Water Research 111 (2017) 297e317
Contents lists available at ScienceDirect
Water Research
journal homepage: www.elsevier.com/locate/watres
Review
From the conventional biological wastewater treatment to hybrid
processes, the evaluation of organic micropollutant removal: A review
ment a, b, c, Isabelle Seyssiecq b, Anne Piram a,
Camille Grandcle
Pascal Wong-Wah-Chung a, Guillaume Vanot c, Nicolas Tiliacos c, Nicolas Roche b, *,
Pierre Doumenq a
a
b
c
Aix-Marseille Univ, CNRS, LCE, Marseille, France
Aix-Marseille Univ, CNRS, Centrale Marseille, M2P2, Marseille, France
^teau-Gombert, H
^le de Cha
Soci
et
e Seakalia SAS, Groupe Ovalee, Technopo
eliopolis, 13013, Marseille, France
a r t i c l e i n f o
a b s t r a c t
Article history:
Received 17 May 2016
Received in revised form
15 December 2016
Accepted 2 January 2017
Available online 6 January 2017
Because of the recalcitrance of some micropollutants to conventional wastewater treatment systems, the
occurrence of organic micropollutants in water has become a worldwide issue, and an increasing
environmental concern. Their biodegradation during wastewater treatments could be an interesting and
low cost alternative to conventional physical and chemical processes. This paper provides a review of the
organic micropollutants removal efficiency from wastewaters. It analyses different biological processes,
from conventional ones, to new hybrid ones. Micropollutant removals appear to be compound- and
process- dependent, for all investigated processes. The influence of the main physico-chemical parameters is discussed, as well as the removal efficiency of different microorganisms such as bacteria or white
rot fungi, and the role of their specific enzymes. Even though some hybrid processes show promising
micropollutant removals, further studies are needed to optimize these water treatment processes, in
particular in terms of technical and economical competitiveness.
© 2017 Elsevier Ltd. All rights reserved.
Keywords:
Trace organic contaminants removal
Process parameters
Membrane bioreactor
Advanced treatment
Enzymatic treatment
Contents
1.
2.
3.
Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 298
Biodegradation of micropollutants in wastewater treatment plants using classical processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 299
2.1.
Removal of micropollutants in wastewater treatment plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 299
2.1.1.
Micropollutants occurrence in wastewaters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 299
2.1.2.
Conventional activated sludge treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 299
2.1.3.
Membrane bioreactor treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 300
2.2.
Effects of operating conditions on removal efficiency . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 301
2.2.1.
Effects of hydraulic retention time and sludge retention time . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 301
2.2.2.
Effect of the dissolved oxygen concentration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 302
2.2.3.
Effects of pH and temperature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 303
2.3.
Feeding effects on removal efficiency: batch vs continuous . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 304
2.4.
Effects of microorganism communities or enzymes extracted from microorganisms on removal efficiency . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 305
2.4.1.
Activated sludge . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 305
2.4.2.
White-rot fungi . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 305
2.4.3.
Bacteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 308
2.5.
Factors limiting the biodegradation in wastewater treatment plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 309
Areas for improvement . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 309
3.1.
Hybrid process description . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 309
* Corresponding author.
E-mail address: nicolas.roche@univ-amu.fr (N. Roche).
http://dx.doi.org/10.1016/j.watres.2017.01.005
0043-1354/© 2017 Elsevier Ltd. All rights reserved.
298
C. Grandclement et al. / Water Research 111 (2017) 297e317
3.2.
4.
Effects of operating conditions on removal efficiency . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3.2.1.
Effects of HRT and SRT . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3.2.2.
Effect of the DO concentration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3.2.3.
Effects of pH and temperature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3.3.
Effects of microorganism communities or enzymes extracted from microorganisms on removal efficiency . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3.3.1.
Biofilm . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3.3.2.
Activated sludge and suspended biofilm carriers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3.3.3.
Enzymatic treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3.4.
Limits of hybrid processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Conclusions and perspectives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Supplementary data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
1. Introduction
Agriculture, industry and domestic practices around the world
are releasing multiple compounds in wastewater, inducing an
increasing environmental concern about pollutants occurrence in
aquatic environments (Kim et al., 2007; Deblonde et al., 2011).
Emerging pollutants, also called trace organic contaminants
(TrOCs), are compounds present in the environment at trace concentrations and whose effects on the environment and human
health are currently unknown. These contaminants include pharmaceuticals, personal care products, industrial chemicals, pesticides, polycyclic aromatic hydrocarbons (PAH), as well as metallic
trace elements. To date, discharge guidelines and standards do not
exist for most of these compounds. However, the EU water
framework directive 2000/06/CE announces in Annex X a list of 45
priority substances or groups of substances. This list which includes
metals, pesticides, phthalates, PAHs, and endocrine disruptors as
well, imposes the removal of these compounds within an objective
of quality and preservation of the good ecological status of water by
2015, not only in receiving waters but in order to remove ecotoxicity of these compounds. Indeed, because of their persistence,
some organic micropollutants could be toxic and bioaccumulate
with potential significant impacts on human health and the environment. This bioaccumulation is typically associated with the high
lipid solubility property of a compound and its ability to accumulate in the fatty tissues of living organisms for a long time period.
These persistent compound move up the food chain, and they increase in concentration as they are processed and metabolized in
certain tissues of organisms, increasing their toxicity in the environment (Burkhardt-Holm, 2011). Furthermore, a watch list of
substances for European Union-wide monitoring was, recently,
reported in the Decision 2015/495/EU of 20 March 2015, including
two pharmaceuticals (diclofenac (DCF) and the synthetic hormone
17-a-ethinylestradiol (EE2)) and a natural hormone (17-b-estradiol
(E2)), three macrolide antibiotics (azithromycin (AZI), clarithromycin (CLA) and erythromycin (ERY)), other natural hormone
(estrone (E1)), some pesticides (methiocarb, oxadiazon, imidacloprid, thiacloprid, thiamethoxam, clothianidin, acetamiprid and
triallate), a UV filter (2-ethylhexyl-4-methoxycinnamate) and, an
antioxidant (2,6-di-tert-butyl-4-methylphenol) commonly used as
food additive (Barbosa et al., 2016).
The increasing concern about the potential accumulation of
micropollutants in the aquatic environment triggered many investigations about their biological degradation in wastewater
treatment systems (Stackelberg et al., 2007). Some mechanisms
such as adsorption on activated sludge flocs or photolysis have been
studied for the removal of micropollutants during water treatment
processes (Radjenovi
c et al., 2009). However, current wastewater
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311
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311
311
312
313
313
313
313
313
treatment plants (WWTPs) using conventional biological processes
are not specifically designed to eliminate recalcitrant TrOCs. Thus,
due to their persistence, many of these molecules are able to pass
through wastewater biological treatment processes. This recalcitrance has often been linked to their molecular properties, which
define their biodegradation abilities by a given strain of microorganism under given operating conditions (Tahri et al., 2013). For
instance, Kimura et al. (2005) suggested that the presence of
chlorine in the molecular structure, and a relatively complex aromatic structure are the reasons for the low degradation rates
observed in the case of clofibric acid (CFA), dichloprop, and DCF.
Moreover, Tadkaew et al. (2011) examined the relationship between chemical structures and the removal of TrOCs using membrane bioreactors (MBRs). Some physico-chemical properties such
as hydrophobicity and the presence of electron withdrawing
(EWGs), or electron donating functional groups (EDGs) appear to be
important factors governing TrOCs biodegradation. This study
shows high removal efficiency for hydrophobic compounds with a
log Kow> 3.2 (at pH ¼ 8.0) and hydrophilic compounds (log Kow<
3.2) which possess only EDGs such as hydroxyl groups or primary
amine groups. In contrast, the removal of hydrophilic compounds
bearing only EWGs is very low (below 20%). For hydrophilic compounds which have both EDGs and EWGs, their removal rate is
variable depending on their functional groups. Beside biodegradation, adsorption can also govern the removal of TrOCs from the
aqueous phase during MBR or conventional activated sludge (CAS)
treatments. According to Fan et al. (2014), removal efficiencies of
five pharmaceuticals by sludge adsorption were positively correlated with their Kow (namely octanolewater partition coefficients).
The removal of pharmaceuticals by sludge adsorption is mainly
affected by the electrostatic interactions between the molecule and
sludge surface, and the hydrophobic/hydrophilic character of the
molecule.
Many review papers have been published regarding the occurrence and fate of micropollutants in the aquatic environment
ge et al., 2009; Oulton et al., 2010; Deblonde et al., 2011;
(Mie
Lapworth et al., 2012; Wijekoon et al., 2013; Luo et al., 2014b,…).
Most of these studies focused on the removal of micropollutants
through CAS processes or MBR treatments (Clara et al., 2005b;
Clouzot et al., 2008), but only a few of them have treated the
removal of micropollutants using recently developed advanced
processes, such as adsorption processes, advanced oxidation processes, or membrane processes (Oulton et al., 2010; Luo et al.,
2014b; Ahmed et al., 2016). Besides, no attempt has been made to
provide a comprehensive review of the removal of contaminants
using hybrid processes, combining different technologies, such as
fixed and free biomasses for instance, or a comparison between
processes using different microorganism's strains. In this context,
C. Grandclement et al. / Water Research 111 (2017) 297e317
the aim of this work is to review the performance of different
processes regarding the removal of emerging contaminants.
Several processes from classical ones (CAS treatment, MBR treatment), to more original approaches such as fixed-bed bioreactors or
hybrid processes will be studied among different scales from laboratory pilot plants to real WWTP. The type of microorganisms used
to efficiently degrade these micropollutants is also reviewed, as
well as the effects of operating conditions on removal efficiencies.
2. Biodegradation of micropollutants in wastewater
treatment plants using classical processes
2.1. Removal of micropollutants in wastewater treatment plants
2.1.1. Micropollutants occurrence in wastewaters
Several review papers report the occurrence of micropollutants
in different water bodies such as influent and effluent from WWTPs
ge et al., 2009; Deblonde et al., 2011; Verlicchi et al., 2012;
(Mie
Benner et al., 2013; Luo et al., 2014b; Evgenidou et al., 2015, …),
but also groundwaters (Lapworth et al., 2012; Luo et al., 2014b; Sui
et al., 2015), surface waters (Luo et al., 2014b), or seawater (ArpinPont et al., 2014). The occurrence and repartition of TrOCs, especially pharmaceuticals, in sewage water and sludge flocs along with
conventional activated sludge, or MBR wastewater treatment processes, have been studied and these compounds are generally
found in concentrations ranging from low ng.L 1 to a few mg.L 1 in
the liquid phase, as well as in solid phase: from a few ng.g 1 to a
few mg.g 1 in sewage sludge (Jeli
c et al., 2011; Verlicchi et al., 2012;
Jiang et al., 2013). For instance, the study of 78 peer-reviewed papers has showed that analgesics and non-steroidal anti-inflammatory drugs (NSAID) concentrations are ranging from 1.60 ng.L 1
to 373 mg.L 1 in the raw influent of municipal WWTPs.
The most commonly investigated compounds were ibuprofen
(IBP), DCF, naproxen (NPX), ketoprofen (KPF) and acetaminophen
(ACE). Regarding antibiotics, variability of their concentrations was
found between 1.0 ng.L 1 and 32 mg.L 1 in the raw influent to
municipal WWTPs, and the most commonly investigated compounds were trimethoprim (TMP), sulfamethoxazole (SMX), ERY,
and ciprofloxacin (CIP) (Verlicchi et al., 2012). In addition, other
review papers, such as Bolong et al. (2009), focused on physical,
biological, or chemical treatment methods for endocrine disrupting
compounds and other pharmaceuticals.
The major part of micropollutants comes from several sources
like domestic or industrial wastewater, hospital effluents, or agricultural run-off (Luo et al., 2014b). Even though the discharge from
WWTPs is only one of the pathways for the introduction of
micropollutants to surface water, WWTPs act as primary barriers
against their spread. Indeed, non-negligible removal rates (from 13
to 100% for some compounds such as atrazine (ATZ), DCF, triclosan
(TCS), estriol (E3)…) have been observed in WWTP's effluents of 14
different countries where they are commonly present in wastewaters at trace concentrations, ranging from a few ng.L 1 to several
mg.L 1 (Luo et al., 2014b).
TrOCs removal is generally dependent on compound physicochemical properties, process-specific factors such as sludge retention time (SRT), or hydraulic retention time (HRT) as well as seasonal parameters such as temperature, precipitation rate, and solar
radiation (Vieno et al., 2005). According to Luo et al. (2014b), a firm
conclusion about the persistency of each compound cannot be
easily drawn, as many compounds showed significantly different
removal rates in different conventional WWTPs. Nevertheless, the
authors presented a simple classification for the removal rates of
these compounds in conventional WWTPs. For instance, ATZ,
diazinon (DZN), DCF, carbamazepine (CBZ), metoprolol (METOP), as
well as mefenamic acid (MFA) are, on the average, removed with
299
poor rates (<40%), while bisphenol A (BPA), caffeine (CFN), IBP, E2,
E1, NPX, nonylphenol (NP), TCS are generally removed with high
rates (>70%).
Some reliable complementary processes can help improving
micropollutant removal. For example, ozonation process can highly
remove molecules such as DCF, CBZ, TCS, E1 (>90%) (Sui et al.,
2010). Nevertheless, this process implies important operating
costs due to high energy requirements. Some by-products, such as
bromate, can also be produced by this treatment from the oxidation
of bromide, through a combination of ozone and OH radical reactions (Von Gunten, 2003). Coagulation and flocculation processes
have also been tested, but most of the time they did not show any
significant removal efficiencies of the tested micropollutants,
whereas activated carbon adsorption can allow important removal
efficiencies especially for hydrophobic compounds with a log Kow >
4. On the contrary, according to Rogers (1996) compounds with log
Kow< 2.5 have a low sorption potential on activated carbon. Using
this process, DCF and CBZ can be removed with efficiencies higher
than 90% (Grover et al., 2011; Kovalova et al., 2013). Besides, organic
micropollutants in aqueous solution will partially be in their ionic
form at a given pH (depending on their pKa), and as log D is pHdependent, log D values are important factors to take into account in the removal by sorption of such micropollutants Tadkaew
et al. (2010).
However, the maintenance cost of adsorption processes is not
negligible. Indeed, granulated activated carbon-based removal
technology will become less efficient over time as the adsorption
bed ages and adsorption sites become less and less regenerated.
Furthermore, micropollutants can be released back into the solution when the influent concentration of a contaminant drops, in
order to restore equilibrium or in case of competition between
organic compounds and other adsorbed species. Bourneuf et al.
