Environ Sci Pollut Res (2015) 22:4383–4395 DOI 10.1007/s11356-014-3646-5
RESEARCH ARTICLE
Lab-scale experimental strategy for determining micropollutant partition coefficient and biodegradation constants in activated sludge M. Pomiès & J. M. Choubert & C. Wisniewski & C. Miège & H. Budzinski & M. Coquery
Received: 15 May 2014 / Accepted: 22 September 2014 / Published online: 11 October 2014 # Springer-Verlag Berlin Heidelberg 2014
Abstract The nitrifying/denitrifying activated sludge process removes several micropollutants from wastewater by sorption onto sludge and/or biodegradation. The objective of this paper is to propose and evaluate a lab-scale experimental strategy for the determination of partition coefficient and biodegradation constant for micropollutant with an objective of modelling their removal. Four pharmaceutical compounds (ibuprofen, atenolol, diclofenac and fluoxetine) covering a wide hydrophobicity range (log Kow from 0.16 to 4.51) were chosen. Dissolved and particulate concentrations were monitored for 4 days, inside two reactors working under aerobic and anoxic conditions, and under different substrate feed conditions (biodegradable carbon and nitrogen). We determined the mechanisms responsible for the removal of the target compounds: (i) ibuprofen was biodegraded, mainly under aerobic conditions by cometabolism with biodegradable carbon, whereas anoxic conditions suppressed biodegradation; (ii) atenolol was biodegraded under both aerobic and anoxic conditions (with a higher biodegradation rate under aerobic conditions), and cometabolism with biodegradable carbon was the main mechanism; (iii) diclofenac and fluoxetine were removed by sorption only. Finally, the abilities of our strategy were evaluated by testing the suitability of the parameters for simulating Responsible editor: Angeles Blanco M. Pomiès : J. M. Choubert (*) : C. Miège : M. Coquery Irstea, UR MALY, 5 rue de la Doua, CS70077, 69626, Villeurbanne Cedex, France e-mail: [email protected] C. Wisniewski UMR Qualisud, Université Montpellier 1, 15 Av. Ch. Flahault BP 14491, 34093, Montpellier Cedex 5, France H. Budzinski EPOC/LPTC UMR5805, Université Bordeaux 1, 351 Cours de la libération, 33405, Talence Cedex, France
effluent concentrations and removal efficiency at a full-scale plant. Keywords Modelling . Pharmaceuticals . Experimental strategy . Wastewater treatment . Activated sludge
Introduction Wastewater treatment plant (WWTP) effluents are a wellknown vector of micropollutant discharge, particularly for pharmaceutical compounds, into the environment (Heberer 2002; Ternes and Joss 2006). Ongoing regulation in Europe regularly reinforces the requirements towards discharge of chemicals into the environment, with lists of priority substances regularly revisited (EC 2008, 2013). A broad strand of research has highlighted how WWTPs could efficiently remove many micropollutants from wastewater (e.g., Miège et al. 2009; Onesios et al. 2009; Luo et al. 2014), particularly by biological processes, even though they were not originally designed for this purpose. However, discharged WWTP effluents still contain several micropollutants due to their high initial concentrations in raw wastewater and/or their low affinity with suspended solids (SS) and their non-biodegradable chemical structure (Martin Ruel et al. 2012). The identification of micropollutant removal mechanisms, as well as their modelling to predict and eventually better control their overall elimination, is essential. Sorption onto sludge and biodegradation are the two main mechanisms dictating the fate of non-volatile micropollutants in biological wastewater treatment systems (Carballa et al. 2004). Micropollutant can be biodegraded, either directly or indirectly, by cometabolism involving macropollutant as a growth substrate (basically biodegradable chemical oxygen demand (bCOD)) (Criddle 1993). Both mechanisms have been modelled using parameters usually based on lab-scale
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experiments (Pomiès et al. 2013). However, experimental conditions can vary greatly from one author to another and no unified experimental protocol is proposed in the literature, leading to possible biased modelling results (Clouzot et al. 2013). The discrepancy in operating conditions of experiments and in the sludge tested makes relatively difficult the determination of typical values for parameters for each micropollutant. Nevertheless, such set of values may be necessary for engineering application of micropollutant modelling (i.e. typical parameter values). This study aimed at giving new insight for determining partition coefficient and biodegradation constant values for modelling micropollutant removal in activated sludge process. We implemented a robust experimental strategy in lab-scale experiments involving micropollutant analysis in both dissolved and particulate phases of biological sludge. The protocol integrated different conditions of nitrifying/denitrifying activated sludge process, i.e. oxido-reduction conditions and presence or absence of biodegradable macropollutants (biodegradable carbon, nitrogen). Our experiments aimed to (i) determine the predominant removal mechanism (sorption or biodegradation) and the most favourable conditions of removal for a selection of micropollutants; (ii) determine the compartment (dissolved or particulate) involved in biodegradation; and (iii) quantify the values of two modelling parameters, the partition coefficient (Kd) and the biodegradation constant (k). Then, we tested the measured parameter values to simulate micropollutant removal in a full-scale activated sludge process. Our work targeted four pharmaceutical compounds (ibuprofen, atenolol, diclofenac, fluoxetine) that share different physicochemical properties and different fates through WWTPs.
Partition coefficient and biodegradation constant: experimental determination Current existing protocols
Two compartments are usually defined for sorption onto sludge: the dissolved compartment (Smp) and the particulate one (Xmp). For each micropollutant, a partition coefficient (Kd) is defined to account for the main sorption mechanisms, i.e. hydrophobic mechanisms and electrostatic interactions (Sipma et al. 2010). It is calculated as follows (Eq. 1):
Kd
Smp SS
=SS
Xmp
Smp
ð1Þ
Partition coefficient of micropollutant onto sludge [L.gSS−1]
Concentration of micropollutant in particulate phase [ng.L−1] Concentration of micropollutant in dissolved phase [ng.L−1] Suspended solids concentration of sludge in biological reactor [gSS.L−1]
Lab-scale experiments are generally used to determine Kd. A common procedure consists in spiking sludge with micropollutant at defined concentrations and then measuring the dissolved and/or particulate micropollutant concentration after a defined short time considering sorption equilibrium is reached rapidly (Ternes et al. 2004; Urase and Kikuta 2005; Lindblom et al. 2009; Abegglen et al. 2009; FernandezFontaina et al. 2014). However, although widely accepted, these procedures carry weaknesses (Stevens-Garmon et al. 2011; Pomiès et al. 2013). First, the mass balance is rarely verified because particulate concentration is often not measured, but simply deduced from dissolved concentration (Hyland et al. 2012). Second, in sorption studies, micropollutant biodegradation is generally inhibited by diverse procedures (Maurer et al. 2007; Wick et al. 2009). However, the biodegradation inhibition may induce potential bias in Kd measurement, as it may influence sorption performances by altering the physicochemical properties of sludge (Stevens-Garmon et al. 2011). Biodegradation The biodegradation rate of a micropollutant is often described with a pseudo-first-order kinetic (Joss et al. 2006; Wick et al. 2009). It is calculated as follows (Eq. 2): dSmp ¼ −k Smp SS ð2Þ dt biodegradation k
Sorption
Kd ¼
Xmp
Biodegradation constant [L.gSS−1.day−1]
The biodegradation process also suffers from a lack of standardization. Biodegradation of micropollutant is usually evaluated from the time-course evolution of micropollutant concentrations after spiking (Lindblom 2009; Plosz et al. 2010). According to the authors, spiking can consist to add real raw wastewater (Plosz et al. 2010) or micropollutants mixed with synthetic substrate (Suarez et al. 2010). The biodegradation constant (k) is mostly determined for micropollutant measured in the dissolved phase only (Wick et al. 2009), while micropollutant in the particulate phase is seldom measured (Urase and Kikuta 2005). Only few studies investigated the effect of oxido-reduction potential (ORP) conditions on biodegradation efficiencies: a significant
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decrease has been shown under anoxic conditions for ibuprofen, estrone and estriol (Joss et al. 2004; Plosz et al. 2010; Suarez et al. 2010), but the difference still needs to be determined for other pharmaceutical compounds.
