Bioresource Technology 121 (2012) 432–440

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Bioresource Technology

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Review

Review on the fate of emerging contaminants during sludge anaerobic digestion

Athanasios S. Stasinakis ⇑Department of Environment, Water and Air Quality Laboratory, University of the Aegean, University Hill, Mytilene 81 100, GreeceEawag, Swiss Federal Institute of Aquatic Science and Technology, Dübendorf 8600, Switzerland

h i g h l i g h t s

" Several emerging contaminants (ECs) have been detected in sludge." For most of them, available data for their fate and effects on AD is yet limited." Factors enhancing biotransformation are not the same for all ECs." A deeper understanding of the role of sorption on ECs removal is required." Combination of AD with other sludge treatment processes could enhance ECs removal.

a r t i c l e i n f o

Article history:Received 19 May 2012Received in revised form 18 June 2012Accepted 21 June 2012Available online 30 June 2012

Keywords:MicropollutantsOccurrenceAnaerobic digestionFateMesophilic

0960-8524/$ - see front matter � 2012 Elsevier Ltd. Ahttp://dx.doi.org/10.1016/j.biortech.2012.06.074

⇑ Tel.: +30 22510 36257; fax: +30 22510 36246.E-mail address: astas@env.aegean.gr

a b s t r a c t

Several research papers have been published during the last years investigating the occurrence, fate andeffects of emerging contaminants (ECs) on sludge anaerobic digestion (AD). Literature review revealedthat research has been mainly focused on specific groups of compounds (linear alkylbenzene sulpho-nates, nonylphenol ethoxylates, some pharmaceuticals, estrogens, phthalates), while there are fewer orno data for others (personal care products, perfluorinated compounds, brominated flame retardants,organotins, benzotriazoles, benzothiazoles, nanoparticles). AD operational parameters (sludge residencetime, temperature), sludge characteristics (type of sludge, adaptation on the compound), physicochemi-cal properties of ECs and co-metabolic phenomena seem to affect compounds’ biodegradation. The use ofsludge pretreatment methods does not seem to enhance ECs removal; whereas encouraging results havebeen reported when AD was combined with other treatment methods. Future efforts should be focusedon better understanding of biotransformation processes and sorption phenomena occurred in anaerobicdigesters, as well as on identification of (bio)transformation products.

� 2012 Elsevier Ltd. All rights reserved.

1. Introduction

A great number of emerging contaminants (ECs) have beendetected in municipal wastewater. Despite the fact that the defini-tion of ‘‘emerging’’ is not clear, as according to Field et al. (2006)‘‘what is emerging is a matter of perspective as well as timing’’,usually organic micropollutants such as surfactants, personal careproducts, pharmaceuticals, estrogens, perfluorinated compounds,phthalate acid esters as well as organic and inorganic nanoparticlesare included in this category (Diaz-Cruz et al., 2009; Clarke andSmith, 2011). During wastewater treatment most of these com-pounds are sorbed to some extent on suspended solids and as aresult they are found in sludge through sedimentation occurringin primary and secondary clarifiers.

ll rights reserved.

Anaerobic digestion (AD) is one of the most widely used pro-cesses for sludge stabilization, while treated sludge is very oftendisposed to the soil or reused for agricultural purposes. In a recentpaper, Kelessidis and Stasinakis (2012) reported that 53% of sludgein EU-27 is used in agriculture directly or after composting, whileCitulski and Farahbakhsh (2010) reported that more than 40% ofproduced biosolids are applied to land in USA and Canada.

The presence of ECs in sewage sludge and the available analyticalmethods for their quantitative determination have been reviewedby Diaz-Cruz et al. (2009). Moreover in a recent review, Clarke andSmith (2011) reported the possible threats for the environmentand human health due to the agricultural use of biosolids containingECs. It is evident that the occurrence of ECs in digested sludge isdependent on their fate during sludge treatment.