(2015) studied the desorption phenomenon of micropollutants
onto activated carbon in water phase. The results showed that
several cycles of adsorption and desorption of methyldiethanolamine and 2,4-dimethylphenol could be successively run on a
column of fixed-bed adsorbent, and that attenuation is largely
dependent on the contaminant (chemical structure of the
pollutant, and notably the occurrence of aromatic moieties, or their
hydrophobicity).
2.1.2. Conventional activated sludge treatment
The activated sludge treatment is commonly used in municipal
WWTPs. It involves the addition of pretreated wastewater and
microorganisms to remove nutrients, and to oxidize carbonaceous
biological matter and nitrogenous matter, mainly ammonium and
nitrogen. The process begins by mixing the polluted influent from
industrial or sewage wastewater with an aerobic bacterial culture
in an aerated reactor (Eckenfelder and Cleary, 2014). The aeration
tank retention time is then adjusted to ensure that the effluent is
sufficiently treated before undergoing a solid/liquid separation in a
gravimetric clarifier (Tchobanoglous et al., 2003). The collected
settled activated sludge is then mainly recycled back to the aeration
tank in order to maintain a fixed concentration of depolluting microorganisms (Eckenfelder and Cleary, 2014).
Three main pathways of degradation exist during activated
sludge treatment: microbial processes (biodegradation, either
metabolic, or co-metabolic), sorption onto sludge flocs, and volatilization (mainly during aeration). However, volatilization can be
considered negligible for the majority of pharmaceuticals and
personal care products (PPCPs), because of the Henry's constant
value of such molecules (Joss et al., 2006). Adsorption onto activated sludge flocs could be a significant pathway for some compounds such as musk fragrances in CAS, or estrogens in MBR due to
the hydrophobicity of such compounds (Carballa et al., 2005;
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C. Grandclement et al. / Water Research 111 (2017) 297e317
Clouzot et al., 2010; Maeng et al., 2013). Besides, organic micropollutants in aqueous solution will partially be in their ionic form at
a given pH (depending on their pKa), and as log D is pH-dependent,
log D values are important factors to take into account in the
removal by sorption of such micropollutants. As said previously,
Tadkaew et al. (2010), as well as Wells (2006), showed that log D is
pH-dependent and suggested that the sorption of a TrOC onto
activated sludge flocs could be assessed by considering the log D
value of the compound at a given pH.
Regarding biodegradation, according to some authors, cometabolic biodegradations could play a major role on the removal
mechanism of micropollutants during activated sludge treatment
of municipal wastewater, since the concentrations in micropollutants could be too low to serve as a direct growth substrate
(Quintana et al., 2005; Xue et al., 2010; Fischer and Majewsky,
2014). In this aerated tank occurs the nitrification which leads to
the conversion of ammonia to nitrates thanks to nitrifying microorganisms. These organisms, such as ammonia-oxidizing bacteria
could possibly co-metabolically oxidize micropollutants thanks to
the presence of an ammonia monooxygenase, and thus improve the
removal of organic micropollutants (Margot et al., 2016).
For PPCPs, the removal of these molecules occurs thanks to a
combination of biodegradation and sorption pathways. Hence,
conventional activated sludge systems give rise to a wide range of
removal efficiencies regarding PPCPs (Verlicchi et al., 2012). Among
the PPCPs and some of their human metabolites, ACE, CFN, digoxigenin, E1, IBP, NPX, and paraxanthine are for instance rather well
removed (>90%); whereas CBZ, EDTA, MFA, gemfibrozil (GFZ), CIP,
E3, ofloxacin (OFX), penicillin V, SMX, TMP are poorly removed
from influent (Bernhard et al., 2006; Radjenovic et al., 2009; Blair
et al., 2015). It is worth noting that the degradation of some
PPCPs like CFN, ACE, and metformin that are highly degradable,
slowed or stopped at trace, but notable, concentrations within an
activated sludge system. A degradation plateau has been observed
with these molecules: 40 ng.L 1 for CFN, 90 ng.L 1 for ACE, and
1000 ng.L 1 for metformin (Blair et al., 2015). This phenomenon
may explain, according to the authors, the continuous low levels of
degradable PPCPs in the effluents of WWTPs. Furthermore, this
study revealed negative mass balances for some PPCPs, such as CBZ
or OFX, whose both soluble and sorbed concentrations increased
over time during aerobic batch experiments (Blair et al., 2015). It
has been hypothesized that some PPCPs can be enclosed in fecal
particles and then, released to the liquid phase when the feces are
broken down by microorganisms (Gobel et al., 2007). Another potential theory is that the undetected PPCPs metabolites are further
transformed back into the parent compounds through microbial
activity (Verlicchi et al., 2012). The deconjugation of conjugates by
hydrolysis during treatment, yielding the parent compound, could
arez et al.,
lead to an additional source of contaminant load (Su
2008; Kovalova et al., 2013). According to Blair et al. (2015), the
negative mass balances are due to a combination of all of these
processes, with the driving factor being compound specific. DCF
removal efficiencies ranging from 0% to 70% have been reported
according to the biological composition of the sludge used (Clara
et al., 2005b; Bernhard et al., 2006; De Wever et al., 2007;
Kimura et al., 2007; Radjenovic et al., 2009). Between 50 and 65%
removal of KPF and NPX have been found in previous studies
(Carballa et al., 2004; Quintana et al., 2005; Radjenovi
c et al., 2009).
A complete removal of NPX was even reported in one study during
a process of wastewater treatments involving disinfection
(Metcalfe et al., 2003). Quintana et al. (2005) used a sludge which
was withdrawn from a reactor treating real municipal wastewater
in which all five selected pharmaceuticals such as KPF, DCF, bezafibrate (BZF), NPX, and IBP were found (Quintana and Reemtsma,
2004). The results of this study showed that KPF could serve as
sole substrate for the microbial growth, which could explain its
high biodegradability, whereas a cometabolic transformation
appeared to be, generally, the important biodegradation pathway in
the case of acidic pharmaceuticals. TMP was removed with around
40% efficiency. This compound is generally considered recalcitrant,
rez et al. (2005) observed its degradation using slow-growing
but Pe
nitrifying bacteria. Some antibiotics such as AZI, ERY, OFX, or TMP
are expected to sorb onto negatively charged surface of sludge flocs
through ionic interactions (Radjenovic et al., 2009).
More generally, the observed important differences between
removal efficiencies for a given molecule from one work to another
are probably due to the differences in operating parameters of the
compared CAS systems, such as the SRT, the HRT, or the solid phase
concentration, but also to the biological composition of sludge flocs
and the chemical composition of wastewaters.
2.1.3. Membrane bioreactor treatment
MBR is another type of common technology for biological
wastewater treatment in which activated sludge treatment is
directly combined to a membrane separation process. It presents
several advantages such as low space requirement and high
effluent quality. Membranes allow a complete retention of particulate matter, but also work at higher solid concentrations without
limitation due to the subsequent solid/liquid separation
(Wisniewski, 2007).
Several observations have been reported on the removal of
micropollutants by MBR treatment. In the case of compounds with
an intermediate removal in CAS treatments (between 15 and 80%),
MBR treatments can generally further reduce micropollutant concentrations by 20e50%. However, in the case of compounds which
are already highly degraded by CAS processes or in the case of
recalcitrant compounds, the results using MBRs did not show any
significant improvements (Hai et al., 2010). Some authors also
concluded that removal rates in MBRs and CAS processes are
comparable for selected pharmaceuticals, fragrances, endocrine
disrupting compounds, naphthalene sulfonates, and benzothiazole-2-sulfonate (Clara et al., 2005b; Joss et al., 2005). On the
contrary, Bernhard et al. (2006) showed significantly better
removal rates of studied persistent polar pollutants such as DCF,
mecoprop, and sulfophenyl carboxylates with MBRs compared to
CAS systems, whereas recalcitrant micropollutants such as EDTA
and CBZ were not eliminated at all during wastewater treatments
by these processes. A better removal efficiency for NP and nonylphenol ethoxylates (NPEO) using a MBR compared to a CAS syslez et al. (2006). Similarly, the two
tem has been noticed by Gonza
MBRs used by Kimura et al. (2007) exhibited better elimination
rates for the six selected acidic pharmaceuticals than the reference
activated sludge process. Kim et al. (2007) also observed that a MBR
system seems to be efficient for hormones (e.g. E3, testosterone,
androstene-dione) and some pharmaceuticals (e.g. ACE, IBP, and
CFN) with approximately 99% removal, but that this process did not
decrease the exit concentration for molecules such as ERY, TMP,
NPX, DCF, and CBZ. Concerning the adsorption phenomenon,
Radjenovi
c et al. (2009) found higher concentrations in MBR sludge
flocs than in CAS sludge flocs for hydrochlorothiazide, AZI, CBZ, and
KPF which could either be explained by a modified intrinsic hydrophobicity (e.g. aliphatic and aromatic groups), an increase of
surface area, or increased electrostatic interactions (e.g. amino
groups) with MBR sludge flocs (Kim et al., 2007). MBR could also
improve the degradation of TrOCs because it allows reaching
different values for process parameters such as HRT or SRT,
compared to CAS system. Because MBRs generally operate at higher
SRTs (at least 15 days) than CAS systems (at most 15 days), higher
removal efficiencies can be achieved as reported by Clara et al.
(2005a), Radjenovi
c et al. (2009), and Weiss and Reemtsma
C. Grandclement et al. / Water Research 111 (2017) 297e317
(2008). However, the relationship between some process parameters is still unclear. A comparison between CAS systems and a MBR
operating at comparable SRT showed no significant differences in
the treatment efficiency (Clara et al., 2005a).
According to Cirja et al. (2008), the solid phase properties also
varies in MBRs compared to CAS systems, both as a function of
wastewater composition and operating conditions, in particular
hydrodynamics, through an increase of the average shear rate.
Indeed, hydrodynamic stress in MBRs reduces floc size which is also
dependent upon mixed liquor suspended solids or exopolymeric
substances concentrations (Zhang et al., 1997). Smaller flocs
(10e100 mm in MBRs against 100e500 mm in CAS systems) and the
presence of some free-living bacteria in MBRs could improve masstransfer kinetics, and thus elimination efficiencies with this process. Indeed, MBRs typically run at lower food/microorganisms (F/
M) ratio than CAS process in order to mitigate membrane fouling
and maintain high oxygen transfer efficiency. The F/M ratio, which
is a balance between substrate consumption and biomass generation, determines the degree of decomposition of organic matter,
and the removal of micropollutants. However, it is hard to
demonstrate that only a low F/M ratio encourages micropollutant
biotransformation, other parameters such as HRT may come into
play (Petrie et al., 2014). Thanks to the presence of smaller flocs and
free-living bacteria, the biomass in a MBR also seems to have a
more viable fraction compared to that of a CAS system (Cicek et al.,
1999). Finally, specific floc surface per unit of reactor volume was
ten times higher in MBRs than in CAS systems. As a consequence,
the contact between microorganisms and pollutants could be
favored with MBRs, which could stimulate enzymatic activities.
Indeed, part of the enzymatic activity seems to increase proportionally with the specific surface area of contact between suspended biosolids and polluted waters (Cirja et al., 2008).
Finally, MBR is able to deliver lower and more stable effluent
concentrations in comparison to CAS systems, generally under
lower HRT, as far as compounds with moderate removals in CAS
systems are concerned (including NPX, DCF, phenazone, CFA).
However, this effect is not important enough to serve as a financial
argument for developing the use of MBRs in municipal WWTPs,
according to Weiss and Reemtsma (2008).
2.2. Effects of operating conditions on removal efficiency
Table 1 (Appendix A: supplementary data) reports examples of
selected micropollutant removal using classical bioreactors (batch
experiments or MBR systems) from an analysis of literature data.
The operating conditions such as temperature, HRT, or SRT of each
study are also compiled in this table. Removal efficiencies given in
Table 1 are grouped according to their references, and represent the
total removal of the species from the liquid phase, so they include
the contributions of biodegradation (both metabolic and cometabolic) and/or adsorption onto activated sludge flocs.
2.2.1. Effects of hydraulic retention time and sludge retention time
The sludge retention time, also known as solid retention time or
sludge age, indicates the mean residence time of microorganisms in
the reactor and is related to the growth rate of microorganisms. It is
calculated through the ratio of the tank volume compared to the
sludge volumetric removal flow rate. High SRTs allow an enrichment of the biomass in slowly growing autotrophic bacteria such as
nitrifiers which can also excrete enzymes that can possibly break
down some low degradable molecules with aromatic rings
(Rosenberger et al., 2002; Cirja et al., 2008). Monod-type kinetics
deal with the relationship between the growth rate of a microbial
species and the concentration of a critical substance sustaining its
growth. Compounds must be sufficiently easy to degrade, and it
301
also must be available in sufficient amounts to result in significant
energy and/or biomass recovery. The degradation of a certain
amount of pollutants enables a proportional enhancement in microbial biomass. On the assumption that the biodegradation of a
given micropollutant is described by a Monod kinetic, a specific SRT
can be associated to this substance even at low concentration, or in
the case of a co-metabolism. Indeed, in a process using biomass
recirculation, the installed SRT corresponds approximately to the
reciprocal of the growth rate (Clara et al., 2005a). Considering this
relationship, according to Clara et al. (2005a), the effluent concentration of some organic micropollutants is dependent on the
selected/operated SRT and independent of influent concentrations.
This is why SRT is a fundamental parameter to design a WWT
process. For example, a minimum value of 10e15 days for the SRT
was proposed by Clara et al. (2005a). Micropollutants can be only
degraded from a critical SRT value, which are determined for
different compounds. If a WWTP operates with SRTs below this
critical value, effluent concentrations of micropollutants are expected to be in the range of influent concentrations. This concept is
useful to allow an estimation of outlet concentrations, and for the
design of WWTP to enhance the removal of organic micropollutants such as pharmaceutical active compounds (PhACs) and
the nitrification process along biological wastewater treatment
systems (Kreuzinger et al., 2004). For instance, it has sometimes
been reported that an increase in SRT could enhance the eliminac et al., 2012b). Indeed, in a MBR,
tion of some pharmaceuticals (Jeli
higher biomass concentration and the presence of slower growing
species, both resulting from higher SRTs, have led to higher removal
efficiencies of some PPCPs, as revealed by Table 1 (FernandezFontaina et al., 2012). According to Xia et al. (2012), higher SRTs
(above 30 days) correspond to the suitable operational condition
for sufficient antibiotics removal (up to 80%). For Tambosi et al.