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Materials and methods Experimental strategy WWTP sampling
Proposed experimental protocol In the proposed experimental strategy, we studied sorption and biodegradation in the same reactor, considering that sorption occurs rapidly and biodegradation takes place at a longer time scale (cf. “Current existing protocols”). As a consequence, we did not need add any biodegradation inhibitor which could alter sludge characteristics. We based our protocol on the analysis of dissolved and particulate phases. Indeed, for a micropollutant that may be present in both phases (i.e. Smp and Xmp, respectively), this is a key issue for understanding its fate. Calculating Kd with Smp and Xmp is more reliable than using the Smp decrease only, as it enables to verify mass conservation. Moreover, it allows pinpointing where the biodegradation reaction occurs. A current assumption in the literature is to consider that only the micropollutant in the dissolved phase is biodegradable (Lindblom et al. 2009; Plosz et al. 2010). Experimental data from our protocol enabled to confirm or invalidate this assumption, particularly for micropollutants known to have some affinity with organic matter and thus can be retained on it. Our experimental strategy, as detailed in the section “Experimental strategy”, also aimed to define the influence of operating conditions, i.e. the nature of the available biodegradable substrate and the ORP conditions. As shown on Table 1, we defined biodegradation rates separately for dissolved (with an S subscript) and particulate phases (with an X subscript) under aerobic (with an Ox subscript) and anoxic (with an Ax subscript) conditions. For each micropollutant, we considered different biodegradation constants (k) according to the absence or the presence of macropollutant as growth substrate (biodegradable carbon, nitrogen as ammonium under aerobic conditions and nitrates under anoxic conditions). This formalism dissociates cometabolism with biodegradable carbon and nitrogen (action of heterotrophic biomass), cometabolism with nitrogen only (action of autotrophic biomass), and the endogenous process. Biodegradation constants used in the biodegradation equations are easier to measure than the parameters of Criddle’s formalism (Criddle 1993; Delgadillo-Mirquez et al. 2011; Fernandez-Fontaina et al. 2014). Pseudo-first-order kinetics were used (Table 1), like in Eq. 2, with the frequently used Monod kinetics functions (switch functions taking values between 0 and 1) to modulate the degradation rate according to the oxygen concentration, the biodegradable COD concentration, and the ammonium concentration.
Sludge and wastewater samples came from a domestic WWTP (2,900 population equivalents (P.E.) based on 60 gBOD5 P.E.−1 as suggested by EC (1991)) consisting in a nitrifying/denitrifying activated sludge process and a secondary clarifier. Samples were collected in June 2011, where the activated sludge process was operated at 50-day solid retention time (hydraulic retention time (HRT) of 4 days), with around 7.3 gSS.L−1 and a temperature of around 23 °C. Raw and treated wastewater samples were collected to constitute 24-h flow proportional composite samples using a refrigerated automatic sampler with cleaned Teflon tubing and glass containers (Choubert et al. 2011). Micropollutants in dissolved and particulate phases were analysed in raw wastewater, whereas only the dissolved phase was analysed in treated wastewater due to the low concentration of suspended solids (7 mgO2.L−1), whereas reactor B was maintained under anoxic conditions. The experiment was carried out for 4 days. The reactors were 200-L plastic tanks, continuously stirred. A subsample was regularly collected in each reactor via a valve port fitted at the bottom of the tanks. In each reactor, we collected a total of nine samples for the analysis of the micropollutant dissolved (Smp) and particulate (Xmp) concentrations, biodegradable carbon (i.e. biodegradable chemical oxygen demand (bCOD) concentration), and nitrogen (i.e. ammonium (NH4-N) and nitrate (NO3-N) concentrations). The experimental protocol was defined according to the constraints relative to on-site production of filtered samples, analytical capability, and expected timecourse evolution of the concentrations versus the sampling and analytical uncertainty. In each reactor, four periods (phases) were applied: a preliminary phase and three other phases corresponding to
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Table 1 Equations involved in biodegradation proposed in this work Conditions
Smp
Xmp
Aerobic
Presence of biodegradable carbon and ammonium Presence of ammonium only
−1 −1
Absence of biodegradable substrate
−1
Presence of biodegradable carbon and nitrates Presence of nitrates only
−1
Absence of biodegradable substrate
−1
Presence of biodegradable carbon and ammonium Presence of ammonium only
−1
Absence of biodegradable substrate
−1
Presence of biodegradable carbon and nitrates Presence of nitrates only
−1
Absence of biodegradable substrate
−1
Anoxic
Aerobic
Anoxic
−1
−1
−1
Reaction rates
B k S;C–N;Ox Smp SS SO þKSOO2;OHO SBSþK 1 SNHx 1 SBKþK k S;N;Ox Smp SS SO þKSOO2;OHO SNHx þK NHx;ANO 1 K2 1 SNHx;ANO k S;endo;Ox Smp SS SO þKSOO2;OHO SBKþK þK2 1 KO2;OHO SNOx B SBSþK SO þK k S;C–N;Ax Smp SS SNOx þK NOx;OHO 1 O2;OHO KO2;OHO SNOx SNHx 1 k S;N;Ax Smp SS SNOx þK SNHx þK SO þK SBKþK NOx;OHO NHx;ANO O2;OHO 1 KO2;OHO SNOx K1 K2 k S;endo;Ax Smp SS SNOx þK KO2;OHO þSO K1 þSB K2 þSNHx NOx;OHO SO SB k X;C−N;Ox Xmp SS SO þKO2;OHO SB þK1 SNHx K1 k X;N;Ox Xmp SS SO þKSOO2;OHO SNHx þK SB þK1 NHx;ANO SO K1 K2 k X;endo;Ox Xmp SS SO þKO2;OHO SB þK1 SNHx þK2 KO2;OHO SNOx SB k X;C−N;Ax Xmp SS SNOx þK S þK S þK NOx;OHO B 1 O O2;OHO KO2;OHO SNOx SNHx K1 k X;N;Ax Xmp SS SNOx þKNOx;OHO SNHx þKNHx;ANO SO þK S þK O2;OHO B 1 KO2;OHO SNOx K1 K2 k X;endo;Ax Xmp SS SNOx þK KO2;OHO þSO K1 þSB K2 þSNHx NOx;OHO
So dissolved oxygen concentration, KO2,OHO oxygen half-saturation coefficient for heterotrophic biomass, SNHx dissolved ammonia nitrogen concentration, KNHx,ANO ammonia half-saturation coefficient for autotrophic biomass, SNOx dissolved nitrate nitrogen concentration, KNOx,OHO nitrate halfsaturation coefficient for heterotrophic biomass, SB dissolved biodegradable COD, K1, K2 half-saturation coefficients for heterotrophic biomass, kS,C-N,Ox, kX,C-N,Ox biodegradation constants for Smp and Xmp in presence of biodegradable COD and nitrogen under aerobic conditions, kS,C-N,Ax, kX,C-N,Ax biodegradation constants for Smp and Xmp in presence of biodegradable COD and nitrogen under anoxic conditions, kS,N,Ox, kX,N,Ox biodegradation constants for Smp and Xmp in presence of nitrogen only, under aerobic conditions, kS,N,Ax, kX,N,Ax biodegradation constants for Smp and Xmp in presence of nitrogen only, under anoxic conditions, kS,endo,Ox, kX,endo,Ox biodegradation constants for Smp and Xmp in absence of biodegradable substrate and aerobic conditions, kS,endo,Ax, kX,endo,Ax biodegradation constants for Smp and Xmp in absence of biodegradable substrate and anoxic conditions