Based on the above, this review aims to give an overview of thedata concerning fate of ECs during sludge AD. The concentrationlevels of these compounds in sludge as well as their possible effectson AD are also presented. The study has been focused on 11 differ-ent categories of ECs, namely; surfactants, personal care products,

Table 1Representative compounds from different groups of emerging contaminants.

Compound Chemical structure

Surfactants

A.S. Stasinakis / Bioresource Technology 121 (2012) 432–440 433

pharmaceuticals, estrogens, phthalate acid esters, perfluorinatedcompounds, organotins, brominated flame retardants, benzotriazoles,benzothiazoles and nanoparticles. Representative compounds ofthese groups are presented in Table 1.

Sodium dodecylbenzenesulfonate,C12LAS

Nonylphenol, NP

Dodecyl trimethyl ammonium chloride,C12TMA

Personal care productsTriclosan, TCS

Galaxolide, HHCB

PharmaceuticalsDiclofenac, DCF

Carbamazepine, CBZ

Sulfamethoxazole, SMX

EstrogensEstrone, E1

Phthalate acid estersBis(2-ethylhexyl) phthalate, DEHP

Perfluorinated compoundsPerfluorooctanoic acid, PFOA

OrganotinsBis(tributyltin)oxide, TBTO

Brominated flame retardantsTetrabromobisphenol A, TBBPA

BenzotriazolesBenzotriazole, BTr

2. Occurrence of emerging contaminants in sludge

The concentration levels of different groups of ECs as well as theconcentrations of some representative compounds are presentedin Fig. 1 and Table S1, respectively. Further data for the occurrenceof emerging chemicals in biosolids can be found in previous reviewpapers (Diaz-Cruz et al., 2009; Clarke and Smith, 2011).

Concentrations of ECs in sludge are significantly different be-tween different groups of compounds, ranging from some lg kg�1

to several g kg�1 dw (Fig. 1). In general, the highest concentrationshave been observed for some surfactants. However, it should bementioned that even for the same compounds, significant differ-ences are often observed between different countries or even be-tween different Sewage Treatment Plants (STPs) of the samecountry. Concentrations in sludge are often correlated with theconcentrations of target compounds in influent wastewater (Fent,1996; Stasinakis et al., 2008), but they are also affected by thephysicochemical properties of the compounds (molecular weight,hydrophobicity, water solubility, pKa, resistance to biodegradation)(Janex-Habibi et al., 2009; Clara et al., 2010), the sludge character-istics (pH, organic matter, cations’ concentration) and the opera-tional parameters of each STP (presence or absence of primarysedimentation, hydraulic residence time in different tanks, sludgeresidence time in bioreactors, method of sludge stabilization)(Esperanza et al., 2007; Stasinakis et al., 2008; Heidler and Halden,2009; Janex-Habibi et al., 2009). Additionally, there are cases inwhich the mass flows of ECs in sludge are higher comparing tothose in influent wastewater (e.g., for perfluorinated compounds),indicating their formation during biological treatment by precur-sors’ transformation (Schultz et al., 2006).

Regarding surfactants, significant differences in their concentra-tion levels have been observed between linear alkyl benzene sul-phonates (LAS), nonylphenol ethoxylates (NPE) and quaternaryammonium-based compounds (QAC), which are representativesgroups of anionic, nonionic and cationic surfactants, respectively.The highest concentrations in sludge have been observed for LAS(Fig. 1, Table S1) and they are mainly due to their wider use. 2.8million tons of LAS were used in 1998, comparing to 0.4 milliontons of NPE in 1997 and 0.5 million tons of QAC in 2004 (Birkettand Lester, 2003; Ismail et al., 2010). On the other hand, NPE arerapidly biotransformed during wastewater biological treatmentto some metabolites such as nonylphenol (NP), nonylphenol mono-ethoxylate (NP1EO) and nonylphenol diethoxylate (NP2EO)(Birkett and Lester, 2003). The concentrations of these compoundsin sludge seem to be differentiated significantly between differentstudies, ranging from few mg kg�1 to more than one g kg�1 dw(Table S1).