(2010), a MBR with a SRT of 30 days presented higher removal efficiencies than a MBR with a SRT of 15 days for all tested pharmaceutical compounds. The same observation has been made by
Kimura et al. (2007). In their study, the MBR with the higher SRT
exhibited the best performances for the removal of pharmaceuticals such as CFA, DCF, KPF, and NPX, as it is collected in Table 1. For
instance, for a SRT of 65 days, the removal efficiencies of KPF and
CFA achieved 99% and 82% respectively, whereas for a SRT of 15 days
the degradation was about 83% and 50% respectively. Moreover, low
effluent concentrations can be achieved in WWTPs operated at
SRTs higher than 10 days, in particular for the biodegradation of
hormones, BZF, and IBP (Clara et al., 2005a). Besides, even if its
influence on the removal of PPCPs has scarcely been reported,
acclimation of biomass is known to be beneficial for degradation of
rez et al., 2012).
xenobiotics (Sua
However, the correlation between the removal rate and the SRT
was not straightforward. Some authors such as Joss et al. (2005)
and Vieno et al. (2007) reported that the effect of an increase in
SRT is not clear and may vary significantly depending on the tested
compounds. Falås et al. (2016) supports this idea, observing no
strong and systematic correlation between the SRT and the rate
constants of more than 20 micropollutants with SRTs ranging from
25 to 80 days. The removal of some pharmaceuticals such as CBZ,
DCF, or ACE during biological treatments did not show any significant dependency on SRT. Regarding CBZ, Bernhard et al. (2006) and
Maeng et al. (2013) have observed that this molecule still remain
recalcitrant regardless of the change of SRT using a MBR reactor.
Concerning DCF, even if most of the studies reported higher elimrez
ination at higher SRTs using MBRs, Clara et al. (2005b) and Sua
et al. (2012) noted no correlation. Besides, Bernhard et al. (2006)
have observed that an enhanced DCF elimination (reaching a
plateau) was obtained at higher SRTs using a MBR reactor. DCF
removal rate was 8e38% when SRT was 20e48 days, 59% at a SRT of
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C. Grandclement et al. / Water Research 111 (2017) 297e317
62 days, and 53% at a SRT of 322 days. Additionally, hydrophilicneutral pharmaceuticals (based on log Kow and pKa values) such
as CFN, phenacetine, or ACE, and hydrophilic-ionic pharmaceuticals
as IBP and estrogens (E1, E2, EE2) can be removed by a MBR operated at a SRT as small as 8 days (up to 90%). Conversely, other
compounds such as KPF, CFA, and EE2 need a higher SRT from 20 to
80 days to be correctly were removed (removal efficiencies achieved 65e90%, 6e34% and 71e78% respectively, as collected in
Table 1) (Maeng et al., 2013). Finally, some studies have found the
SRT to be a determining factor as far as biodegradation kinetics of
micropollutants are concerned. Majewsky et al. (2011) compared
biodegradation kinetics of some pharmaceuticals such as CFN, DCF,
CBZ, ACE, as well as SMX, using activated sludge from two WWTPs
notably differing by their SRT. The results, collected in Table 1,
showed that PhAC removal was more important under high concentration of heterotrophic microorganisms at a low SRT. Besides,
according to Sipma et al. (2010), the biodegradation of some
micropollutants is mostly due to co-metabolism processes since
their low concentrations are not likely to sustain microorganisms
growth. Since SRT is the relevant parameter to achieve an efficient
biodegradation of the primary substrate, this could explain the fact
that an increase in SRT beyond 30 days does not seem to give any
improvement in terms of removal efficiencies of different compounds (Sipma et al., 2010). For example, an increase in SRT appears
to be a relevant parameter for an efficient biodegradation in the
case of very low concentrated compounds such as pharmaceuticals
(concentration range from ng.L 1 to a few mg.L 1) (Sipma et al.,
2010). Gobel et al. (2007) demonstrated that the combination of a
high SRT with reduced F/M ratios may induce an increased biodiversity, and thus enhance elimination of compounds such as TMP,
and CLA by co-metabolism processes.
The hydraulic retention time (HRT) corresponds to the mean
residence time of the liquid phase in the reactor. This parameter has
an impact on the reaction volume and on the F/M ratio, but not on
the Kbiol coefficient of the compound. However, even with a lack of
information about temperature and sludge age, the influence of this
parameter on the biodegradation efficiencies of different micropollutants was largely investigated. Bernhard et al. (2006) found no
significant correlation between the removal of micropollutants
such as pharmaceuticals and the HRT in a MBR, but noticed that the
tested MBR showed better removal efficiencies (even if its HRT was
lower) than a 22 h HRT reference WWTP. Vieno et al. (2007) also
noticed that the relationship between HRT and removal efficiency
was not straightforward for all selected compounds, sampled in
different WWTPs in Finland, having a SRT between 2 and 20 days.
They observed that a decrease of the HRT reduces the elimination
for some beta-blockers such as METOP, and atenolol, but the effect
was not so evident for sotalol. On the contrary, no significant effects
were found by Weiss and Reemtsma (2008), who studied the
variation of HRT in the range of 7 he14 h on different TrOCs
removal rates, using a MBR. Weiss and Reemtsma (2008) assumed
that a combination of a high SRT and a reduced F/M ratio at low
HRT, which may force microorganisms to utilize poorly degradable
polar compounds as substrates, induces an increased biodiversity
in MBRs. Indeed, a lower F/M ratio results in stronger substrate
limitation. This could explain why removal efficiencies of some
persistent PhACs are higher in MBRs operated under such feeding
conditions than in CAS systems, and why this can be obtained even
under low HRT (Weiss and Reemtsma, 2008).
Gros et al. (2010) calculated PhACs removal efficiencies in
Spanish WWTPs (SRT data are unknown) and the corresponding
PhAC half-life times t1/2, assuming that compound degradation
followed a pseudo-first order kinetic. Indeed, they assumed that the
decrease of the concentration through time is proportional to the
concentration remaining in the matrix used. On the one hand, Gros
et al. (2010) concluded that degradation kinetics of compounds
with high pseudo-first order biological degradation rate constants
(Kbiol) (or low t1/2) and low log Kow (low sorption abilities) are more
influenced by HRT, while degradations kinetics of compounds with
low Kbiol and high log Kow are more influenced by SRT. On the other
hand, there are some exceptions such as IBP which is a high Kbiol
and low log Kow molecule that remains well removed whatever the
HRT and SRT values are (Gros et al., 2010). Besides, the HRT value
does not influence removal efficiencies for compounds with high t1/
2 like CBZ always showing poor or no elimination, whereas for
compounds with medium t1/2 (between 10 and 20 h), the HRT value
seems to play a role on the achieved percentage of degradation
(Gros et al., 2010). Joss et al. (2006) also observed a pseudo firstorder degradation kinetics for many organic micropollutants
down to ng.L 1 concentrations, indicating that their biodegradation
is directly influenced by micropollutant concentration. As a
consequence of this pseudo first-order kinetic, micropollutant
concentration decreases exponentially with time with a constant
directly related to Kbiol so that the effects of operating conditions
are
less
obvious
for
low
degradable
compounds
(Kbiol < 0.1 L.gSS1 .d 1) as well as highly degradable compounds
(Kbiol > 10L.gSS1 .d 1). They concluded by proposing three groups of
micropollutants:
- Kbiol < 0.1 L.gSS1 .d 1: no substantial removal by degradation
(<20%), but for strongly sorbing compounds with Kd > 1 L.gSS1
the removal may be higher due to transfer to sludge.
- 0.1 < Kbiol < 10 L.gSS1 .d 1: partial removal (20e90%)
- Kbiol > 10 L.gSS1 .d 1: more than 90% removal by biological
degradation; specific degradation efficiency strongly dependent
on reactor configuration.
If Kbiol for a micropollutant is known, the HRT could be adjusted
to ensure efficient removal of this compound. However, this makes
sense in very specific contexts where one or a few pollutants are of
particular concern, such as in industrial wastewaters. Indeed, HRT
cannot be increased to the extent to remove some of the very
recalcitrant compounds in municipal WWTP.
2.2.2. Effect of the dissolved oxygen concentration
Biodegradation experiments described in the literature were
mostly carried out under aerobic conditions. It is indeed known
that some ammonia-oxidizing microorganisms such as nitrifying
microorganisms, whose growth is favored under high dissolved
oxygen (DO) environments, have the potential to degrade some
TrOCs (Ren et al., 2007).
Since conventional WWTPs combine the existence of aerobic
and anoxic conditions, and since different metabolites could be
formed under these conditions, it is interesting to investigate the
potential removal mechanism of various pollutants under different
redox conditions. For instance, for compounds with amide groups,
the first transformation step of primary and secondary amides is
usually a hydrolysis of the amide group, while the primary transformation step of tertiary amides is an oxidation. Hydrolysis of
primary and secondary amides can occur under both oxic and
anoxic conditions, whereas oxidation of tertiary amides specifically
requires the presence of molecular oxygen (Helbling et al., 2010). In
rez et al. (2010) studied the removal of some PPCPs
this context, Sua
under both nitrifying and denitrifying conditions. Under nitrifying
conditions, aerobic bacteria using inorganic chemicals as an energy
source were found, whereas anaerobic or heterotrophic facultative
anaerobic bacteria formed a denitrifying biomass under anoxic or
anaerobic conditions. They observed an increase of DCF removal
from 0% to 74% in an aerobic reactor due to the development of the
nitrifying biomass, while an efficient aerobic (95%) and anoxic
C. Grandclement et al. / Water Research 111 (2017) 297e317
transformation of IBP (75%) was observed after an acclimatization
period. Oxygen may also directly participate in biochemical reactions, or play a role by regulating the enzymatic activity. Xue et al.
(2010) reported that the first-order biodegradation rate constants
were positively related to the DO level for most of the studied
compounds. The DO concentration level may, as a consequence, be
crucial in promoting the overall degradation.
Lahti and Oikari (2011) also compared removal efficiencies of
micropollutants under both aerobic and anaerobic conditions.
Biotransformation of NPX and, to a lesser degree, bisoprolol (BSP)
was observed under both aerobic and anaerobic environmental
conditions. The biotransformation using inocula from activated
sludge processes achieved about 40% in aerobic and 97.3% in
anaerobic conditions for NPX after 75 and 161 days respectively,
and about 35% in aerobic and 14% in anaerobic conditions for BSP
rez et al. (2010) indicated
after 75 and 161 days respectively. Sua
that fluoxetine, natural estrogens (E1, E2, EE2), and musk fragrances, galaxolide (HHCB), tonalide (AHTN), and celestolide
(ADBI), were transformed to a large extent under both dissolved
oxygen conditions (aerobic (>75%) and anoxic (>65%) conditions).
However, NPX, EE2, roxithromycin (ROX), and ERY were only
significantly transformed in the aerobic reactor (>80%). The transformation rate of BSP, especially anaerobically, was slow, but rose
immediately under aerobic conditions. DCF was recalcitrant under
both aerobic and anaerobic conditions (Lahti and Oikari, 2011).
Some other compounds, such as CBZ, diazepam (DZP), SMX, and
TMP, also showed high resistance to biological transformation
rez et al., 2010), whatever the DO concentration.
(Sua
Moreover, Falås et al. (2016) noticed that many micropollutants
such as atenolol, and BZF are almost ubiquitously degraded under
aerobic treatment systems, whereas TMP, DCF, and DIU seem to be
degraded under specific aerobic treatment processes. On the contrary, demethylation and deiodination of some micropollutants
with high aerobic persistence, such as venlafaxine, or diatrizoate,
can be achieved under anaerobic conditions. Thus, a combination of
different aerobic and anaerobic treatment conditions could expand
the spectrum of organic micropollutants susceptible to biological
degradation at WWTPs.
Regarding pesticides, Stasinakis et al. (2009) investigated the
impacts of aerobic and anaerobic conditions on diuron (DIU)
degradation, using activated sludge reactors. The results showed
that, under aerobic conditions, DIU could be biodegraded by activated sludge (Table 1) and that the role of sorption onto biomass
was not significant, while under anoxic conditions DIU seems to act
as a source of carbon and energy for the microorganisms used in
this study. Besides, the degradation of DIU was enhanced by
acclimatization of the biomass under anoxic conditions. Almost
50% of DIU was degraded after a 140 h batch experiment.
2.2.3. Effects of pH and temperature
The pH of an aqueous body can influence both the solubility of
micropollutants present in this environment and the activity of
microorganisms, in particular the microbial enzymatic activities.
Alterations in pH can inactivate some microbial enzymes that are
essential to complex molecules biodegradation. It can also denature
proteins within the cells, thus preventing microbial activity from
occurring (Sylvia, 2005). Consequently, the fate of micropollutants
during bioreactor treatments can be affected by pH variations.
Chemical, physical, or biological processes involving such micropollutant molecules can show some changes, notably in terms of
their kinetics, depending on the pH value (Cirja et al., 2008). Indeed,
depending on their pKa values, PPCPs can exist in various protonation states as a consequence of pH variations. At pH 6.0e7.0, some
micropollutants are deprotonated and adsorption sludge becomes
an important removal mechanism. Besides pH values varied from
303
neutral to acidic in MBR as nitrification became significant, which
improve the degradation of some pharmaceuticals such as KPF or
IBP (Cirja et al., 2008). Urase et al. (2005) reported a considerable
enhancement in removal efficiency of some TrOCs when MBRs
were operated under acidic (pH ¼ 4.3e5.0) rather than basic conditions (pH ¼ 7.5e8.0) (see Table 1). Higher removal of acidic
pharmaceuticals was achieved under low pH conditions due to an
increase of their adsorption onto sludge particles. According to
Tadkaew et al. (2010), who studied the effects of pH variations
between 5.0 and 9.0 on the removal of different TrOCs, removal
efficiencies of acidic pharmaceuticals such as DCF, KPF, or IBP by
submerged MBR are strongly pH-dependent. The pKa values of
these three compounds are ranging from 4.2 to 4.4. That is why at
pH ¼ 5.0 these molecules are predominantly present as deprotonated species. Consequently, they can readily adsorb onto the
activated sludge flocs improving the MBRs removal efficiencies of
these compounds by adsorption. However, the removal mechanisms are quite different for ionizable and non-ionizable compounds. Indeed, the removal efficiencies of BPA and CBZ remained
relatively constant and independent of the mixed liquor. High
removal efficiency of BPA could be attributed to both high biodegradability and adsorption, while CBZ does not readily adsorb onto
sludge flocs pH (Tadkaew et al., 2010). Moreover, Gulde et al. (2014)
investigated the influence of pH on the biotransformation of 15
micropollutants with cationic-neutral speciation in batch experiment using activated sludge. One control micropollutant with
neutral-anionic speciation, and two neutral micropollutants at pHs
6.0, 7.0, and 8.0 were also performed in same operating conditions.