different macropollutant supply conditions (phase I, no macropollutant addition; phase II, biodegradable carbon and nitrogen addition; phase III, nitrogen addition). The different phases of the experiment are presented in the following sections and illustrated in Fig. 1 (using the example of reactor A). The pH was practically stable during the experiments, with 8.3±0.2 (reactor A) and 8.1±0.1 (reactor B), due to buffer effect or addition of calcium carbonates. Preliminary phase (t=0 to 18 h) Reactors A and B were filled with 180 L of sludge collected in the WWTP aeration tank (t=0 h). Intermittent aeration was applied for 18 h to attempt to eliminate possible nonbiodegraded residual substrate contained in the liquid sludge. The initial micropollutant concentration in the sludge (Xmp) was measured at t=0 h. Phase I “without any substrate” (t=18 to 43 h) At t= 18 h, each reactor was spiked with a mixture of micropollutants (10 μg.L−1 each) dissolved in 10 mL of
methanol solvent. These levels are in the usual high concentration range of raw wastewater (Miège et al. 2009). ORP conditions were monitored in both reactors. Reactor A was supplied with air via a diffuser located at the bottom of the reactor. In reactor B, N2 gas was supplied to initiate anoxic conditions. No chemical agent (e.g. inhibitor) was used. In each reactor, sludge was sampled three times at t=18.5 h (for dissolved and particulate analysis), at t=23 h (for dissolved analysis, reactor B only) and at t=42 h (for dissolved analysis). Phase I aimed at determining partition coefficient and biodegradation of micropollutants without biodegradable carbon and nitrogen conditions (endogenous conditions). Phase II “with biodegradable carbon and nitrogen substrates” (t=43 to 66 h) At t=43 h, a volume of 50 L of raw wastewater was added to each reactor (raw wastewater to sludge ratio 1:4 v/v). Raw wastewater was a source of biodegradable substrate (bCOD= 175 mg.L−1, NH4-N=72 mg.L−1). In reactor B, we also added NaNO3 (NO3-N=18 mg.L−1). Suspended solids concentration in both reactors was 5.4 gSS.L−1 after raw wastewater
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Sampling for Smp analysis Sampling for Xmp analysis
Xmp and Smp [ng.L-1]
Preliminary phase
Phase I
Phase II
kS,endo,Ox
Phase III
kS,C-N,Ox kS,N,Ox
Kd
kX,endo,Ox
0
18 18.5
Sludge addition (180 L)
Micropollutant spiking (10 µg.L -1)
23
kX,C-N,Ox
42 43 43.5
47.5
Raw wastewater addition (50 L)
kX,N,Ox
66 66.5
70
71.5
Time (h)
NH4-N addition (40 mgN.L-1)
Fig. 1 Successive phases applied to reactor A with sampling type (for analysis of dissolved Smp and particulate Xmp micropollutant concentrations) and the sorption and biodegradation parameters
addition. In each reactor, sludge was sampled at t=43.5 h (for dissolved and particulate analysis) and t=47.5 h (for dissolved analysis). Phase II aimed at determining biodegradation of micropollutants under biodegradable carbon and nitrogen conditions. Phase III “with nitrogen substrate only” (t=66 to 72 h) At t=66 h, we increased the NH4-N concentration in reactor A to 40 mgN.L−1 by adding ammonium chloride (NH4Cl) to ensure non-limiting nitrogen substrate conditions and to stimulate nitrification. In reactor B, we increased the NO3-N concentration to 30 mgN.L−1 by adding sodium nitrate (NaNO3) to stimulate denitrification. In each reactor, sludge was sampled at t=66.5 h (for dissolved analysis), 70 h (for dissolved analysis) and 71.5 h (for dissolved and particulate analysis). Phase III aimed at determining biodegradation of micropollutants under nitrogen-only conditions. Studied micropollutants and chemical analysis Four pharmaceutical compounds (ibuprofen, atenolol, diclofenac and fluoxetine) were chosen based on their differing physicochemical properties as shown in Table 2. Sorption onto sludge may differ due to varying log Kow values (0.16 for
atenolol to 4.51 for diclofenac) and pKa values (from 4.15 for diclofenac to 10.1 for fluoxetine) (Suarez et al. 2008). Also, biodegradation potential may be influenced by the molecular feature of the micropollutants, as some of the functional groups are known to favour their biodegradability (e.g. carboxylic acid, amide or hydroxyl groups and phenyl ring), whereas other groups increase their resistance to biodegradability (e.g. extensive alkyl chain branching or halogens) (Boethling et al. 2007). Ibuprofen is usually removed at more than 90 % by biodegradation (Joss et al. 2005; Soulier et al. 2011). Atenolol is also removed by biodegradation, but at a lower level (50–70 %) (Jelic et al. 2011). Removal efficiency of diclofenac is discrepant according to the studies (from 5 to 80 %); Jelic et al. (2011) considered that biodegradation is the main removal mechanism but sorption can occur also. Fluoxetine is partly removed by sorption (33 and 50 %) (Radjenovic et al. 2009; Soulier et al. 2011) and/or biodegradation (Suarez et al. 2012). For the analysis of the dissolved phase of the sludge from lab-scale experiments, we sampled and filtered directly in situ with a filtration module (polysulfone, 0.1 μm, 0.16 m2, Polymem, France). Filtrates were analysed 24 h after sampling (or frozen before analysis). Concerning the dissolved phase of raw and treated wastewater samples from full-scale experiments, samples were filtered in laboratories less than 24 h after
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Table 2 Physicochemical properties (log Kow) and limits of quantification (LQ) of the studied pharmaceutical compounds and literature data on removal efficiency (RW) through activated sludge process Micropollutant (family)
Ibuprofen (antiinflammatory) Atenolol (beta-blocker) Diclofenac (antiinflammatory) Fluoxetine (antidepressant)
Formula
Molecular weight [g/mol]
pKa
*
Log Kow
Log Dow a
LQ in dissolved phase [ng.L-1]
LQ in particulate phase [ng.L1]
C13 H18 O2
Removal efficiency by activated sludge process found in the literature (RW , wastewater line) [%]
S uggested removal mechanism in literature S orption
Biodegradation
(3)
Molecular features influen cing biodegradability Positive effect (increase) Carboxylic acid group (COOH)
Negative effect (decrease ) Extensive alkyl chain branching
206
4.91
3.971,2
0.68
1
2
90 99 (4) 98 (5)
-
X (9)
266
9.67
0.161
-1.32
2.5
1.5
71 (6) 61 (4) 61 (5)
-
X (10)
Amide group and hydroxyl group (OH)
Extensive alkyl chain branching
296
4.15
4.511,2
0.46
0.5
1
69 (7) 5-80 (8) 22 (4) 9 (5)
X (11)
X (12)
Carboxylic acid group (COOH)
2 chlorine atoms
309
9.8
4.052
2.44
1
1
33 (4) 50 (5)
X (13)
X (13)
Phenyl ring
3 fluorine atoms
C14H 22N2O3
C14H 11Cl2NO 2
C17H 18F3NO
14
Verlicchi et al. (2012); 2 Hyland et al. (2012); Ternes et al. (2004); 4 Radjenovic et al. (2009); 5 Soulier et al. (2011); Carucci et al. (2006); 7 Ternes (1998); Paxeus (2004); Salgado et al. (2012); 10 Maurer et al. (2007); 11 Suarez et al. (2012); Suarez et al. 2008; Fernandez-Fontaina et al. (2012); 14 Boethling et al. (2007) a
Calculated for acids with equation log Dow=log Kow−log(1+10(pH−pKa) ) and for basis with log Dow=log Kow−log(1+10(pKa–pH) ) using the mean pH of activated sludge in reactors (=8.2), pKa and log Kow of each micropollutant