Regarding personal care products, the highest concentrationshave been reported for the widely used hydrophobic compounds;triclosan (TCS), triclocarban (TCC) and galaxolide (HHCB), whilelower concentrations of other synthetic musks have been observed(Fig. 1, Table S1). Concentrations of pharmaceuticals in sludge arevaried between different compounds, as well as between differentcountries, indicating differences on their physicochemical charac-teristics and usage rates. The presence of natural estrogens (estrone,E1; 17b-estradiol, E2 and estriol, E3) and 17a-ethinylestradiol (EE2)in sludge samples has been investigated in several studies(Andersen et al., 2003; Citulski and Farahbakhsh, 2010). In mostcases, concentrations of these compounds were in the range of somelg kg�1 dw (Table S1).

(continued on next page)

Table 1 (continued)

Compound Chemical structure

BenzothiazolesBenzothiazole, BT

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Among ECs that mainly originate from industrial applications,phthalic acid esters (and especially bis(2-ethylhexyl) phthalate,DEHP) are usually detected in sludge at higher concentrationscomparing to brominated flame retardants, organotins and per-fluorinated compounds (Fig. 1, Table S1). Finally so far, there arelimited or even no data for the occurrence of benzotriazoles (BTrs),benzothiazoles (BTs) and nanoparticles in sludge (Table S1). BTrsand BTs are high production volume chemicals and their existencehas been documented in wastewater at lg L�1 level. Beside the factthat they are polar compounds, some recent studies have showntheir presence in sludge samples (Wick et al., 2010; Liu et al.,2012). Regarding nanoparticles, despite the increasing trends thatare observed on their use; no data is yet available for their concen-tration levels in sludge.

3. Effects of emerging contaminants on sludge anaerobicdigestion

The effects of ECs on AD have been studied using simple batchexperiments, as well as in pilot-scale bioreactors operating in thepresence of these compounds for a long period of time. Parameterssuch as biogas and methane production, accumulation of volatile

Fig. 1. Reported ranges of emerging contami

fatty acids (VFAs) and volatile suspended solids (VSS) removal havebeen used as endpoints for evaluating micropollutants’ inhibition(Gruden et al., 2001; Garcia et al., 2006; Watson et al., 2012; Yanget al., 2012). Most of the published results are referred to some sur-factants and pharmaceuticals (mainly antibiotics used by humansand animals), whereas there are much fewer data for the othergroups of compounds.

Regarding the effects of surfactants on sludge AD, Angelidakiet al. (2004) reported that butyrate and propionate utilizing bacteriawere much more sensitive to LAS compared to methanogens. Addi-tion of LAS homologues to the anaerobic digesters increased biogasproduction at concentrations below 10 mg C L�1 (or 5 g of surfactantkg�1 dry sludge); while partial or total inhibition of methanogenicactivity was observed at higher surfactant loads (Garcia et al.,2006). A direct relation was observed between toxicity increaseand increase of alkyl chain length, when the available surfactantfraction in the aqueous phase was used to evaluate toxicity of LAShomologues (Garcia et al., 2006). In contrast to LAS, toxicity ofQAC on methanogens was decreased by increasing alkyl chainlength (Garcia et al., 1999). Tezel et al. (2006) reported that QACwere inhibitory to this group of microorganisms at concentrationof 25 mg L�1, while methanogenesis was more susceptible to thesecompounds comparing to acidogenesis. Watson et al. (2012) inves-tigated inhibitory effect of ester QAC on AD. Acetylcholine chloridedid not inhibit methanogenesis for concentrations up to 300 mg L�1,whereas, decreased methane production and accumulation of VFAswas observed for concentration of lauroylcholine chloride equal to100 mg L�1.

Regarding pharmaceuticals, in an early study, Sanz et al. (1996)studied the toxicity of different antibiotics on AD. According to theresults, the inhibiting concentrations as well as the way of

nants’ concentrations in sludge samples.