The authors noticed that biotransformation was pH-dependent and
correlated qualitatively with the neutral fraction of the ionizable
micropollutants. At the same time, they observed that the sorption
coefficients derived from control experiments were small and
showed no notable pH-dependence. They concluded that, pHdependent removal of polar, ionizable organic micropollutants in
activated sludge systems is less likely an effect of pH-dependent
sorption but rather of pH-dependent biotransformation (Gulde
et al., 2014). Furthermore, in the case of MBR using white rot
fungi such as Trametes versicolor, the pH of the medium was found
to be the most important factor, followed by the initial substrate
concentration (Tavares et al., 2006). The optimal pH for T. versicolor
activity was shown to be acidic (pH ¼ 4.5). However, it should be
taken into account for a good pH regulation that the addition of
carbon and nitrogen sources, aiming at boosting the enzymatic
activity (production of laccase), also results in pH variations. Zhang
and Geißen (2012) showed a relationship between a pH decrease
and an increase in the activity of acidogenic bacteria present in a
non-sterile wastewater. Other authors observed a pH decrease
during the growth under carbon consumption of T. versicolor.
Another parameter that can influence the degradation of
micropollutants is the temperature. However, only a few studies
have investigated the effects of temperature variations on the
performances of wastewater treatment processes in the case of
micropollutants. Temperature fluctuations can arise from hot industrial effluents mixed with municipal wastewaters, or diurnal
and seasonal variations, and affect treatment performances. Temperature fluctuations can play a role on microbial activity, solubility,
other physicochemical properties of micropollutant molecules, and
on the reaction rate which can be expressed by the Arrhenius
equation. On the one hand, an increase of temperature of the
effluent can decrease DO concentration, and encourage the development of specific microorganisms. Temperature upshifts (from
35 C to 45 C) are known to cause an increase in suspended solid
levels in the effluents, caused by sludge defloculation and a
decrease (up to 20%) of chemical oxygen demand. On the other
hand, temperature upshifts (from 35 C to 45 C) and periodic
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C. Grandclement et al. / Water Research 111 (2017) 297e317
temperature oscillations (from 31.5 C to 40 C, 6 day period, for 30
days) caused the decrease in bioflocs ability to settle, due to filamentous bacteria proliferation (Morgan-Sagastume and Allen,
2003). Moreover, temperature modifications can also impact
other phenomenon such as membrane fouling (Zhang et al., 2006;
Hai et al., 2011) but the link between temperature variations and
membrane fouling is not very clear. Vieno et al. (2005) observed the
effects of seasonal variations on the remaining concentration of
different pharmaceuticals in effluent waters: the total concentration of all studied pharmaceuticals was 3e5 times higher in
wintertime than during other seasons. Even though the inlet PhAC
concentrations are higher during wintertime than during summertime because of an important consumption of antibiotics for
instance, a slowdown of microbial activity was also observed during wintertime. However, because of the absence of controlled
experimental conditions, the overall effects of temperature variations are still unclear. Other factors such as photodegradation or
precipitation rate can also play a part in the observed seasonal
variations on the overall degradation of micropollutants. According
to Hai et al. (2011), a temperature increase (from 10 to 45 C) caused
an increase in total organic carbon (TOC) and total nitrogen (TN)
levels in the bioreactor supernatant, as well as higher concentrations of soluble microbial products released in the mixed liquor.
Besides, results of experiments measuring the removal of micropollutants at different temperatures in a batch mode demonstrated
the existence of a temperature dependent correlation between
hydrophobicity, molecular properties, and micropollutant removal.
Experiments conducted at 45 C allowed a good removal of some
less hydrophobic (log Kow < 3.2) micropollutants possessing strong
EWGs. On the contrary, the removal of most of the hydrophobic
compounds (log Kow > 3.2) was stable around 80e100% for experiments conducted in a temperature range of 10e35 C, but became
very low for temperature above 45 C (<40% for E1, BPA, EE2 for
rez et al. (2012) concluded that the influinstance). Moreover, Sua
ence of temperature is inversely proportional to the biological
degradation rate constants of PPCPs, and that temperature is a
relevant factor for the elimination of PPCPs with moderate to low
Kbiol. Finally, Kruglova et al. (2014) studied the removal of three
pharmaceuticals using a nitrifying activated sludge at a 12 C
operated at full-scale in a WWTP and in a laboratory-scale
sequencing batch reactor. Under this temperature, CBZ showed
no biodegradation, IBP was almost completely removed (up to
99%), and DCF showed high concentration fluctuations, as revealed
by Table 1. This latter phenomenon could be caused by time variations in nitrite concentration during the development of the nitrifying biomass (Barbieri et al., 2012). Since, each biomass such as
carbon oxidizing heterotrophs, nitrifiers, or denitrifiers, have their
own temperature correction factor, and the effect of temperature
on the reaction rate of a biological process can be expressed by the
Arrhenius equation, nitrification is the most temperature sensitive
process in biological system. The conversion of ammonia into nitrate due to the presence of autotrophic biomass may have slowed
down due to a temperature decrease. The temperature correction
factor of 1.072 is widely recently accepted for designing wastewater
treatment plants (Melcer and Water Environment Federation,
2003; Hwang and Oleszkiewicz, 2007). Hwang and Oleszkiewicz
(2007) investigated the effect of temperature decrease on nitrification. A sudden temperature decrease from 20 C to 10 C had an
important effect on nitrification, more intense than predicted by
the commonly used temperature correction factor. With this abrupt
10 C temperature decrease, a 20% decrease of the nitrification rate
was observed. On the contrary, a gradual temperature change of
minus 2 C per day induced a nitrification rate decrease similar to
the prediction with the temperature correction factor of 1.072
(Hwang and Oleszkiewicz, 2007). Thus, consequence on
nitrification biomass may have an impact on micropollutant
removal.
To conclude on this point, even though these two parameters
have an influence on the removal of organic micropollutants, the
modification or regulation of pH and temperature requires a large
amount of energy, and acid and base products, which is hardly
economically feasible for municipal WWTPs. However, these parameters could be monitored and regulated for concentrated industrial wastewaters, which have a low hydraulic flow.
2.3. Feeding effects on removal efficiency: batch vs continuous
A bioreactor may be classified as batch, fed batch, or continuous.
A typical batch reactor consists in an agitated tank, equipped with a
temperature regulating system, in which a bioreaction is carried on
without any addition until the reaction is considered to be complete. A fed-batch reactor is a process during which one or more
substrates are added to the bioreactor during the cultivation, while
the products remain in the bioreactor until the end of the experiment. Finally, a continuous reactor is one in which substrates are
continuously fed into the reactor, and from which a continuous
stream of products is drawn (Nanda and Pharm, 2008). These
feeding modes can largely influence the removal efficiencies of
different micropollutant families. A few authors have performed
experiments with bioreactors operated according to different
feeding modes and noticed significant differences on the biodegradation percentages obtained. Jeli
c et al. (2012a) studied the
degradation of CBZ and its metabolites using an air pulsed fluidized
bed bioreactor (FBR) inoculated with T. versicolor and operated in
fed-batch and continuous mode. A unique metabolite was found,
and CBZ was well removed (about 96%) after 2 days of FBR operated
in fed-batch mode. This percentage of degradation is higher than
the percentage obtained using Erlenmeyer flasks (94% after 6 days
of incubation), because glucose was continuously added, pH was
controlled, and the air pulses supplied allowed the fungus used in
the fed-batch reactor to thrive. However, using a continuous mode
operation with a hydraulic retention time of 3 days, only 54% of the
inlet concentration was degraded after the reactor reached a steady
state (25 days). This corresponds to a CBZ degradation rate of
11.9 mg CBZ g 1.dry weight pellets.d 1. A sufficient supply of nutrients was also considered as a crucial parameter for an effective
removal of CBZ by Zhang and Geißen (2012), who used a bioreactor
inoculated with Phanerochaete chrysosporium in both batch and
continuous modes. Under continuous operation, and thus input of
nutrients, a high elimination of CBZ (60e80%) was achieved, and
the elimination rate was stabilized around 100 days. Regarding
batch experiments, a high elimination was achieved after 4 h
(around 80%), mostly due to an adsorption onto the foam. The
proportion of biotransformation in CBZ elimination during the
batch experiment varied between 21 and 68%. In addition, the
elimination of some pharmaceutical compounds was also studied
by Rodarte-Morales et al. (2012), using a fed-batch reactor and a
continuous stirred tank reactor. A continuous feeding in the stirred
tank reactor operated with free pellets of P. Chrysosporium allowed
a complete degradation of three NSAID: DCF, IBP, and NPX; a partial
elimination of CBZ, but no degradation of DZP. Using fixed-bed
reactors under either continuous air flow or oxygen pulses, DCF,
IBP, and NPX were well removed under both aeration conditions,
while CBZ and DZP were only partially (60e90%) degraded
throughout these experiments.
Nevertheless, feeding effects, such as sequencing batch reactor
versus continuous flow, have hardly an impact on the removal of
micropollutants, which are predominated by the influence of SRT,
temperature, and batch or plug flow reactor. Besides, because of the
low concentration of organic micropollutants in wastewaters and
C. Grandclement et al. / Water Research 111 (2017) 297e317
their first order reaction, batch or plug flow reactors seem to be
more efficient than completely mixed reactors, especially regarding
the toxicity of influents.
2.4. Effects of microorganism communities or enzymes extracted
from microorganisms on removal efficiency
Table 2 (Appendix A: supplementary data) presents the removal
of selected micropollutants using classical bioreactors (batch experiments or MBR systems), depending on microorganism communities or enzyme extracted from microorganisms.
2.4.1. Activated sludge
Most studies dealing with the problem of micropollutant
degradation have thus used batch or membrane bioreactors inoculated with activated sludge from classical wastewater treatment,
in order to investigate removal efficiencies of these molecules. Luo
et al. (2015) notably investigated the performance of a conventional
MBR, and membrane fouling during the treatment of different
micropollutants. The results showed moderate or low removal of
KPF, CBZ, primidone (PRM), BPA (50%, 10%, 58%, 50% respectively)
and a significant membrane fouling as compared to the hybrid
moving bed biofilm reactoreMBR. Wijekoon et al. (2013) investigated the relationship between molecular properties and the fate of
29 micropollutants such as UV-filter, pesticides, phytoestrogens, or
pharmaceuticals using a MBR inoculated with municipal activated
sludge. Adsorption is the dominant removal mechanism from the
aqueous phase for hydrophobic (log Kow > 3.2) compounds (up to
50%), while biodegradation is the most important removal mechanism from the aqueous phase for hydrophilic compounds (up to
70%) (see Table 2). Compounds with a moderate hydrophobicity
that remains recalcitrant to biodegradation, such as CBZ, accumulated significantly onto the solid phase while highly hydrophobic,
but readily biodegradable compounds (up to 75%), such as E1 and
E2, did not accumulate onto activated sludge solids (<20%)
(Wijekoon et al., 2013). Fan et al. (2014) investigated the removal
efficiencies of five pharmaceuticals from synthetic domestic
wastewater using a submerged MBR. They studied separately the
contributions of sludge adsorption and biodegradation, as provided
in Table 2. The results of batch adsorption experiments at different
reaction times of 0e6 h, using sterilized sludge, showed that the
removal efficiencies of ACE, E2, NPX, DCF, and CBZ by sludge
adsorption were 28, 68, 60, 40, and 72% respectively. Besides, these
adsorption percentages were positively correlated to the molecules
Kow. The results of batch experiments using activated sludge
showed that 83% of ACE, 98% of E2, and 47% of NPX were removed
due to a combination of sludge adsorption and biodegradation,
while adsorption of these molecules onto the sludge solid phase
was only 1.8, 1.3, and 7.0% respectively. Regarding the continuous
process, the average removal efficiencies observed in the submerged MBR for ACE, E2, NPX, and DCF was about 92, 90, 55, 39%
respectively and low removal efficiency of CBZ (<5%) was also
observed. Biodegradation thus seems to be the main way of
degradation for ACE, E2, and NPX. On the other hand, the removal of
DCF was mainly achieved by sludge adsorption. Indeed, the total
removal efficiency of DCF was 19.7% and the contributions of sludge
adsorption and biodegradation were 14.9 and 4.8% respectively.
Regarding CBZ, this compound still remains recalcitrant and its
removal efficiency only achieved 8.9% (see Table 2). This implies
that, in the operating conditions studied by the authors, neither
sludge adsorption nor biodegradation was very effective for the
removal of CBZ.
However, a few studies have evaluated the influence of different
micropollutants on the bacterial community. The experiments
were based on evaluating the influence of some compounds on the
305
endogenous and exogenous respiration using a biomass initially
sampled from a CAS process. For instance, Aubenneau et al. (2010)
evaluated the potential effect of CBZ on the heterotrophic microorganisms taken from CAS and a pilot-scale MBR. During batch
tests, they noticed some effects on both the respiratory activity of
the bacterial community and on the floc size. Moreover, no inhibition, and no significant difference on chemical oxygen demand
(COD) removal, sludge production, or oxygen requirement were
observed with or without 1 mg.L 1 of CBZ, during a MBR wastewater
treatment experiment. The authors have chosen this concentration,
which was higher than WWTP influents, in order to induce a strong
biomass reaction. On the one hand, under endogenous conditions,
the observed increase of oxygen uptake rate (OUR) suggests an
increase in maintenance requirements, essentially to manage the
chemical stress induced by the CBZ's presence. On the other hand,
under exogenous conditions, an OUR decrease was noticed. This
observation could suggest a change in the metabolic pathways of
the substrate or in the active bacterial species (Aubenneau et al.,
2010). However, further studies would be useful to predict the influence of micropollutants on WWTP bacterial communities, but
the concentration of CBZ found in municipal wastewaters should
have no effect on biomass that treats wastewaters with several
hundred mg.L 1 of COD.
2.4.2. White-rot fungi
A biological alternative to activated sludge and a promising
process may be based on the use of white rot fungi (WRF) cultures.