b
From http://chem.sis.nlm.nih.gov/chemidplus/ (for ibuprofen, diclofenac), or from http://www.drugbank.ca (for atenolol and fluoxetine)
sampling on fibreglass filters (GF/F, 0.7 μm). For particulate phase from full- and lab-scale experiments, sludge samples were centrifuged in laboratories within 24 h after sampling; centrifuged sludge and suspended solids collected in the fibreglass filters were frozen before being freeze-dried, then ground. They were stored in the dark at room temperature before analysis. Atenolol was analysed separately from ibuprofen, diclofenac and fluoxetine. Details of the methodology are presented in the studies of Gabet-Giraud et al. (2010) for atenolol and of Soulier (2012) for ibuprofen, diclofenac and fluoxetine. The general analytical scheme was similar in both cases. For dissolved concentration, the filtrates were acidified and extracted by solid phase extraction (SPE), evaporated to dryness, and reconstituted in a solution of internal standards. For particulate concentration, samples were extracted by pressurized solvent extraction for atenolol and microwaveassisted extraction for ibuprofen, diclofenac and fluoxetine, purified by SPE, and reconstituted in a solution containing internal standards. Analysis was performed by liquid chromatography coupled with tandem mass spectrometry (LCMS/MS). These methods reached very low limits of quantification (LQ): between 0.5 and 2 ng.L−1 in the dissolved phase and between 1 and 2 ng.gSS−1 (dry weight) in the particulate phase (Table 2). Suspended solids (SS) and macropollutant concentrations, like the biological chemical oxygen demand (bCOD), ammonium (NH4-N), nitrites (NO2-N), and nitrates (NO3-N), were analysed in each sample according to standard methods of the American Public Health Association (APHA 2012) to verify the presence or absence of bCOD or nitrogen forms during the experiment.
Calculation Determination of partition coefficient and biodegradation constant Partition coefficient (Kd) was determined using Eq. 1 with the dissolved and particulate concentrations of micropollutants measured 30 min after spiking, i.e. at t = 18.5 h. Biodegradation constants were obtained with the slopes of dissolved and particulate concentrations over time (Fig. 1), as measured during the different phases: – –
–
During phase I, kS,endo,Ox, kS,endo,Ax, kX,endo,Ox and kX,endo,Ax were determined. During phase II, while the elimination of both COD and nitrogen (NH4-N for reactor A, and NO3-N for reactor B) took place, kS,C-N,Ox, kS,C-N,Ax, kX,C-N,Ox and kX,C-N,Ax were determined. During phase III, while only N substrate was consumed, kS,N,Ox, kS,N,Ax, kX,N,Ox and kX,N,Ax were determined.
To ensure the reliability of the biodegradation constant values, we considered the uncertainty related to both sampling and analysis steps. A variation of concentration during a given phase below 10 % was considered not significant; consequently, the biodegradation constant was not calculated. In this study, the duration of sampling periods during each phase (5 h for phases I and III, 4 h for phase II) was a practical compromise to study different micropollutants in the same experiment. So, we were able to determine a reliable biodegradation constant value only above a determined threshold. For example, if Smp was 1,000 ng.L−1 at the beginning of
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phase II, the lowest value of kS,C-N that could be determined was 0.09 L.gSS−1.day−1. This protocol could obviously be adapted to increase the sensitivity of the method for slowly biodegradable micropollutants. Based on the previous example, with a 10-h-long phase II, the lowest calculable value for kS,C-N would reach 0.03 L.gSS−1.day−1. Otherwise, sampling period might be shortened for highly biodegradable micropollutants (to less than 2 h for example), to ensure the presence of micropollutant (>LQ) for the last sampling. Removal efficiency measured on the WWTP We calculated the removal efficiency of the wastewater line of the WWTP (RW, Eq. 3) and the total removal efficiency including sludge line (RT, Eq. 4) thanks to measurements of raw wastewater, treated wastewater and sludge concentrations. Corresponding flow rates were also measured to enable fluxes calculation. RW ¼
Fmp;INF −Fmp;EFF Fmp;INF
ð3Þ
RT ¼
Fmp;INF − Fmp;EFF − Fmp;S Fmp;INF
ð4Þ
RW RT Fmp,INF Fmp,EFF Fmp,S
Removal efficiency in the wastewater line [%] Total removal efficiency including sludge line [%] Micropollutant flux in raw wastewater [ng.day−1] Micropollutant flux in treated wastewater [ng.day−1] Micropollutant flux in the extracted sludge [ng.day−1]
Ibuprofen Firstly, ibuprofen was mainly present in the dissolved phase throughout the experiments (Smp >90 % of total concentration), whereas log Kow of ibuprofen (3.97, cf. Table 2) would suggest a tendency to sorb on particulate matter. This is confirmed by other studies (Urase and Kikuta 2005; Fernandez-Fontaina et al. 2012), and this can be explained by the ionisation of ibuprofen according to the pH with a resulting log Dow equal to 0.83. Secondly, residual Smp and Xmp at the end of the experiment in reactor A were significantly lower than in reactor B (232 and 16 ng.L−1, respectively, in reactor A and 9,245 and 781 ng.L−1, respectively, in reactor B). So, ibuprofen was better removed under aerobic conditions (reactor A) than under anoxic conditions (reactor B). Thirdly, in reactor A, Xmp decreased during phase III due to biodegradation (from 369 to 16 ng.L−1). The decrease in Smp was much higher, from 2,907 to 232 ng.L−1. So, we conclude that biodegradation of ibuprofen took place almost exclusively in the dissolved phase. Reactor A showed complete removal of ibuprofen by biodegradation (99 %). This result is in accordance with the removal efficiency observed at a full-scale bioreactor (cf. Table 2). In contrast, reactor B showed only 46 % removal of ibuprofen. Thus, ibuprofen was very efficiently removed under aerobic conditions but much less under anoxic conditions. Suarez et al. (2010) also reported a great difference in removal efficiency of ibuprofen between aerobic and anoxic conditions: in 2-L bioreactors continuously fed with synthetic substrate spiked in micropollutant, ibuprofen removal was above 80 % under aerobic conditions, whereas it was lower than 20 % under anoxic conditions. Only ibuprofen in dissolved phase was analysed by these authors. Atenolol
With RW, we quantified the removal efficiency from wastewater, that is to say the removal by biodegradation and by sorption. RT was the mass balance calculation of the total WWTP, including the flux in withdrawn sludge. Hence, in case of non-volatile micropollutant, RT allowed to determine the removal efficiency by biodegradation only.
Results and discussion Fate of micropollutants Figure 2 presents the time-course evolution of micropollutant concentrations in dissolved and particulate phases of reactors A and B. The fate of the four pharmaceutical compounds is detailed below.
