

model the impact of temperature on DOC removal rate constant (R2 =0.97-0.99). The temperature

189

activity coefficients were estimated to be 1.0429 (1.0429-1.0430) for BIEX biofilters and 1.0675

(1.0660-1.0690) for BAC biofilters. For both BIEX and BAC filters, the temperature activity

coefficients were significantly greater than 1, indicating that temperature impacted DOC removal

for both filters. Also, the estimated temperature activity coefficient for the BAC filters was

significantly greater than that for the BIEX filters, indicating that temperature has a greater impact

on DOC removal in BAC filtration than in BIEX filtration. For the BIEX and BAC filters, the rate

constants at a reference temperature of 20 °C were calculated to be 0.24 ± 0.016 min-1 and 0.06

± 0.003 min-1, respectively (Table A2).

Previous studies have not reported the impact of temperature on biofilters using temperature

activity coefficients. Nonetheless, it was possible to calculate temperature activity coefficients for

some previous studies based on reported data. As presented in table 2, the temperature activity

coefficient of BIEX filters obtained in the present study (i.e., bench-scale tests under constant

temperature conditions) is similar to that calculated from Amini et al. (2018) for pilot-scale tests

conducted under seasonally dynamic temperature conditions. These results suggest that feed

water characteristics have minimal impact on the temperature activity coefficient for BIEX filters.

We hypothesize that ion exchange is less impacted by the feed water characteristics in terms of

temperature activity coefficient. However, the temperature activity coefficient of BAC filters

obtained in the present study is greater than that calculated from Amini et al. (2018), suggesting

that feed water characteristics impact the temperature activity coefficient for BAC filters. In

contrast, the temperature activity coefficient estimated in the present study for BAC filters is

similar to that calculated based on Terry and Summers, (2018) who reported results from an

extensive literature review that assessed DOC removal in biofilters. It is likely that, in addition to

temperatures, the characteristic of the DOC also impacts the temperature activity coefficient.

However, more studies are still needed to evaluate the impact of DOC characteristics for

temperature activity coefficient.

Overall, the results from the present study confirmed that the performance of BIEX filters, in

terms of DOC removal, is significantly impacted by the temperature. However, the impact is not

as great as that for BAC filters. This is likely because DOC removal in BIEX filters results from both

ion exchange, which is less significantly impacted by temperature (Amini et al., 2018), and

190

biodegradation, which is significantly impacted by temperature (Seger and Rothman, 1996; Urfer,

1997; Terry and Summers, 2018; Moona et al., 2019).

Figure A5 - Rate constants vs temperature. Note that the rate constants are the same as those

listed in table 2 but presented with equation 2 fitted to the data. Confidence intervals were

omitted because of overlapping with data symbols.

A.3.4 Implication on the operation of biofilters for NOM removal

The performance of biofilters for NOM removal can vary throughout the year especially in Nordic

countries where winter temperature is much lower than summer temperature. The low

temperature during winter seasons could bring challenges for the operation of biofilters for NOM

removal. The EBCT could be extended to address the impact of low temperature on biofilters. The

EBCT required to achieve a given extent of DOC removal in a biofilter at a given temperature could

be modeled by combining equations A1 and A2, yielding equation A3.

;<-= = ,-(%!&%#%"&%#)

"&#.((&(#) (A3)

where Ct is the DOC concentration for the biofilter effluent at an operating temperature of T (°C);

C0 is the DOC concentration for biofilter influent; CR is the residual DOC concentration (mg C/L),

and kR is the DOC removal rate constant at the reference temperature of TR (e.g., 20 °C). For the

present study, to consistently achieve an effluent DOC concentration of 2 mg/L (treatment

0

0.1

0.2

0.3

0 5 10 15 20 25

Rate

con

stan

t (m

in-1

)

Temperature (°C)

BIEXBAC

kT= 0.24×1.0429(T-20)

R2= 0.9974

kT= 0.06×1.0675(T-20)

R2 = 0.9720

191

objective based on USEPA, 1999) over range of temperature considered (4 °C-20 °C), the EBCT

required for BIEX filters was estimated to range from 5 min to 9 min. BAC filters could not achieve

the objective of 2 mg/L for any of the conditions investigated.

A.4 Conclusion The present study is the first to systematically study the impact of temperature and empty bed

contact time (EBCT) on biological ion exchange (BIEX) filters for natural organic matter (NOM)

removal with a benchmark of conventional biological activated carbon (BAC) filters. The key

findings are summarized as follows.

• BIEX achieved a higher average dissolved organic matter (DOC) removal (73 ± 6 %) than

did BAC (22 ± 9 %) at steady state with an EBCT of 30 min.

• Higher temperature can significantly improve the DOC removal for BIEX and BAC filters

operated at lower EBCTs (i.e., 7.5 min and 15 min).

• Higher EBCT could significantly improve DOC removal for BIEX and BAC filters operated at

lower temperatures (i.e., 4 °C and 10 °C).

• Temperature had a greater impact on BAC filters than BIEX filters for NOM removal based

on temperature activity coefficients (1.0675 vs.1.0429).

• The impact of temperature and EBCT must be considered in parallel when using drinking

water biofilters for NOM removal.

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