These microorganisms were reported to degrade a wide range of
xenobiotics due to the action of fungal oxidative enzymes, such as
manganese peroxidase (MnP), lignin peroxidase (LiP), versatile
peroxidase (VP), or laccase. MnP (Mn(II): hydrogen-peroxide oxydoreductase, EC 1.11.1.13) is a heme glycoprotein enzyme which
catalyzes the oxidation of organic compounds in the presence of
H2O2 (Wong, 2009). LiP (diarylpropane: oxygen, hydrogenperoxide oxidoreductase (CeC-bond-cleaving), EC 1.11.1.14) catalyzes the H2O2-dependent oxidative depolymerization of lignin. LiP
has been shown to eliminate several recalcitrant aromatic compounds such as PAH and phenolic compounds (Christian et al.,
2005). VP (EC 1.11.1.16) is a hemoprotein which combines the
substrate-specificity characteristics of the two other ligninolytic
peroxidases, MnP and LiP. It is able to involve multiple binding sites
for substrates in order to oxidize phenolic and non-phenolic substrates, hydroquinones, and both low- and high-redox-potential
dyes (Camarero et al., 1999). Finally, laccase (benzenediol: oxygen
oxidoreductase, EC 1.10.3.2) belongs to a family of multicopper
enzymes of low-specificity. The enzyme catalyzes the oxidation of
hydrogen-donating substrates such as lignin, phenol, or acrylamines via the four-electron reduction of O2 to H2O. Laccase oxidizes phenolic compounds in the presence of O2 (Wong, 2009; Yang
et al., 2013b). All fungal species cannot secrete all four extracellular
enzymes which have been reported to oxidize persistent TrOCs.
Apart from theses enzymes, intracellular enzyme systems, such as
cytochrome P450, have also been reported to play important roles
in the removal of some TrOCs (Golan-Rozen et al., 2011).
In previous studies, the removals of TrOCs by different species of
WRF have been reported. Bouchiat et al. (2016) investigated the
removal of four emergent pollutants (di(2-ethylhexyl)phthalate
(DEHP), fluoranthene (Fl), aminomethylphosphonic acid (AMPA),
and E1) by filamentous fungi (Fusarium oxysporum, Geotrichum
galactomyces, Trichoderma harzianum, and Fusarium solani) in
mineral medium for 10 days. Except for E1 which was not degraded
by any fungi, AMPA was degraded at 69% by T. harzianum, and DEHP
was completely degraded by F. oxysporum and F. solani after 10 days
of incubation. Fl was not significantly degraded by G. galatomyces
and T. harzianum, whereas the degradation by F. oxysporum and
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C. Grandclement et al. / Water Research 111 (2017) 297e317
F. solani was moderate, as revealed Table 2 (42 and 12%
respectively).
However, previous works showed that the results varied
depending on the tested enzyme systems. Trametes versicolor,
which seems to have a good potential for the degradation of
et al., 2013; Nguyen et al., 2014b),
micropollutants (Cruz-Morato
secretes all four types of extracellular enzyme systems (laccase may
be the predominant one in some strains). The cytochrome P450
system may be also involved in the first step of the degradation of
some pharmaceuticals (Marco-Urrea et al., 2009). It allowed for
high removal rates, especially with some of the most recalcitrant
compounds such as CBZ, CFA, DCF, DIU in batch experiments
(Bending et al., 2002; Marco-Urrea et al., 2009; Tran et al., 2010;
Jeli
c et al., 2012a; Margot et al., 2013b). At least two different
mechanisms using cytochrome P450 or laccase were described by
Marco-Urrea et al. (2009) for the almost complete (94%) removal
of DCF during the first hour of incubation, using T. versicolor mycelia
pellets in Erlenmeyer flasks. The cytochrome P450 system may also
be involved in the first step of CFA and CBZ oxidation by T. versicolor
(which reached 91% and 57% respectively after 7 days of incubation). On the contrary, extracellular fungal enzyme systems did not
appear to play a significant role during the first step of degradation
(Marco-Urrea et al., 2009). Intracellular enzymes may be involved
in the biodegradation of KPF, propyphenazone (PPZ), fenoprofen
(FEP), and GFZ, while laccase preferentially removed DCF, NPX, and
indomethacin (IDM) among the targeted PhACs degraded by the
whole fungal culture. Tran et al. (2010) noticed a complete removal
of DCF, NPX, IDM, IBP, and FEP and partial degradation of other
selected PhACs after 48 h of incubation with the 7-day-old liquid
fungal culture, both in the presence and absence of a laccase
mediator
ABTS
(2,20 -azino-bis-(3-ethylbenzothiazoline)-6sulfonate) (see Table 2). T. versicolor also seems to be able to
degrade pesticides. Important degradations of DIU, ATZ, and terbuthylazine were achieved after 42 days of batch experiments (99%,
>86%, and 63% respectively), whereas for metalaxyl less than 44%
was reached (Bending et al., 2002). At last, T. versicolor also showed
a good ability to remove endocrine-disrupting compounds such as
BPA, NP, or EE2 (Cajthaml et al., 2009) even recalcitrant anticancer
drugs such as azathioprine, etoposide (more than 97% after 8 days
in batch experiment) (Ferrando-Climent et al., 2015). Polychlorinated biphenyls (PCBs) can also be degraded by T. versicolor.
Ruiz-Aguilar (2002) studied the degradation of a mixture of PCBs at
high initial concentrations from 600 to 3000 mg.L 1, in the presence of a non-ionic surfactant (Tween 80). PCB degradation ranged
from 29 to 70% using T. versicolor in 10-day incubation tests.
Nevertheless, only a few studies related to fungi have focused on
the degradation of PhACs from real urban wastewater under nonsterile conditions, in the presence of mixtures of contaminants at
low concentrations (ng.L 1 to mg.L 1) as well as other active mi et al. (2013) used a batch fluidized bed
croorganisms. Cruz-Morato
bioreactor to evaluate the effects of non-sterile urban wastewater
substrate on a T. versicolor culture. They concluded that T. versicolor
can remain active when fed with real wastewater where bacteria
and contaminants are present, if a source of nutrients such as
glucose and nitrogen is also added to maintain a significant biological fungus activity. Using this batch FBR, around half of the
detected PhACs (at environmentally relevant concentrations) achieved a complete removal, while 25% were partially removed
(average removal of 35% after 8 days). Regarding the other compounds, no degradation or very low degradation (<20%) was
observed. For instance, CBZ showed no removal from the real
wastewater used. Furthermore, its concentration increased to 37%
after 8 days due to deconjugation of CBZ intermediates (Kovalova
et al., 2012). Yang et al. (2013a) studied the removal of DCF and
BPA using a MBR inoculated with T. versicolor and operated during
three months under non-sterile conditions. They confirmed that
biodegradation is the main mechanism for the removal of both
compounds. Relatively stable removals of BPA (80e90%) and DCF
(~55%) were achieved by applying a HRT of two days. Besides,
T. versicolor also seems to be able to degrade CBZ in aqueous medium using an air pulsed fluidized bed bioreactor operated in batch
or continuous mode. Using the batch reactor, the CBZ removal
achieved 96% within 2 days, whereas CBZ concentration decreased
by more than a half (54%), using the air pulsed fluidized bed
bioreactor operated at steady state in continuous mode with a HRT
of 3 days. However, according to Erlenmeyer flask batch experiments, CBZ (at 9 mg.L 1) was almost completely eliminated (94%)
after 6 days, while close to environmentally relevant concentrations (50 mg.L 1), 61% of the contaminant was degraded after 7
days (Jelic et al., 2012a). Studies on the relative contribution of
biosorption compared to various modes of biodegradation (e.g.,
extracellular enzyme dependent or independent) during fungal
removal of TrOC remain scarce. Nguyen et al. (2014b) confirmed
that biodegradation was, over all studied compounds, the main
mechanism of removal. However, hydrophilic compounds generally
remained poorly removed which may indicate the importance of
biosorption in subsequent degradation by whole-cell cultures. In
addition, inhibiting the intracellular cytochrome P450 during the
degradation of some TrOCs by whole-cell cultures resulted in a
reduction in biodegradation efficiencies which may point out the
importance of extracellular enzyme-independent catalytic pathways. The degradation profile of the tested TrOCs using a WRF
fungal culture is quite different from that obtained using activated
sludge processes.
The capacity of P. chrysosporium to remove TrOCs has also been
evaluated by some authors despite the lack of laccase and VP in
their enzymatic system (Hatakka, 1994). P. chrysosporium has been
reported to achieve high removals in the case of some pharmaceuticals. The elimination of NPX and DCF was 100% after 4 days of
incubation in batch (Rodarte-Morales et al., 2011). In a subsequent
study, Rodarte-Morales et al. (2012) used a stirred-tank reactor
inoculated with free pellets of P. chrysosporium to remove PhACs.
High removal efficiencies, collected in Table 2, were achieved for
DCF, NPX, and IBU (94, 94, and 100% respectively); CBZ removal
varied from 24 to 63% between 20 and 50 days of operation. Using
P. chrysosporium immobilized on polyurethane foam in a stirredtank reactor, removal efficiencies were 93% for DCF and IBP, up to
90% for NPX during the first 3 days then decreased between 65 and
77%. Since some metabolites of these compounds have been identified in the reactors, their back transformation into parent compounds could explain the observed decrease in removal efficiency
after day 3. A chemical balance between precursors and metabolites could also disadvantage the biodegradation of the active
product. For antiepilectic and tranquilizers, the removal percentages were up to 50% (Rodarte-Morales et al., 2012). The same FBR
was operated 100 days with immobilized P. chrysosporium, DCF, IBP,
and NPX were completely removed regardless of the continuous air
flow (1 L.min 1) or pulsation of oxygen, and high removal efficiencies were observed for CBZ and diazepam (60e90%) (RodarteMorales et al., 2012). Li et al. (2015) also reported an almost complete removal of NPX and a 60e80% removal of CBZ after two
weeks, using a fixed-bed bioreactor packed with a mixture of WRF
pellets and wood chips (see Table 2). However, after the 14th day,
the removal efficiencies for both compounds suddenly dropped due
to a possible contamination by other microorganisms. According to
the studies, CBZ biodegradation experiments using WRF bioreactors presented high variations. CBZ either showed no removal
at all after 7 days of incubation in batch experiment inoculated with
a blended mycelial suspension of P. chrysosporium (Marco-Urrea
et al., 2009); limited degradation (<10%) during in vitro batch
C. Grandclement et al. / Water Research 111 (2017) 297e317
experiments, using LiP from P. chrysosporium (Zhang and Geißen,
2010), but high elimination was achieved with a novel plate
bioreactor, using P. chrysosporium grown on polyether foam under
non-sterile conditions (Zhang and Geißen, 2012). Endocrine disrupting
compounds
have also
been
degraded
using
P. chrysosporium cultures. For example, almost complete removal of
NP (up to 90%) has been reported using a 3 day-batch experiment
(Cajthaml et al., 2009; Subramanian and Yadav, 2009), while only
30e50% of removal has been reported by Soares et al. (2005) after
25 days of incubation in aerobic batch experiments. Regarding
pesticides, poor removal of CFA has been observed after 7 days of
incubation with P. chrysosporium (Marco-Urrea et al., 2009). Besides, almost complete removal of DIU has been reported after 10
days in Erlenmeyer flasks and the results showed that the presence
of this herbicide did not cause any drop in the biomass production
(Coelho-Moreira et al., 2013). Although P. chrysosporium has been
intensively studied as a model for the white-rot group of basidiomycete, there is an increasing evidence in the literature that
T. versicolor can degrade xenobiotics more efficiently than
P. chrysosporium or other species of WRF like Bjerkandura adusta, or
Pleurotus ostreatus (Soares et al., 2005; Cajthaml et al., 2009;
Marco-Urrea et al., 2009).
Enzymatic treatments seem very attractive as far as the removal
of TrOCs from wastewaters is concerned. They consume less
chemicals, water and energy and produce less wastes than other
chemically catalyzed bioprocesses. Recent studies have investigated the capacity of laccase solutions to degrade a wide range of
TrOCs that are persistent, using other biological processes. On the
one hand, Margot et al. (2013a) investigated the ability of four
strains of the bacterial genus Streptomyces (S. cyaneus, S. ipomoea,
S. griseus and S. psammoticus) and the white-rot fungus T. versicolor
to produce active extracellular laccase in biologically treated
wastewater with different carbon sources. They concluded that
T. versicolor was the most promising strain. Indeed, this fungus
produced more than 20-times more laccase activity than S. cyaneus,
the best candidate of the Streptomyces strains evaluated (especially
in treated wastewater with forestry waste as the sole substrate).
Besides, laccase from T. versicolor was more active than that from
S. cyaneus near neutral pH and between 10 and 25 C (conditions
usually found in municipal wastewater) and presented faster
degradation kinetics of DCF, BPA, and MFA (Margot et al., 2013a). On
the other hand, purified or commercial laccase solutions showed
high removal of DCF, NPX, TCS, NP, E1, EE2, and BPA, as collected in
Table 2 (Kim and Nicell, 2006; Lloret et al., 2010; Tran et al., 2010),
but the removal of CBZ, IBP, and CFA still remains low (<40%) (Tran
et al., 2010). Besides, a complete degradation of DCF was observed
using LiP from P. chrysosporium at pH 3.0e4.5 with 3e24 ppm H2O2
(Zhang and Geißen, 2010). However, CBZ degradation was limited
(mostly below 10%) using LiP from P. chrysosporium (Zhang and
Geißen, 2010) whereas it seems to be well oxidized (98%) by MnP
and VP produced by P. ostreatus after 32 days of incubation in
Erlenmeyer flasks (Golan-Rozen et al., 2011). MnP solutions can
also degrade efficiently methoxychlor (69% after 24 h) (Hirai et al.,
2004), BPA (100% after 12 h) (Tsutsumi et al., 2001), and hormones
such as EE2, E1 (>80% after 8 h) (Suzuki et al., 2003). An innovative
strategy based on the induction of hydroxyl radicals in T. versicolor
using the quinone redox cycling have been studied by (Marco-Urrea
et al., 2010). The results of this study showed a high percentage of
CBZ removal (80%) after 6 h of batch experiment.
Redox mediators act as electron shuttles between the oxidizing
enzyme and target compounds to enhance fungal enzyme-catalysis
depending on both TrOC and mediator molecular structures (Kim
and Nicell, 2006). In the case of laccase, two oxidative steps are
involved. First, laccase oxidizes the mediator which finally transfers
the electron to the substance of interest. The most commonly used
307
redox mediators are 1-hydroxybenzotriazole (HBT), ABTS or violuric acid (VA) (Fabbrini et al., 2002). The influence of different
mediators (synthetic and natural) and of their concentration on the
laccase-based oxidation system were evaluated by Lloret et al.
(2010). Among the different selected natural or synthetic mediators, syringaldehyde (SA) or HBT as well, greatly enhanced the action of the laccase enzyme, in the case of the biodegradation of
estrogens and DCF (100% after 15min and 1 h of incubation
respectively). The other natural mediators (vanillin, p-coumaric
acid, or ferulic acid) presented significantly high efficiencies,
allowing to achieve percentages of removal ranging from 80% to
100% after 24 h of enzymatic reaction on DCF, NPX, EE2, E3, and E1.
HBT addition also improved the removal of pentachlorophenol
(31e91%), DCF (70e95%), and NPX (20e98%) (Nguyen et al., 2013b).