Throughout the experiments, atenolol Xmp remained low (
RESEARCH ARTICLE
Lab-scale experimental strategy for determining micropollutant partition coefficient and biodegradation constants in activated sludge M. Pomiès & J. M. Choubert & C. Wisniewski & C. Miège & H. Budzinski & M. Coquery
Received: 15 May 2014 / Accepted: 22 September 2014 / Published online: 11 October 2014 # Springer-Verlag Berlin Heidelberg 2014
Abstract The nitrifying/denitrifying activated sludge process removes several micropollutants from wastewater by sorption onto sludge and/or biodegradation. The objective of this paper is to propose and evaluate a lab-scale experimental strategy for the determination of partition coefficient and biodegradation constant for micropollutant with an objective of modelling their removal. Four pharmaceutical compounds (ibuprofen, atenolol, diclofenac and fluoxetine) covering a wide hydrophobicity range (log Kow from 0.16 to 4.51) were chosen. Dissolved and particulate concentrations were monitored for 4 days, inside two reactors working under aerobic and anoxic conditions, and under different substrate feed conditions (biodegradable carbon and nitrogen). We determined the mechanisms responsible for the removal of the target compounds: (i) ibuprofen was biodegraded, mainly under aerobic conditions by cometabolism with biodegradable carbon, whereas anoxic conditions suppressed biodegradation; (ii) atenolol was biodegraded under both aerobic and anoxic conditions (with a higher biodegradation rate under aerobic conditions), and cometabolism with biodegradable carbon was the main mechanism; (iii) diclofenac and fluoxetine were removed by sorption only. Finally, the abilities of our strategy were evaluated by testing the suitability of the parameters for simulating Responsible editor: Angeles Blanco M. Pomiès : J. M. Choubert (*) : C. Miège : M. Coquery Irstea, UR MALY, 5 rue de la Doua, CS70077, 69626, Villeurbanne Cedex, France e-mail: [email protected] C. Wisniewski UMR Qualisud, Université Montpellier 1, 15 Av. Ch. Flahault BP 14491, 34093, Montpellier Cedex 5, France H. Budzinski EPOC/LPTC UMR5805, Université Bordeaux 1, 351 Cours de la libération, 33405, Talence Cedex, France
effluent concentrations and removal efficiency at a full-scale plant. Keywords Modelling . Pharmaceuticals . Experimental strategy . Wastewater treatment . Activated sludge
Introduction Wastewater treatment plant (WWTP) effluents are a wellknown vector of micropollutant discharge, particularly for pharmaceutical compounds, into the environment (Heberer 2002; Ternes and Joss 2006). Ongoing regulation in Europe regularly reinforces the requirements towards discharge of chemicals into the environment, with lists of priority substances regularly revisited (EC 2008, 2013). A broad strand of research has highlighted how WWTPs could efficiently remove many micropollutants from wastewater (e.g., Miège et al. 2009; Onesios et al. 2009; Luo et al. 2014), particularly by biological processes, even though they were not originally designed for this purpose. However, discharged WWTP effluents still contain several micropollutants due to their high initial concentrations in raw wastewater and/or their low affinity with suspended solids (SS) and their non-biodegradable chemical structure (Martin Ruel et al. 2012). The identification of micropollutant removal mechanisms, as well as their modelling to predict and eventually better control their overall elimination, is essential. Sorption onto sludge and biodegradation are the two main mechanisms dictating the fate of non-volatile micropollutants in biological wastewater treatment systems (Carballa et al. 2004). Micropollutant can be biodegraded, either directly or indirectly, by cometabolism involving macropollutant as a growth substrate (basically biodegradable chemical oxygen demand (bCOD)) (Criddle 1993). Both mechanisms have been modelled using parameters usually based on lab-scale
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experiments (Pomiès et al. 2013). However, experimental conditions can vary greatly from one author to another and no unified experimental protocol is proposed in the literature, leading to possible biased modelling results (Clouzot et al. 2013). The discrepancy in operating conditions of experiments and in the sludge tested makes relatively difficult the determination of typical values for parameters for each micropollutant. Nevertheless, such set of values may be necessary for engineering application of micropollutant modelling (i.e. typical parameter values). This study aimed at giving new insight for determining partition coefficient and biodegradation constant values for modelling micropollutant removal in activated sludge process. We implemented a robust experimental strategy in lab-scale experiments involving micropollutant analysis in both dissolved and particulate phases of biological sludge. The protocol integrated different conditions of nitrifying/denitrifying activated sludge process, i.e. oxido-reduction conditions and presence or absence of biodegradable macropollutants (biodegradable carbon, nitrogen). Our experiments aimed to (i) determine the predominant removal mechanism (sorption or biodegradation) and the most favourable conditions of removal for a selection of micropollutants; (ii) determine the compartment (dissolved or particulate) involved in biodegradation; and (iii) quantify the values of two modelling parameters, the partition coefficient (Kd) and the biodegradation constant (k). Then, we tested the measured parameter values to simulate micropollutant removal in a full-scale activated sludge process. Our work targeted four pharmaceutical compounds (ibuprofen, atenolol, diclofenac, fluoxetine) that share different physicochemical properties and different fates through WWTPs.
Partition coefficient and biodegradation constant: experimental determination Current existing protocols
Two compartments are usually defined for sorption onto sludge: the dissolved compartment (Smp) and the particulate one (Xmp). For each micropollutant, a partition coefficient (Kd) is defined to account for the main sorption mechanisms, i.e. hydrophobic mechanisms and electrostatic interactions (Sipma et al. 2010). It is calculated as follows (Eq. 1):
Kd
Smp SS
=SS
Xmp
Smp
ð1Þ
Partition coefficient of micropollutant onto sludge [L.gSS−1]
Concentration of micropollutant in particulate phase [ng.L−1] Concentration of micropollutant in dissolved phase [ng.L−1] Suspended solids concentration of sludge in biological reactor [gSS.L−1]
Lab-scale experiments are generally used to determine Kd. A common procedure consists in spiking sludge with micropollutant at defined concentrations and then measuring the dissolved and/or particulate micropollutant concentration after a defined short time considering sorption equilibrium is reached rapidly (Ternes et al. 2004; Urase and Kikuta 2005; Lindblom et al. 2009; Abegglen et al. 2009; FernandezFontaina et al. 2014). However, although widely accepted, these procedures carry weaknesses (Stevens-Garmon et al. 2011; Pomiès et al. 2013). First, the mass balance is rarely verified because particulate concentration is often not measured, but simply deduced from dissolved concentration (Hyland et al. 2012). Second, in sorption studies, micropollutant biodegradation is generally inhibited by diverse procedures (Maurer et al. 2007; Wick et al. 2009). However, the biodegradation inhibition may induce potential bias in Kd measurement, as it may influence sorption performances by altering the physicochemical properties of sludge (Stevens-Garmon et al. 2011). Biodegradation The biodegradation rate of a micropollutant is often described with a pseudo-first-order kinetic (Joss et al. 2006; Wick et al. 2009). It is calculated as follows (Eq. 2): dSmp ¼ −k Smp SS ð2Þ dt biodegradation k
Sorption
Kd ¼
Xmp
Biodegradation constant [L.gSS−1.day−1]
The biodegradation process also suffers from a lack of standardization. Biodegradation of micropollutant is usually evaluated from the time-course evolution of micropollutant concentrations after spiking (Lindblom 2009; Plosz et al. 2010). According to the authors, spiking can consist to add real raw wastewater (Plosz et al. 2010) or micropollutants mixed with synthetic substrate (Suarez et al. 2010). The biodegradation constant (k) is mostly determined for micropollutant measured in the dissolved phase only (Wick et al. 2009), while micropollutant in the particulate phase is seldom measured (Urase and Kikuta 2005). Only few studies investigated the effect of oxido-reduction potential (ORP) conditions on biodegradation efficiencies: a significant
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decrease has been shown under anoxic conditions for ibuprofen, estrone and estriol (Joss et al. 2004; Plosz et al. 2010; Suarez et al. 2010), but the difference still needs to be determined for other pharmaceutical compounds.