Tetracycline antibiotics were completely eliminated after 1 h using
a treatment with a laccase-HBT system (Suda et al., 2012).
Furthermore, Hata et al. (2010) suggested that repeated addition of
laccase and HBT is effective in CBZ removal. They observed a
decrease of 22% of CBZ concentration after 24 h using a single
treatment, and a drop of 60% after 48 h using a repeated treatment.
Even though the use of redox mediators seems to be relevant to
improve micropollutant removal in an enzymatic system, these
substances are somewhat toxic and further studies are needed to
evaluate their chronic toxicity and the effluent toxicity.
Previous studies have confirmed significant removal of TrOCs by
WRF under sterile batch test conditions. Only a few studies have
been conducted using a continuous flow fungal reactor in a nonsterile environment or a combination of white-rot fungi and activated sludge. Yang et al. (2013a) focused their study on the removal
of DCF and BPA using a fungal MBR in non-sterile conditions. In
these conditions and with a HRT of 2 days, relatively stable removal
rates for BPA (80e90%) and DCF (about 55%) were observed. The
degradation of 30 TrOCs using a WRF-augmented MBR was investigated by Nguyen et al. (2013b) and collected in Table 2. The obtained results suggest that activated sludge and WRF would be
complementary. Indeed, TrOCs resistant to bacterial degradation
such as DCF, TCS, NPX, and ATZ could be degraded by laccase and
further enhanced using HBT as a redox mediator (from 22 to 93%).
Nevertheless, a low removal was observed for some compounds
that are well removed by simple CAS treatment such as IBP, GBZ,
and amitriptyline. CAS and TrOCs degradation ability of the fungalenzyme was also studied during batch tests using crude enzyme
extracts (laccase). Over the 30 tested molecules, 13 significant
enzymatic degradations were observed. Some other molecules
showed low or negligible degradation. The variation of enzymatic
degradation efficiencies is attributed to the differences in chemical
structure of the selected TrOCs.
As in the case of CAS or MBR treatments, TrOC removal by
enzymatic systems is dependent upon a range of operating factors
such as pH, temperature, molecular structure, and so on. For
example, because laccase promotes the single electron oxidation of
phenols (Yang et al., 2013b), TrOC molecules with a hydroxyl group
attached to a benzene ring are highly degraded (70e90%) by laccase
extracts. The optimal temperature for laccase production is
25e30 C and the optimal temperature for peroxidases production
is 37e40 C (Cabana et al., 2007). PH is another controlling factor
that can influence the development of fungal cultures, and thus the
removal efficiencies of TrOCs using such cultures. Margot et al.
(2013b) investigated the optimal conditions for the transformation of two pharmaceuticals, DCF and MFA, one biocide, TCS,
and one plastic additive (BPA) by laccase from T. versicolor. Batch
experiments were conducted in spiked solutions at pH varying
from 3.0 to 9.0, enzyme concentration from 70 to 1400 U.L 1, reaction times (0e26 h) and temperature (10, 25 and 40 C). They
concluded that all four factors had a significant effect on the
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C. Grandclement et al. / Water Research 111 (2017) 297e317
micropollutant removal, but that the greatest effect was obtained
with pH. Even though, optimal conditions were compounddependent, they were found to be between pH ¼ 4.5 to 6.5 and
between 25 C to more than 40 C (Margot et al., 2013b). Other
studies evaluated the efficiency of some enzymes depending on pH
values. For instance, DCF was completely degraded by LiP in the pH
range of 3.0e4.5 whereas only 10% was degraded at pH ¼ 6.0 due to
the inactivation of LiP at higher pH (Zhang and Geißen, 2010).
However, pH ¼ 6.0 was reported as the optimum pH for laccasecatalyzed treatment of estrogens (Auriol et al., 2007) and BPA by
purified laccase from Trametes villosa (Fukuda et al., 2001). The
optimum pH for the degradation of TCS by laccase from T. versicolor
was observed at pH ¼ 5.0 (Kim and Nicell, 2006), while the optimal
pH for laccase to degrade chlorophenols was around 5.5 (Zhang
et al., 2008).
A few studies have investigated the degradation of TrOCs under
continuous operation using an enzymatic membrane reactor
(EMR). Lloret et al. (2012a,b) first used a fed-batch reactor to
evaluate the effect of process parameters such as gas composition
(air or oxygen), pH, enzyme concentration; then the continuous
degradation of E1 and E3 was investigated by an EMR, composed of
a stirred tank reactor coupled with an ultrafiltration membrane.
The highest removal rates under steady state operation reached up
to 95% for E1 and nearly complete degradation for E3. Nguyen et al.
(2014a) studied the effect of a redox mediator addition to a
continuous EMR on the removal of different TrOCs. Under these
conditions, high removals of BPA and DCF were achieved (>85% and
>60%, respectively). They were improved to >95% and >80%,
respectively, by adding a natural redox-mediator compound, SA
(5 mM), to the culture medium. In addition, the use of EMR can
facilitate the separation of enzymes from products and substrates
due to the semi-permeable membrane, and thus decrease the losses in enzymes that are often washed out with the treated effluent
when using conventional bioreactors (Lloret et al., 2012a,b).
Even though, the use of WRF or immobilized enzyme showed
interesting efficiencies regarding the removal of organic micropollutants, their implementation in a biological treatment seems
hardly effective because of the overgrowth by the normal biomass.
Moreover, this biological treatment could present a significant cost
for municipal WWTPs, but may be an interesting alternative for
industrial wastewaters.
2.4.3. Bacteria
The use of specific bacteria has often been investigated to
et al.,
remove PCBs or PAHs (Haritash and Kaushik, 2009; Murínova
2014; Isaac et al., 2015; Kuppusamy et al., 2016). However, only a
few authors, whose results are gathered in Table 2, have tried to use
specific bacteria isolated from activated sludge to remove PhACs. Li
et al. (2013) studied the degradation of CBZ by a bacterium which
can use CBZ as its sole source of carbon and energy. This strain was
identified as Pseudomonas sp. by the 16S rRNA gene sequence.
Pseudomonas sp. CBZ-4 can effectively degrade CBZ under optimal
conditions (pH 7.0, 10 C, mechanical stirring). After 144 h of incubation, the average removal rate of CBZ reached 46.6% (Li et al.,
2013). Another strain of Pseudomonas sp., Pseudomonas putida,
can be used for the oxidation of some micropollutants. As an
example, DCF is rapidly degraded during ongoing manganese
oxidation by P. putida MnB6 (Meerburg et al., 2012). A co-metabolic
removal of DCF has been proved to be the main degradation path
during active Mn2þ oxidation by P. putida. Regarding P. putida,
Kuddus et al. (2013) showed the production of laccase enzyme from
P. putida MTCC 7525 was achieved at 30 C at pH ¼ 8.0 after 108 h of
incubation. Besides, the optimal activity of the purified enzyme was
observed at pH ¼ 8.0 and 40 C. This bacterium, isolated from soil
samples containing sawdust and dairy effluents, was used in order
to treat synthetic dyes and industrial effluents. The results
respectively showed 74e93% and 58e68% of decolorization within
24 h of incubation (Kuddus et al., 2013). Furthermore, YanzeKontchou and Gschwind (1994) observed up to 50% of mineralization for ATZ using Pseudomonas DSM93-99 in batch experiments.
Streptomyces sp. is also of some interest for the degradation of
micropollutants. Streptomyces MIUG 4.89 was studied by Popa
Ungureanu et al. (2014) for its ability in CBZ biodegradation during cultivation in a submerged system under aerobic conditions at
an initial CBZ concentration of 0.2 mg.L 1. A 30% of CBZ biotransformation was yielded under optimal conditions: liquid medium
containing 6.5 g.L 1 glucose and 2 g.L 1 yeast extract, inoculated at
7% (v/v) and cultivated at pH 6.0, during 7 days of incubation at
25 C and 150 rpm. Besides, Castillo et al. (2006) observed the
degradation of DIU by 17 Streptomyces strains isolated from agricultural and non-agricultural soils. Twelve strains degraded the
herbicide by up to 50% and four of them by up to 70%. Strain A7-9,
belonging to the S. albidoflavus cluster, was the most efficient organism (95% of degradation after 5 days of incubation and complete
degradation after 10 days).
DIU can also be degraded by Sphingomonas sp., even at low
pollutant concentrations (mg.L 1). Sphingomonas sp. SRS2 is capable
of DIU mineralization with an initial degradation pathway consisting of two successive N-demethylations, followed by a cleavage
of the urea group that gives 3,4-dichloroaniline (3,4-DCA). 86% of
14
C-carbonyl-diuron was mineralized to 14CO2 within 72 days.
Moreover, the mineralization activity can be enhanced by
combining SRS2 with the 3,4-DCA-mineralizing Variovorax sp.
SRS16 (Sorensen et al., 2003). In other studies, Sphingomonas sp. has
been used to remove some PAHs. Rentz et al. (2008) showed that
concentrations around 1.2 mg.L 1 of benzo[a]pyrene (BaP) were
completely removed within 20 h of batch experiments when
Sphingomonas yanoikuyae JAR02 was grown on salicylate. S.
yanoikuyae JAR02 uses salicylate as an inducer, as well as a carbon
and energy source. Indeed, aerobic bioremediation of high molecular weight PAH uses a co-metabolic degradation that requires a
carbon/energy source, an inducer of catabolic enzymes, and oxygen
(Rentz et al., 2008). Guo et al. (2010) also studied the degradation of
a mixture of PAHs comprising phenanthrene (Phe), Fl, and pyrene
(Pyr) by Sphingomonas strains isolated from mangrove sediments
(see Table 2). Phe was degraded by more than 50% and Fl was
degraded between 30 and 60%, but Pyr degradation was less than
30% after 7 days of batch experiments. A co-culture of Sphingomonas and Mycobacterium strains enhanced the degradation of all
three PAHs (complete removal after 7 days) (Guo et al., 2010).
Regarding pharmaceuticals, Murdoch and Hay (2005) studied the
degradation of IBP using Sphingomonas sp. strain Ibu-2 isolated
from a WWTP, based on its ability to use IBP as a sole carbon and
energy source. They observed a complete removal after 80 h of
batch experiments.
Widehem et al. (2002) isolated and characterized Arthrobacter
sp. N2 from soil treated over several years with DIU. This strain was
able to aerobically transform DIU into 3,4-DCA in pure culture,
either alone, or in the presence of alternatives carbon sources.
Besides, Arthrobacter globiformis D47 was shown to degrade a range
of substituted phenylurea herbicides in soils because of two plasmids of approximately 47 kb and 34 kb. This strain was tested by
Turnbull et al. (2001) for its ability to degrade DIU, which demonstrated that the degradative genes were located on the 47-kb
plasmid. When A. globiformis D47 was added to soil samples, the
strain was able to degrade other urea pesticides (>90% after 10
days), such as chlortoluron, isoproturon, linuron, monolinuron, and
monuron, initially introduced at 20 mg.L 1.
C. Grandclement et al. / Water Research 111 (2017) 297e317
2.5. Factors limiting the biodegradation in wastewater treatment
plants
Several methods have been tested in WWTPs in order to remove
micropollutants from effluents, but these physical, chemical, or
biological treatments did not show significant results. The
advanced oxidation processes using O3, UV, Fenton showed high
removals, but generated some byproducts whose toxic effects and
risks on health are still unknown (Benner et al., 2013). Regarding
treatments using filtration, membrane fouling is the main limit
because of the high organic matter content characteristic of
wastewaters. Besides, these processes involve high energy ren
~ ez et al.,
quirements and important maintenance costs (Ordo
2014). Submerged membrane systems need frequent air scouring
to reduce cake deposit on them and to generate localized cross-flow
conditions along the membrane surface. Some studies investigating
membrane fouling in MBR processes have reported the significance
of colloidal particles as an important factor contributing to fouling
development. Colloids are responsible for 25e50% of the total
measured fouling (Defrance et al., 2000; Bouhabila, 2001).
Furthermore, as stated above, the removal of compounds, and
specially xenobiotics, using CAS is often not mainly due to
biodegradation, but also to adsorption on activated sludge flocs,
and to a less extent to air stripping. Many micropollutants must be
considered stable in biological processes for municipal wastewater
treatment (Falås et al., 2016), and due to adsorption they are just
transferred to another phase, and thus still released in the environment. They are not degraded into less toxic species and they
might cause health problem again (Dionisi, 2014). Biodegradation
can only occur when the substrate is dissolved in the liquid phase.
Because of the competition between air stripping and adsorption
on microorganisms, the concentration in the liquid, and thus the
substrate concentration available for biodegradation is reduced
(Byrns, 2001).
Regarding enzymatic membrane processes, membrane fouling,
enzyme retention, and enzyme activity decay are responsible for
strong limitations on the performance of EMRs. This seems to be in
tight connection with several phenomena such as catalyst leakage,
but also enzyme denaturation due to various factors including
physical ones (pH, temperature, shear stress), or chemical and
biological inhibitors. Regarding shear stress, the effect of this factor
has been a subject of discussion for Rios et al. (2004) since shear
stress seems more difficult to characterize than enzyme leakage.
Other authors, such as Jaspe and Hagen (2006), found no evidence
of relationship between shear rates and the destabilization of the
studied folded protein (horse cytochrome c, 104 amino acids).
Moreover, Mendoza et al. (2011) observed that denaturation enzymes may be further exacerbated when a wastewater containing
the target pollutant is continuously introduced into the reactor.
Lloret et al. (2012b) observed no enzyme denaturation within a
short 8 h-period during the evaluation of continuous TrOC removal
by an EMR, but beyond 24 h of continuous operation, a gradual drop
in enzymatic activity was recorded. The observed decrease was due
to enzyme denaturation rather than to the permeation of enzyme
through the membrane, because no enzyme was detectable in the
permeate.
To ensure the technical and economical viability of such EMR
processes at industrial scale, more studies need to be conducted.
Because of the huge volume of wastewaters, and thus the important quantities of enzymes that are needed, reactors with free enzymes may not be an economically viable solution for wastewater
treatment. Besides, a complementary treatment may be needed at
the end of such processes to remove the enzymes from the effluent.
Nevertheless, for industrial-scale requests, using immobilized enzymes could be an interesting solution to decrease the cost, by
309
reusing the biocatalyst. In addition, enzyme immobilization
generally results in an enhancement of the biocatalyst stability
even if enzymatic decay is still observed with time. This also increases the contact surface between enzymes and substrates, and
maintains a good catalytic efficiency over many reaction cycles (de
Cazes et al., 2014). As a consequence, processes with enzymes
supported on a solid phase, or using cross-linking enzymes aggregates, represent interesting options to remove micropollutants.
3. Areas for improvement
3.1. Hybrid process description
As it has been previously explained in this paper, conventional
processes based on activated sludge are often not sufficient to
ensure high removals for most organic micropollutants. As a
consequence, different alternative technologies, such as hybrid
processes which are a combination of two or more treatment
processes, have been studied that may appear to be effective to
remove micropollutants. Indeed, the removal of some recalcitrant
compounds can be improved with the combination of two processes due to synergistic effects. For instance, the addition of activated carbon can enhance the elimination of poorly biodegradable
organic compounds by adsorption (Alvarino et al., 2016a). The
combination of biofilm with suspended biomass can also improve
the potential biodegradation of organic micropollutants due to an
enhancement of the biodiversity into the systems.
The use of a biofilter system containing a fixed biofilm has been
studied with a main focus on porous media biofilm processes such
Casas and Bester, 2015). As SRT is an imporas sand filters (Escola
tant factor with respect to TrOCs' removal in classical systems based
on suspended biomass cultures, interesting results can be expected
using low loaded biofilm processes that will tend to promote a
more diverse bacterial population. Joss et al. (2004) evaluated the
removal of estrogens obtained with a full-scale submerged biofilm
reactor using a Biostyr™ system as a support and a reference for
activated sludge process. They showed only slightly lower removal
in the biofilm reactor, despite a much longer HRT in the activated
sludge process. These results suggested that the shorter reaction
time in the biofilm reactor can be compensated by a higher biomass
concentration and/or a higher pharmaceutical removal capacity per
unit of biomass, probably associated to the development of slow
growing bacteria in the biofilm. Attached-growth processes thus
offer a number of advantages mostly linked to an enlargement of
the range of possible active strains, due to the development of slowgrowing microorganisms on the carrier media. The acidic pharmaceuticals such as IBP, DCF KPF, or NPX removals during batch
experiments using activated sludge on the one hand, and suspended biofilm carriers on the other hand (AnoxKaldnes™ type K1
media) were compared by Falås et al. (2012). In their subsequent
study, during batch experiments, Falås et al. (2013) evaluated the
efficiency of a hybrid suspended & attached growth process obtained by combination of biofilm carriers with a free activated
sludge. Results were used to extrapolate the micropollutant
removal at full-scale. The model estimations indicated that, in
hybrid biofilm activated sludge processes, the attached biomass can
significantly contribute to the removal of some micropollutants,
such as DCF. In this process, two different communities of bacteria
have been observed such as a slow growing community in the
carrier biomass, and ammonia and nitrite oxidizing bacteria in the
free biomass (Falås et al., 2013). Along with such biofilm technologies, moving bed biofilm reactors (MBBR) also seem to be a
promising solution to remove micropollutants. In this context,
Casas et al. (2015) proposed to remove pharmaceuticals from
hospital wastewaters using a MBBR. In this system, the biofilm
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C. Grandclement et al. / Water Research 111 (2017) 297e317
grew on small (1e4 cm diameter) plastic carriers which are suspended in a reactor. In this case, the process can be as robust as
activated sludge treatment (because of the enhancement of nitrification), and has the advantage of a biofilm reactor regarding the
presence of slow-growing bacteria. In a subsequent study Escol
a
Casas et al. (2015) evaluated the ability of a combination of suspended activated sludge and biofilm (polyethylene carriers for
biofilm growth are suspended within activated sludge) on the
removal of different micropollutants. The hybrid process Hybas™
(VeoliaWater Technology), based on the integrated fixed-film
activated sludge technology, contains two separate biomasses.
This process combines a fast growing biomass with low sludge age
in free activated sludge flocs, and a slow-growing biomass with
high sludge age on MBBR-carriers. For this study, a pilot plant
consisting in a series of one activated sludge reactor, two hybrid
processes and one MBBR have been established and successfully
processed during 10 months under continuous operation (Escol
a
Casas et al., 2015).
Apart from plastic biofilm carriers, other materials can be used
for attached-growth microorganisms, such a polyurethane sponge.
The efficiency of sponge-based MBBRs in removing organic matter,
among dissolved organic carbon, nitrogen and phosphorus have
been investigated by some authors such as Ngo et al. (2008). Luo
et al. (2014a) evaluated the short-term removal rates of five
micropollutants during 24 h-batch experiments using nonacclimatized and acclimatized sponge supported biomasses
(acclimatization to the synthetic wastewater without addition of
micropollutants for 20 days until TOC, TN, and PO4-P removal
became stable). Then, a continuous bench-scale MBBR was set for a
long-term assessment (100 days' period) of selected micropollutant
removal. In their subsequent study, Luo et al. (2015) compared the
removal of micropollutants using a conventional MBR and a hybrid
MBBR-MBR system. Results notably showed that the hybrid
MBBReMBR system could effectively remove most of the studied
micropollutants thanks to biodegradation pathways, while the
conventional MBR was less effective for compounds such as KPF,
CBZ, PRM, BPA, and E3. Besides, membrane fouling was minimized
with the hybrid system because of the alteration of the soluble
microbial products and extracellular polymeric substances.
Furthermore, an enhancement of the organic micropollutant
removal was achieved with an innovative plant configuration based
on an upflow anaerobic sludge blanket (UASB) reactor coupled to a
hybrid aerobic MBR at ambient temperature and low HRT. This
process demonstrated to be a sustainable and robust system which
achieved high COD removal performances and better micropollutant removal efficiencies than conventional technologies
(Alvarino et al., 2016b). The use of biofilm surfaces in a hybrid
process seems interesting for the enhancement of the removal of
organic micropollutants in small WWTPs, especially if they have to
be extended in order to improve nutrient removal. Besides, the cost
of such process should be moderated compared to an additional
treatment as activated carbon adsorption.
Finally, due to the increasing interest in using enzymes to
degrade micropollutants from wastewater, some novel processes of
enzymatic treatment have been suggested, combining filtration
and enzyme reactors. Ba et al. (2014) proposed a hybrid bioreactor
(HBR) containing cross-linked enzymes aggregates of laccase
combined with polysulfone hollow suspended fibers operated
continuously to remove three pharmaceuticals (ACE, CBZ, and
MFA). Synergistic action of the microfiltration and cross-linked
enzymes aggregates of laccase (CLEA-Laccase) achieved significant eliminations from aqueous solution. The HBR demonstrated
elimination rates up to 93% after 72 h for CBZ and near complete
elimination was achieved within 24 h of treatment for ACE and
MFA. Furthermore, Nguyen et al. (2015) evaluated the laccase-
catalyzed degradation of 30 TrOCs using an EMR equipped with
an ultrafiltration membrane. Using this process, phenolic compounds were more effectively eliminated than the non-phenolic
ones due to the formation of a dynamic layer of laccase over the
membrane surface which facilitated their subsequent enzyme
degradation.
3.2. Effects of operating conditions on removal efficiency
Table 3 (Appendix A: supplementary data) presents the efficiency of some hybrid process on the removal of organic micropollutants found in selected studies.
3.2.1. Effects of HRT and SRT
Contrary to conventional treatments, the influence of process
parameters such as HRT and SRT was not often evaluated in the case
of newly developed processes. Although for hybrid systems
including biofilm, the evaluation of the SRT is harder than in CAS,
the biofilm biomass typically has a higher age than the suspended
biomass, and the biodiversity in the biofilm is enhanced. The COD
load per surface of biofilm should be an important parameter to
monitor for the evaluation of a biofilm system.
Only a few studies used hybrid processes with different HRTs or
SRTs, but studying the variation of these parameters did not appear
as the aim of the study. For instance, Falås et al. (2012) evaluated
the removal of DCF, KPF, GFZ, NPX, IBP, MFA, and CFA, whose results
are collected in Table 3, using suspended biofilm carriers in order to
compare the removal rates of theses micropollutants per unit of
biomass to the removal rates obtained with a nitrifying activated
sludge sampled from different WWTPs. Four of the seven selected
WWTPs are using MBBR treatment operated at different HRTs
(from 6e7 h to 35 h) to remove micropollutants. Usually typical
aerobic HRTs for nitrifying activated sludge processes are around
5e10 h and around 2e4 h for MBBR processes. Results showed in
the case of several pharmaceuticals that considerably higher
removal rates can be expected with MBBR processes compared to
nitrifying activated sludge processes. All the selected compounds
were removed faster from wastewater using low HRT in MBBR
treatment (complete removal was achieved after 5 h for KPF, GFZ,
NPX, IBP and more than 60% was achieved after 10 h for DCF, MFA,
and CFA). Falås et al. (2012) suggested that high sludge ages and
microbial adaptation to the substrate gradients in biofilms could
favor degradation of some pharmaceuticals.
Furthermore, Di Trapani et al. (2013) studied organic matter
removal and nitrification using a hybrid MBBR fed with municipal
wastewater previously subjected to primary clarification. This
process was operated at different values of the mixed liquor SRT
and temperature in order to highlight the influence of these parameters. The authors hypothesized that nitrification could be
maintained at far lower SRT's than in conventional activated sludge
systems and under the application of high organic loading rates.
The pilot plant showed very high nitrification activity and was
capable of removing the organic matter at loading rates up to
3 kg.TCOD.m 3.day 1. Thanks to ammonia uptake rate batch tests,
an increase of biofilm nitrification activity was observed when the
mixed liquor SRT decreased. Results suggested that the hybrid
reactor should be run under low mixed liquor SRT values in order to
enhance ammonium removal efficiency, thus confirming that
nitrification could be maintained at far lower SRT's than in CAS
systems.
The influence of process parameters such as SRT and HRT has
been scarcely studied for hybrid processes and further researches
seem necessary in order to confirm and complete the results suggested by conventional processes' investigations. However, the
biomass retention time in biofilm systems is not easily controlled
C. Grandclement et al. / Water Research 111 (2017) 297e317
even if low loaded biofilm processes tend to favor slow-growing
bacteria, which seems promising for the pharmaceutical removal.
A shorter reaction time in the biofilm reactor is nonetheless
compensated by a higher biomass concentration and/or a higher
micropollutant removal capacity per unit of biomass.
3.2.2. Effect of the DO concentration
Biofilm reactors produce an effluent with different particulate
characteristics compared to activated sludge, in terms of floc
structure, particle size distribution, and so on. Some studies have
shown that a too strong aeration can have an influence on biofilm
breakage and can promote the formation of colloidal particles
which could enhance membrane fouling (Leiknes and Ødegaard,
2007). However, redox conditions within the biofilm may also
have an influence on the removal of different micropollutants using
attached-growth processes. Indeed, if controlled properly,
attached-growth processes can lead to different redox conditions at
different thicknesses within the biofilm. The coexistence of oxic
and anoxic conditions in the overall biofilm volume can facilitate
nutrient removal, and enhance the elimination of a wider spectrum
of micropollutants. For instance, oxic conditions prevailing at the
surface of the biofilm and among free biomass, improve the
removal of molecules such as NPX, EE2, ROX, and ERY. On the
contrary, anoxic conditions prevailing in the depth of the biofilm,
favor the degradation of molecules such as CBZ, CFA, DCF, and
iodinated X-ray contrast media such as tri-iodinated benzene dearez
rivatives (Drewes et al., 2001; Zwiener and Frimmel, 2003; Su
et al., 2010). Falås et al. (2013) noticed that the anoxic and oxic
conditions successively applied during nitrogen removal cycle
affected the micropollutant removal capacity. Some compounds
such as BZF, atenolol, CLA could be removed under both oxic and
anoxic conditions whereas other compounds were only removed
under oxic conditions (KPF, METOP, MFA, or valsartan). KPF, MFA,
and valsartan were degraded faster by the attached biomass than
the suspended biomass, but it was the opposite for METOP and 4-/
5-methylbenzotriazole (see Table 3). The rate constants obtained
for these selected micropollutants indicate that the presence of
available molecular oxygen is critical for the degradation of several
micropollutants.
Furthermore, an integrated process comprising of an anaerobic
pre-treatment before an aerobic process may be an alternative to
enhance micropollutant removal. Alvarino et al. (2016b) investigated the fate of 16 TrOCs in an integrated anaerobic/aerobic process. During 6 months of operation an UASB reactor coupled to a
hybrid aerobic MBR showed promising results compared to a
conventional process (see Table 3). CBZ, DZP, DCF, EE2, and fluoxetine were poorly removed (<40%), E1 was recalcitrant under
anaerobic operation (<20%), but well removed during aerobic step
(>65%), while some molecules such as AHTN and ADBI were
significantly removed by the UASB reactor (about 50%). Regarding
degradation pathways, biotransformation seemed to be the main
removal mechanism except for musk fragrances.
In sum, in addition to being substance specific and dependent
on the composition of the biomass, micropollutant degradation is
also dependent on the redox conditions. The degradation capacity
can differ significantly between the suspended and attached
biomass in hybrid biofilm/activated sludge processes.
3.2.3. Effects of pH and temperature
Di Trapani et al. (2013) investigated the removal of organic
matter and nitrification through a MBBR process using different
SRT values and different temperatures (between 10 and 14 C).
Their results showed that the use of this process under low mixed
liquor SRT values and low temperatures can achieve a high
ammonium removal efficiency, since a large part of nitrification
311
activity will take place in the slow growing biofilm. Temperature
plays a key role on the nitrification activity, even if under low
temperatures, the increased oxygen solubility could likely hinder
the drop in nitrifiers biological activity.
To date, the influence of pH and temperature on the micropollutant biodegradation using hybrid processes has been very
scarcely examined. Further investigations have to be undertaken to
support the conclusions found using conventional processes, or to
complete and expand the current knowledge.
3.3. Effects of microorganism communities or enzymes extracted
from microorganisms on removal efficiency
As for the bioreactor configuration, a few studies tend to assess
what are the best types of microorganisms to remove some given
organic micropollutants. Some of them used activated sludge to
form a suspended biofilm, while others tried to use enzymes produced by WRF and combined with activated sludge.
Table 4 (Appendix A: supplementary data) presents the removal
of selected micropollutants using hybrid bioreactors, depending on
microorganism communities or enzyme extracted from
microorganisms.
3.3.1. Biofilm
Today's knowledge on micropollutant and specially PhAC
removal using biofilm systems is rather limited. However, Zwiener
and Frimmel (2003) investigated the biodegradation of three active
pharmaceuticals using a biofilm reactor formed from activated
sludge biomass during a 48 h-period. The biodegradation obtained
for CFA, IBP, and DCF using an oxic biofilm reactor was close to the
one obtained using a reference pilot activated sludge plant. With
the reference pilot plant, CFA and DCF were not eliminated (about
5%), whereas the concentration of IBP was decreased to approximately 35%. On the contrary, using the anoxic BFR, all three substances, showed elimination resulting in a decrease of their
concentration to values between 60 and 80% of their initial concentration (see Table 4).