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Materials and methods Experimental strategy WWTP sampling
Proposed experimental protocol In the proposed experimental strategy, we studied sorption and biodegradation in the same reactor, considering that sorption occurs rapidly and biodegradation takes place at a longer time scale (cf. “Current existing protocols”). As a consequence, we did not need add any biodegradation inhibitor which could alter sludge characteristics. We based our protocol on the analysis of dissolved and particulate phases. Indeed, for a micropollutant that may be present in both phases (i.e. Smp and Xmp, respectively), this is a key issue for understanding its fate. Calculating Kd with Smp and Xmp is more reliable than using the Smp decrease only, as it enables to verify mass conservation. Moreover, it allows pinpointing where the biodegradation reaction occurs. A current assumption in the literature is to consider that only the micropollutant in the dissolved phase is biodegradable (Lindblom et al. 2009; Plosz et al. 2010). Experimental data from our protocol enabled to confirm or invalidate this assumption, particularly for micropollutants known to have some affinity with organic matter and thus can be retained on it. Our experimental strategy, as detailed in the section “Experimental strategy”, also aimed to define the influence of operating conditions, i.e. the nature of the available biodegradable substrate and the ORP conditions. As shown on Table 1, we defined biodegradation rates separately for dissolved (with an S subscript) and particulate phases (with an X subscript) under aerobic (with an Ox subscript) and anoxic (with an Ax subscript) conditions. For each micropollutant, we considered different biodegradation constants (k) according to the absence or the presence of macropollutant as growth substrate (biodegradable carbon, nitrogen as ammonium under aerobic conditions and nitrates under anoxic conditions). This formalism dissociates cometabolism with biodegradable carbon and nitrogen (action of heterotrophic biomass), cometabolism with nitrogen only (action of autotrophic biomass), and the endogenous process. Biodegradation constants used in the biodegradation equations are easier to measure than the parameters of Criddle’s formalism (Criddle 1993; Delgadillo-Mirquez et al. 2011; Fernandez-Fontaina et al. 2014). Pseudo-first-order kinetics were used (Table 1), like in Eq. 2, with the frequently used Monod kinetics functions (switch functions taking values between 0 and 1) to modulate the degradation rate according to the oxygen concentration, the biodegradable COD concentration, and the ammonium concentration.
Sludge and wastewater samples came from a domestic WWTP (2,900 population equivalents (P.E.) based on 60 gBOD5 P.E.−1 as suggested by EC (1991)) consisting in a nitrifying/denitrifying activated sludge process and a secondary clarifier. Samples were collected in June 2011, where the activated sludge process was operated at 50-day solid retention time (hydraulic retention time (HRT) of 4 days), with around 7.3 gSS.L−1 and a temperature of around 23 °C. Raw and treated wastewater samples were collected to constitute 24-h flow proportional composite samples using a refrigerated automatic sampler with cleaned Teflon tubing and glass containers (Choubert et al. 2011). Micropollutants in dissolved and particulate phases were analysed in raw wastewater, whereas only the dissolved phase was analysed in treated wastewater due to the low concentration of suspended solids (7 mgO2.L−1), whereas reactor B was maintained under anoxic conditions. The experiment was carried out for 4 days. The reactors were 200-L plastic tanks, continuously stirred. A subsample was regularly collected in each reactor via a valve port fitted at the bottom of the tanks. In each reactor, we collected a total of nine samples for the analysis of the micropollutant dissolved (Smp) and particulate (Xmp) concentrations, biodegradable carbon (i.e. biodegradable chemical oxygen demand (bCOD) concentration), and nitrogen (i.e. ammonium (NH4-N) and nitrate (NO3-N) concentrations). The experimental protocol was defined according to the constraints relative to on-site production of filtered samples, analytical capability, and expected timecourse evolution of the concentrations versus the sampling and analytical uncertainty. In each reactor, four periods (phases) were applied: a preliminary phase and three other phases corresponding to
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Table 1 Equations involved in biodegradation proposed in this work Conditions
Smp
Xmp
Aerobic
Presence of biodegradable carbon and ammonium Presence of ammonium only
−1 −1
Absence of biodegradable substrate
−1
Presence of biodegradable carbon and nitrates Presence of nitrates only
−1
Absence of biodegradable substrate
−1
Presence of biodegradable carbon and ammonium Presence of ammonium only
−1
Absence of biodegradable substrate
−1
Presence of biodegradable carbon and nitrates Presence of nitrates only
−1
Absence of biodegradable substrate
−1
Anoxic
Aerobic
Anoxic
−1
−1
−1
Reaction rates
B k S;C–N;Ox Smp SS SO þKSOO2;OHO SBSþK 1 SNHx 1 SBKþK k S;N;Ox Smp SS SO þKSOO2;OHO SNHx þK NHx;ANO 1 K2 1 SNHx;ANO k S;endo;Ox Smp SS SO þKSOO2;OHO SBKþK þK2 1 KO2;OHO SNOx B SBSþK SO þK k S;C–N;Ax Smp SS SNOx þK NOx;OHO 1 O2;OHO KO2;OHO SNOx SNHx 1 k S;N;Ax Smp SS SNOx þK SNHx þK SO þK SBKþK NOx;OHO NHx;ANO O2;OHO 1 KO2;OHO SNOx K1 K2 k S;endo;Ax Smp SS SNOx þK KO2;OHO þSO K1 þSB K2 þSNHx NOx;OHO SO SB k X;C−N;Ox Xmp SS SO þKO2;OHO SB þK1 SNHx K1 k X;N;Ox Xmp SS SO þKSOO2;OHO SNHx þK SB þK1 NHx;ANO SO K1 K2 k X;endo;Ox Xmp SS SO þKO2;OHO SB þK1 SNHx þK2 KO2;OHO SNOx SB k X;C−N;Ax Xmp SS SNOx þK S þK S þK NOx;OHO B 1 O O2;OHO KO2;OHO SNOx SNHx K1 k X;N;Ax Xmp SS SNOx þKNOx;OHO SNHx þKNHx;ANO SO þK S þK O2;OHO B 1 KO2;OHO SNOx K1 K2 k X;endo;Ax Xmp SS SNOx þK KO2;OHO þSO K1 þSB K2 þSNHx NOx;OHO
So dissolved oxygen concentration, KO2,OHO oxygen half-saturation coefficient for heterotrophic biomass, SNHx dissolved ammonia nitrogen concentration, KNHx,ANO ammonia half-saturation coefficient for autotrophic biomass, SNOx dissolved nitrate nitrogen concentration, KNOx,OHO nitrate halfsaturation coefficient for heterotrophic biomass, SB dissolved biodegradable COD, K1, K2 half-saturation coefficients for heterotrophic biomass, kS,C-N,Ox, kX,C-N,Ox biodegradation constants for Smp and Xmp in presence of biodegradable COD and nitrogen under aerobic conditions, kS,C-N,Ax, kX,C-N,Ax biodegradation constants for Smp and Xmp in presence of biodegradable COD and nitrogen under anoxic conditions, kS,N,Ox, kX,N,Ox biodegradation constants for Smp and Xmp in presence of nitrogen only, under aerobic conditions, kS,N,Ax, kX,N,Ax biodegradation constants for Smp and Xmp in presence of nitrogen only, under anoxic conditions, kS,endo,Ox, kX,endo,Ox biodegradation constants for Smp and Xmp in absence of biodegradable substrate and aerobic conditions, kS,endo,Ax, kX,endo,Ax biodegradation constants for Smp and Xmp in absence of biodegradable substrate and anoxic conditions
different macropollutant supply conditions (phase I, no macropollutant addition; phase II, biodegradable carbon and nitrogen addition; phase III, nitrogen addition). The different phases of the experiment are presented in the following sections and illustrated in Fig. 1 (using the example of reactor A). The pH was practically stable during the experiments, with 8.3±0.2 (reactor A) and 8.1±0.1 (reactor B), due to buffer effect or addition of calcium carbonates. Preliminary phase (t=0 to 18 h) Reactors A and B were filled with 180 L of sludge collected in the WWTP aeration tank (t=0 h). Intermittent aeration was applied for 18 h to attempt to eliminate possible nonbiodegraded residual substrate contained in the liquid sludge. The initial micropollutant concentration in the sludge (Xmp) was measured at t=0 h. Phase I “without any substrate” (t=18 to 43 h) At t= 18 h, each reactor was spiked with a mixture of micropollutants (10 μg.L−1 each) dissolved in 10 mL of
methanol solvent. These levels are in the usual high concentration range of raw wastewater (Miège et al. 2009). ORP conditions were monitored in both reactors. Reactor A was supplied with air via a diffuser located at the bottom of the reactor. In reactor B, N2 gas was supplied to initiate anoxic conditions. No chemical agent (e.g. inhibitor) was used. In each reactor, sludge was sampled three times at t=18.5 h (for dissolved and particulate analysis), at t=23 h (for dissolved analysis, reactor B only) and at t=42 h (for dissolved analysis). Phase I aimed at determining partition coefficient and biodegradation of micropollutants without biodegradable carbon and nitrogen conditions (endogenous conditions). Phase II “with biodegradable carbon and nitrogen substrates” (t=43 to 66 h) At t=43 h, a volume of 50 L of raw wastewater was added to each reactor (raw wastewater to sludge ratio 1:4 v/v). Raw wastewater was a source of biodegradable substrate (bCOD= 175 mg.L−1, NH4-N=72 mg.L−1). In reactor B, we also added NaNO3 (NO3-N=18 mg.L−1). Suspended solids concentration in both reactors was 5.4 gSS.L−1 after raw wastewater
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Sampling for Smp analysis Sampling for Xmp analysis
Xmp and Smp [ng.L-1]
Preliminary phase
Phase I
Phase II
kS,endo,Ox
Phase III
kS,C-N,Ox kS,N,Ox
Kd
kX,endo,Ox
0
18 18.5
Sludge addition (180 L)
Micropollutant spiking (10 µg.L -1)
23
kX,C-N,Ox
42 43 43.5
47.5
Raw wastewater addition (50 L)
kX,N,Ox
66 66.5
70
71.5
Time (h)
NH4-N addition (40 mgN.L-1)
Fig. 1 Successive phases applied to reactor A with sampling type (for analysis of dissolved Smp and particulate Xmp micropollutant concentrations) and the sorption and biodegradation parameters
addition. In each reactor, sludge was sampled at t=43.5 h (for dissolved and particulate analysis) and t=47.5 h (for dissolved analysis). Phase II aimed at determining biodegradation of micropollutants under biodegradable carbon and nitrogen conditions. Phase III “with nitrogen substrate only” (t=66 to 72 h) At t=66 h, we increased the NH4-N concentration in reactor A to 40 mgN.L−1 by adding ammonium chloride (NH4Cl) to ensure non-limiting nitrogen substrate conditions and to stimulate nitrification. In reactor B, we increased the NO3-N concentration to 30 mgN.L−1 by adding sodium nitrate (NaNO3) to stimulate denitrification. In each reactor, sludge was sampled at t=66.5 h (for dissolved analysis), 70 h (for dissolved analysis) and 71.5 h (for dissolved and particulate analysis). Phase III aimed at determining biodegradation of micropollutants under nitrogen-only conditions. Studied micropollutants and chemical analysis Four pharmaceutical compounds (ibuprofen, atenolol, diclofenac and fluoxetine) were chosen based on their differing physicochemical properties as shown in Table 2. Sorption onto sludge may differ due to varying log Kow values (0.16 for
atenolol to 4.51 for diclofenac) and pKa values (from 4.15 for diclofenac to 10.1 for fluoxetine) (Suarez et al. 2008). Also, biodegradation potential may be influenced by the molecular feature of the micropollutants, as some of the functional groups are known to favour their biodegradability (e.g. carboxylic acid, amide or hydroxyl groups and phenyl ring), whereas other groups increase their resistance to biodegradability (e.g. extensive alkyl chain branching or halogens) (Boethling et al. 2007). Ibuprofen is usually removed at more than 90 % by biodegradation (Joss et al. 2005; Soulier et al. 2011). Atenolol is also removed by biodegradation, but at a lower level (50–70 %) (Jelic et al. 2011). Removal efficiency of diclofenac is discrepant according to the studies (from 5 to 80 %); Jelic et al. (2011) considered that biodegradation is the main removal mechanism but sorption can occur also. Fluoxetine is partly removed by sorption (33 and 50 %) (Radjenovic et al. 2009; Soulier et al. 2011) and/or biodegradation (Suarez et al. 2012). For the analysis of the dissolved phase of the sludge from lab-scale experiments, we sampled and filtered directly in situ with a filtration module (polysulfone, 0.1 μm, 0.16 m2, Polymem, France). Filtrates were analysed 24 h after sampling (or frozen before analysis). Concerning the dissolved phase of raw and treated wastewater samples from full-scale experiments, samples were filtered in laboratories less than 24 h after
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Table 2 Physicochemical properties (log Kow) and limits of quantification (LQ) of the studied pharmaceutical compounds and literature data on removal efficiency (RW) through activated sludge process Micropollutant (family)
Ibuprofen (antiinflammatory) Atenolol (beta-blocker) Diclofenac (antiinflammatory) Fluoxetine (antidepressant)
Formula
Molecular weight [g/mol]
pKa
*
Log Kow
Log Dow a
LQ in dissolved phase [ng.L-1]
LQ in particulate phase [ng.L1]
C13 H18 O2
Removal efficiency by activated sludge process found in the literature (RW , wastewater line) [%]
S uggested removal mechanism in literature S orption
Biodegradation
(3)
Molecular features influen cing biodegradability Positive effect (increase) Carboxylic acid group (COOH)
Negative effect (decrease ) Extensive alkyl chain branching
206
4.91
3.971,2
0.68
1
2
90 99 (4) 98 (5)
-
X (9)
266
9.67
0.161
-1.32
2.5
1.5
71 (6) 61 (4) 61 (5)
-
X (10)
Amide group and hydroxyl group (OH)
Extensive alkyl chain branching
296
4.15
4.511,2
0.46
0.5
1
69 (7) 5-80 (8) 22 (4) 9 (5)
X (11)
X (12)
Carboxylic acid group (COOH)
2 chlorine atoms
309
9.8
4.052
2.44
1
1
33 (4) 50 (5)
X (13)
X (13)
Phenyl ring
3 fluorine atoms
C14H 22N2O3
C14H 11Cl2NO 2
C17H 18F3NO
14
Verlicchi et al. (2012); 2 Hyland et al. (2012); Ternes et al. (2004); 4 Radjenovic et al. (2009); 5 Soulier et al. (2011); Carucci et al. (2006); 7 Ternes (1998); Paxeus (2004); Salgado et al. (2012); 10 Maurer et al. (2007); 11 Suarez et al. (2012); Suarez et al. 2008; Fernandez-Fontaina et al. (2012); 14 Boethling et al. (2007) a
Calculated for acids with equation log Dow=log Kow−log(1+10(pH−pKa) ) and for basis with log Dow=log Kow−log(1+10(pKa–pH) ) using the mean pH of activated sludge in reactors (=8.2), pKa and log Kow of each micropollutant