Moreover, Paje et al. (2002) evaluated the degradation of DCF by
a river biofilm. Degradation was possible after acclimatization.
Adapted biofilms showed that a removal of 10e25% of the initial
concentration could be achieved within 4 days. Besides, the results
showed that DCF can inhibit many microorganisms such as Staphylococcus epidermidis (Perilli et al., 2000) that would usually
compromise a lotic biofilm. Indeed, DCF disrupted normal biofilm
development in lotic systems, while some microorganisms such as
Cytophaga-Flavobacterium were able to survive and even to degrade
this compound.
Still little is known about the biomass capacity to remove
pharmaceuticals in biofilm systems and whether this capacity differs from that of activated sludge.
3.3.2. Activated sludge and suspended biofilm carriers
The acidic pharmaceutical removal during batch experiments
using activated sludge and suspended biofilm carriers (AnoxKaldnes™ type K1 media) were compared by Falås et al. (2012).
Similar removal rate constants for IBP (around 2e5 L.g 1 of biomass.d 1) and NPX (around 0.5e1 L.g 1 of biomass.d 1) were found
in both biofilm carriers and activated sludge biomasses, whereas
significant higher rate constants for DCF, KPF, GFZ, CFA, and MFA
were found with the carriers biomass (0.06e0.38, 0.9e3.6, 0.6e2.1,
0.05e0.17 and 0.08e0.48 L.g 1 of biomass.d 1, respectively), as
compared to the activated sludge biomass (0e0.02, 0.01e0.32,
0.01e0.27, 0e0.04 and 0e0.06 L.g 1 of biomass.d 1, respectively).
In their subsequent study, Falås et al. (2013) evaluated the efficiency of a hybrid suspended/attached growth process obtained by
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C. Grandclement et al. / Water Research 111 (2017) 297e317
combination of biofilm carriers and activated sludge. In most cases,
considerably higher micropollutant removal rates were observed
for the biofilm compared to the free biomass. This study confirmed
that a reactor with a fixed biomass achieved rapid removals for DCF
(1.3e1.7 L.g 1 of biomass.d 1) and TMP (1.0e3.3 L.g 1 of biomass.d 1), while the elimination of both compounds in the
suspended-free biomass reactor was insignificant (0.1 L.g 1 of
biomass.d 1) (Falås et al., 2013). Results of this study demonstrated
that the degradation rate of organic micropollutants in biological
wastewater treatment is substance specific and dependent on the
composition of the biomass.
Casas et al. (2015) also evaluated the ability of a combination of
suspended activated sludge and biofilm on the removal of different
micropollutants from hospital wastewater using three MBBR in
series. The authors noticed that the degradation of these micropollutants occurred in parallel with the removal of COD and nitrogen which suggest a co-metabolism pathway. Besides, the
efficiency of each MBBR reactor was also evaluated. While the
amount of biomass was decreasing from the first to the last reactor,
the specific activities (Kbio) of the biomass, which are the removal
rate constants corrected by the amount of biomass per reactor
volume, were increasing along the reactors succession. In a sub Casas et al. (2015) evaluated the efficiency of
sequent study, Escola
a pilot plant consisting in a series of one activated sludge reactor,
two hybrid processes, and one MBBR during 10 months under
continuous operation. Results, showed that removal of organic
matter and nitrification mainly occurred in the first reactor which is
well designed for COD or nitrogen removal and other compounds
that are easily degraded by activated sludge biomass. Pharmaceuticals were globally removed efficiently, as revealed in Table 4.
Batch experiments showed highest removal rate constants of the
pharmaceuticals in the activated sludge reactor. However, during
the continuous flow experiments, a concentration increase of
compounds such as CBZ, venlafaxine, METOP, or SMX was observed
in the first reactor with activated sludge. This phenomenon can
occur due to a de-conjugation by bacterial enzymes of the compounds formed by sulfation, glucuronidation and acetylation during phase II of human metabolites, and eliminated via urine or
feces. Another possibility may be the transformation of metabolites
from other parent compounds (Kovalova et al., 2012). Besides, a
better removal (close to 20%) was noticed for these compounds in
the other reactors containing activated sludge and biofilm carriers,
which improved the amount of biomass per reactor volume.
The micropollutant removal rates obtained by Luo et al. (2014a)
using a continuous bench-scale MBBR was of the same order of
magnitude than the ones obtained with classical processes (activated sludge and MBR). IBP, salicylic acid (SLA), PRM, and NPX were
efficiently eliminated using this particular process (93.7%, 91.1%,
83.5%, and 81.1% respectively). The high removal efficiency could be
ascribed to the presence of strong electron donating (readily
biodegradable) functional groups (e.g., eOH) on these compounds.
KPF, ACE, metronidazole, and GFZ were well removed (50.0e75.0%)
by the MBBR, while DCF and CBZ were resistant to the MBBR
treatment. The average removal of DCF by the MBBR was only 45.7%
and CBZ showed an even lower removal of 25.9%. A subsequent
study Luo et al. (2015) showed that the MBBR-MBR coupled system
had lower fouling tendency than a conventional MBR, and the
compound-specific removal efficiencies varied significantly
ranging from 25.5 to 99.5% with a HRT of 24 h (see Table 4). Previous batch experiments using non-acclimatized and acclimatized
(for attached microbial growth) sponge biomasses showed that
several micropollutants can be adsorbed on non-acclimatized
sponge cubes, and that acclimatized sponge can improve the
removal of some of the less hydrophobic (log Kow < 2.5) compounds
like CBZ (Luo et al., 2014a). Besides, the removal efficiency achieved
by the MBBR depends on physicochemical properties of the tested
compounds, but the obtained degradation is comparable with other
conventional processes. BPA, E1, E2, EE2, 4-n-NP, 4-tert-octylphenol, and TCS were considerably eliminated (>80.0%) during the first
two hours in the experiments with either non-acclimatized sponge
or acclimatized sponge. Thus, sorption played a significant role in
the removal of these compounds. ACE, DCF, GBZ, IBP, KPF, NPX, and
SLA were hardly removed (mostly <20%) with non-acclimatized
sponge, but showed markedly improved reduction when acclimatized sponge was used.
3.3.3. Enzymatic treatment
Only a handful of studies have investigated TrOC removal in
EMRs operating in continuous flow. Ba et al. (2014) evaluated the
removal, collected in Table 4, of three pharmaceuticals ACT, CBZ,
and MFA using microfiltration alone and a combination with CLEALac. The MF alone showed significant removals of the three compounds in the filtrate varying approximately from 50 to 90% after a
time-period of 8 h. Synergistic action of the MF and CLEA-Lac
during operation achieved eliminations from aqueous solution up
to 85% for ACT, MFA, and CBZ, of around 99% for ACT and nearly
100% for MFA. Under continuous operation, the HBR demonstrated
elimination rates of the drugs from filtered wastewater up to 93%
after 72 h for CBZ and near complete elimination of ACT and MFA
was achieved within 24 h of treatment. Besides, the TrOCs removal
efficiencies of EMRs depend on some factors such as the chemical
structure of the targeted compounds. Nguyen et al. (2015) investigated the removal of 30 TrOCs using an EMR equipped with a
nanofiltration membrane. They noticed that phenolic compounds
were more effectively removed than the non-phenolic ones due to
the formation of a dynamic layer of laccase over the membrane
surface. Thus, TrOCs were retained and their degradation was
facilitated. The addition of a redox-mediator (SA or HBT) to the EMR
significantly improved the TrOC degradation. In a subsequent
study, Nguyen et al. (2016b) investigated the removal of 14 phenolic
and 17 non-phenolic compounds using an EMR with different
TrOCs concentration values under SA loadings. The evaluation of
the toxicity of laccase, SA, TrOCs, and treated effluent was also
investigated by the authors. Results showed that 10 mM of SA
addition could improve TrOCs removal. However, a high concentration of SA (50 or 100 mM) did not show significant improvements
regarding TrOCs removal, but increased effluent toxicity, due to the
presence of unconsumed SA and radicals generated from SAoxidation by laccase. In parallel, Nguyen et al. (2016a) studied the
degradation of four micropollutants in a packed-bed enzyme
reactor using laccase immobilized on granular activated carbon.
Results of this investigation showed high removals for all studied
compounds, as described Table 4 (up to 90% after 24 h). Besides,
since enzyme immobilization seems to be a good option for longterm operational stability, Ji et al. (2016) used a membrane hybrid
reactor with T. versicolor laccase immobilized on suspended biocatalytic TiO2 nanoparticles to investigate CBZ removal. Even if the
highest ratio of 71% within 96 h was observed using optimized
operating conditions, and that the toxicity of CBZ was also removed,
more improvements on this hybrid process and studies on the CBZ
degradation at environmentally relevant concentration are still
required.
It is clear that further investigations are needed to advance in
the design of EMRs, in particular to demonstrate the viability of
n et al. (2015) focused
such process at full-scale in WWTPs. Abejo
their study on the evaluation of the economic aspects of EMR based
on laccase immobilized over ceramic membranes and applied to
the degradation of antibiotics. Results from a mathematical cost
estimation model showed that the process is still far from economic
competitiveness because of membrane conditioning costs. To
C. Grandclement et al. / Water Research 111 (2017) 297e317
achieve competitive economical results, some improvements on
enzymatic activity, on the effective lifetime of the enzymatic reactors, and on membrane conditioning or regeneration costs have
to be made.
3.4. Limits of hybrid processes
Studies about the efficiencies of a hybrid process to remove
micropollutants are recent and further studies are needed in order
to fill the gap regarding the influence of hydraulic parameters as
suggested by Ba et al. (2014). On the one hand, a proper choice of
the main operating parameters, such as pH, HRT, or temperature
might lead to a substantial improvement of the hybrid process
performances. On the other hand, further researches should target
the evaluation of the costs associated to the functioning of such
processes. Indeed, the optimization in terms of technical and
economical competitiveness of such water treatment processes
could lead to the emergence of environmentally and economically
sustainable water treatment processes, even though improvements
regarding hybrid processes are still needed in order to maximize
the removal of some of the more recalcitrant micropollutants.
Casas et al. (2015) suggested to add a complementary
Escola
advanced process to the treatment such as ozonation which could
break down some bonds, and thus facilitates the subsequent
removal by biodegradation. In that field, Navaratna et al. (2016)
have investigated for seven months the elimination of s-triazine
herbicide using a laboratory-scale MBR combined with ultraviolet
disinfection and sorption onto granular activated carbon. More
than 80.0% of the targeted herbicide was removed by this hybrid
MBR through the biodegradation pathway, only with different HRT:
from 1.5 to 7.5 days. Regarding pharmaceutical compounds, the
complementary effects of adsorption and enzymatic degradation
have been highlighted using granular activated carbon-bound laccase (Nguyen et al., 2016a). In a previous study, Nguyen et al.
(2013a) evaluated the removal of TrOCs by an MBR-based hybrid
treatment process using UV oxidation or nanofiltration/reverse
osmosis membrane filtration. Results confirmed that UV oxidation
is effective for the degradation of chlorinated TrOCs and TrOCs
containing a phenolic group, but less effective for the removal of
TrOCs containing an amide group such as CBZ. Only 30.0% of CBZ
was removed by UV oxidation whereas a complete removal was
achieved for pentachlorophenol and TCS. However, the hybrid
process achieved 85.0% of removal efficiency for all 22 selected
TrOCs. For instance, 96.0% of CBZ was eliminated with a contacting
time of 7.5 min. Furthermore, as it was studied by Nguyen et al.
(2013b), using a MBR, the efficiency of a hybrid process
comprising of mixed culture of bacteria and WRF could be evaluated, and the addition of a redox mediator could improve the
removal of some recalcitrant TrOCs.
Moreover, only few studies evaluated the toxicity of the effluent
after a biodegradation process using a conventional process, but
none using a hybrid process. It seems obvious that further experiments should be performed to evaluate the toxicity of by-products
after a hybrid process. Jeli
c et al. (2012a) showed, using Vibrio
fischeri luminescence reduction tests, that TrOC transformation via
WRF often leads to detoxification, but T. versicolor can, for instance,
produced 1,2-hydroxy ibuprofen, the main metabolites of IBP,
which is more toxic than the parent compound (Marco-Urrea et al.,
2009). Microtox® tests also showed that metabolites of DIU could
also be more toxic than the parent compound (Tixier et al., 2002).
The toxic effect of the DIU's metabolites was also demonstrated on
two phytoplanktonic microorganisms, the green alga Dunaliella
tertiolecta and the diatom Navicula forcipata (Gatidou and
Thomaidis, 2007). Besides, Nguyen et al. (2016b) noticed that the
use of a high dose of redox mediator such as SA can increase the
313
effluent toxicity.
4. Conclusions and perspectives
During the past decade, a relevant number of studies have
evaluated the efficiency of biodegradation processes to remove
organic micropollutants from wastewaters. No significant difference exists between CAS and MBR treatments. The two systems are
efficient to remove hydrophobic compounds and hydrophilic ones
which possess only EDGs. In contrast, the removal of hydrophilic
compounds bearing EWGs is still very low (below 20%). Besides,
few authors noticed that the use of WRF or a mixed culture of
activated sludge and WRF could improve the performances of a
MBR, but the operating conditions play a key role especially on
enzymatic activity. Thus, pH, aeration conditions, HRT, SRT have to
be optimized depending on the selected micropollutants and their
physico-chemical characteristics, e.g. hydrophobicity, chemical
structure, pKa, and so on. However, membrane fouling, recalcitrance of some hydrophilic compounds, and adsorption on activated sludge flocs are still important factors limiting the
biodegradation of such pollutants using conventional processes.
Recent studies suggested improvements regarding micropollutant
degradation using hybrid processes. These processes containing
biofilm carriers, suspended/attached growth system, or crosslinked enzymes aggregates showed better removal of micropollutants, even on recalcitrant compounds such as CBZ. Further
studies need to be performed in order to evaluate which system is
actually the more cost-benefit efficient, and to investigate the influence of operating conditions and the toxicity of effluents after
treatment as well. However, even if a lack of studies at full-scale has
been noticed, these processes could be a sustainable prospective
treatment to improve the degradation of micropollutants from
wastewaters. This could be facilitated by addition of a pretreatment
step such as ozonation.
Acknowledgements
The authors sincerely thank the French public investment bank
(Bpifrance) for its financial support.
Appendix A. Supplementary data
Supplementary data (tables 1, 2, 3 and 4) related to this article
can be found at http://dx.doi.org/10.1016/j.watres.2017.01.005.
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