b
From http://chem.sis.nlm.nih.gov/chemidplus/ (for ibuprofen, diclofenac), or from http://www.drugbank.ca (for atenolol and fluoxetine)
sampling on fibreglass filters (GF/F, 0.7 μm). For particulate phase from full- and lab-scale experiments, sludge samples were centrifuged in laboratories within 24 h after sampling; centrifuged sludge and suspended solids collected in the fibreglass filters were frozen before being freeze-dried, then ground. They were stored in the dark at room temperature before analysis. Atenolol was analysed separately from ibuprofen, diclofenac and fluoxetine. Details of the methodology are presented in the studies of Gabet-Giraud et al. (2010) for atenolol and of Soulier (2012) for ibuprofen, diclofenac and fluoxetine. The general analytical scheme was similar in both cases. For dissolved concentration, the filtrates were acidified and extracted by solid phase extraction (SPE), evaporated to dryness, and reconstituted in a solution of internal standards. For particulate concentration, samples were extracted by pressurized solvent extraction for atenolol and microwaveassisted extraction for ibuprofen, diclofenac and fluoxetine, purified by SPE, and reconstituted in a solution containing internal standards. Analysis was performed by liquid chromatography coupled with tandem mass spectrometry (LCMS/MS). These methods reached very low limits of quantification (LQ): between 0.5 and 2 ng.L−1 in the dissolved phase and between 1 and 2 ng.gSS−1 (dry weight) in the particulate phase (Table 2). Suspended solids (SS) and macropollutant concentrations, like the biological chemical oxygen demand (bCOD), ammonium (NH4-N), nitrites (NO2-N), and nitrates (NO3-N), were analysed in each sample according to standard methods of the American Public Health Association (APHA 2012) to verify the presence or absence of bCOD or nitrogen forms during the experiment.
Calculation Determination of partition coefficient and biodegradation constant Partition coefficient (Kd) was determined using Eq. 1 with the dissolved and particulate concentrations of micropollutants measured 30 min after spiking, i.e. at t = 18.5 h. Biodegradation constants were obtained with the slopes of dissolved and particulate concentrations over time (Fig. 1), as measured during the different phases: – –
–
During phase I, kS,endo,Ox, kS,endo,Ax, kX,endo,Ox and kX,endo,Ax were determined. During phase II, while the elimination of both COD and nitrogen (NH4-N for reactor A, and NO3-N for reactor B) took place, kS,C-N,Ox, kS,C-N,Ax, kX,C-N,Ox and kX,C-N,Ax were determined. During phase III, while only N substrate was consumed, kS,N,Ox, kS,N,Ax, kX,N,Ox and kX,N,Ax were determined.
To ensure the reliability of the biodegradation constant values, we considered the uncertainty related to both sampling and analysis steps. A variation of concentration during a given phase below 10 % was considered not significant; consequently, the biodegradation constant was not calculated. In this study, the duration of sampling periods during each phase (5 h for phases I and III, 4 h for phase II) was a practical compromise to study different micropollutants in the same experiment. So, we were able to determine a reliable biodegradation constant value only above a determined threshold. For example, if Smp was 1,000 ng.L−1 at the beginning of
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phase II, the lowest value of kS,C-N that could be determined was 0.09 L.gSS−1.day−1. This protocol could obviously be adapted to increase the sensitivity of the method for slowly biodegradable micropollutants. Based on the previous example, with a 10-h-long phase II, the lowest calculable value for kS,C-N would reach 0.03 L.gSS−1.day−1. Otherwise, sampling period might be shortened for highly biodegradable micropollutants (to less than 2 h for example), to ensure the presence of micropollutant (>LQ) for the last sampling. Removal efficiency measured on the WWTP We calculated the removal efficiency of the wastewater line of the WWTP (RW, Eq. 3) and the total removal efficiency including sludge line (RT, Eq. 4) thanks to measurements of raw wastewater, treated wastewater and sludge concentrations. Corresponding flow rates were also measured to enable fluxes calculation. RW ¼
Fmp;INF −Fmp;EFF Fmp;INF
ð3Þ
RT ¼
Fmp;INF − Fmp;EFF − Fmp;S Fmp;INF
ð4Þ
RW RT Fmp,INF Fmp,EFF Fmp,S
Removal efficiency in the wastewater line [%] Total removal efficiency including sludge line [%] Micropollutant flux in raw wastewater [ng.day−1] Micropollutant flux in treated wastewater [ng.day−1] Micropollutant flux in the extracted sludge [ng.day−1]
Ibuprofen Firstly, ibuprofen was mainly present in the dissolved phase throughout the experiments (Smp >90 % of total concentration), whereas log Kow of ibuprofen (3.97, cf. Table 2) would suggest a tendency to sorb on particulate matter. This is confirmed by other studies (Urase and Kikuta 2005; Fernandez-Fontaina et al. 2012), and this can be explained by the ionisation of ibuprofen according to the pH with a resulting log Dow equal to 0.83. Secondly, residual Smp and Xmp at the end of the experiment in reactor A were significantly lower than in reactor B (232 and 16 ng.L−1, respectively, in reactor A and 9,245 and 781 ng.L−1, respectively, in reactor B). So, ibuprofen was better removed under aerobic conditions (reactor A) than under anoxic conditions (reactor B). Thirdly, in reactor A, Xmp decreased during phase III due to biodegradation (from 369 to 16 ng.L−1). The decrease in Smp was much higher, from 2,907 to 232 ng.L−1. So, we conclude that biodegradation of ibuprofen took place almost exclusively in the dissolved phase. Reactor A showed complete removal of ibuprofen by biodegradation (99 %). This result is in accordance with the removal efficiency observed at a full-scale bioreactor (cf. Table 2). In contrast, reactor B showed only 46 % removal of ibuprofen. Thus, ibuprofen was very efficiently removed under aerobic conditions but much less under anoxic conditions. Suarez et al. (2010) also reported a great difference in removal efficiency of ibuprofen between aerobic and anoxic conditions: in 2-L bioreactors continuously fed with synthetic substrate spiked in micropollutant, ibuprofen removal was above 80 % under aerobic conditions, whereas it was lower than 20 % under anoxic conditions. Only ibuprofen in dissolved phase was analysed by these authors. Atenolol
With RW, we quantified the removal efficiency from wastewater, that is to say the removal by biodegradation and by sorption. RT was the mass balance calculation of the total WWTP, including the flux in withdrawn sludge. Hence, in case of non-volatile micropollutant, RT allowed to determine the removal efficiency by biodegradation only.
Results and discussion Fate of micropollutants Figure 2 presents the time-course evolution of micropollutant concentrations in dissolved and particulate phases of reactors A and B. The fate of the four pharmaceutical compounds is detailed below.
Throughout the experiments, atenolol Xmp remained low (