A.S. Stasinakis / Bioresource Technology 121 (2012) 432–440 435

inhibition were differentiated significantly according to the com-pound. Erythromycin did not inhibit biogas production for concen-trations up to 250 mg L�1, while some antibiotics such asdoxycline, tylosine, streptomycine and neomycine had partiallyinhibitory effects, interfering the activity of propionic-acid and bu-tyric-acid degrading bacteria. Moreover, chlortetracycline andchloramphenicol were powerful inhibitors of AD, inhibiting aceto-clastic methanogenic Archaea. Lallai et al. (2002) investigated theeffects of amoxicillin, oxytetracycline and thiamphenicol on biogasproduction and methane concentration in batch reactors. Thiam-phenicol had a considerable effect on methane production at con-centrations of 80 mg L�1, a 25% decrease of methane productionwas observed when 60 mg L�1 of amoxicillin were added, whileno effect was observed for oxytetracycline concentrations up to250 mg L�1. Stamatelatou et al. (2003) observed no effect of car-bamazepine on the AD process when 1 and 10 mg L�1 were addedin short term and long term experiments, respectively. Fountoula-kis et al. (2004) studied the effect of six pharmaceuticals on aceto-clastic methanogens for concentrations ranging up to 400 mg L�1.A 50% inhibition on specific methanogenic activity was observedfrom 30 mg L�1 (for propranolol) to more than 400 mg L�1 (for clo-fibric acid and sulfamethoxazole). The inhibition mechanism wasdirectly correlated with the affinity of pharmaceuticals to sorbonto anaerobic sludge. Gartiser et al. (2007) studied inhibition of16 antibiotics on anaerobic bacteria using ISO 13641 test and re-ported that most antibiotics showed only moderate inhibition ef-fects with EC50 values ranging between 24 mg L�1 (imipenem)and more than 1000 mg L�1 (amoxicillin, benzylpenicillin and ce-furoxime). On the contrary, metronidazol was much more toxicto anaerobic bacteria with an EC50 of 0.7 mg L�1.

Only sporadically data is available for the effects of other organ-ic ECs on sludge AD. Gavala et al. (2003) reported that accumula-tion of DEHP in the anaerobic digester had a negative effect ondi-n-butyl phthalate ester (DBP) and DEHP removal rates as wellas on biogas production. Experiments with methylbenzotriazole(MeBTr) showed significant decrease in methanogenic activityand volatile solids production in the presence of 300 mg L�1 of thiscompound (Gruden et al., 2001).

During the last years, the effects of nanoparticles on sludge ADhave been investigated in some papers. Nyberg et al. (2008) re-ported that the exposure of anaerobic sludge on C60 fullereneshad no significant effect on methanogenesis and community struc-ture. Mu and Chen (2011) reported that the presence of 30 mg ZnOnanoparticles g�1 TSS inhibited methane production by 18.3%. Thetoxic effects of these nanoparticles were mainly due to the releaseof Zn2+, while the steps of hydrolysis and methanation on sludgeAD were mainly affected. A lower abundance of methanogenesis Ar-chaea was observed when higher concentrations of ZnO nanoparti-cles were used. In another study, nano-TiO2, nano-Al2O3 and nano-SiO2 showed no inhibitory effect on sludge AD in doses up to150 mg g�1 TSS (Mu et al., 2011). Yang et al. (2012) reported no ef-fect of silver nanoparticles on biogas and methane production forconcentrations up to 40 mg Ag L�1, while methanogenic populationand diversity remained also unchanged. Finally, Garcia et al. (2012)investigated the effects of CeO2, TiO2, Ag and Au nanoparticles onmesophilic and thermophilic anaerobic bacteria. Nano-Au andnano-TiO2 presented no or low toxicity to anaerobic biomass, whilenano-CeO2 was the most toxic (EC50 value equal to 0.26 mg mL�1

for mesophilic biomass, <0.32 mg mL�1 for thermophilic biomass).

4. Fate of emerging contaminants during sludge anaerobicdigestion

The comparison of published data reveals considerable varia-tions on the removal of ECs during sludge AD. Data originated from

the operation of lab-scale anaerobic digesters under different con-ditions is presented in Table 2. The role of biodegradation andsorption on ECs’ removal and the factors that affect these mecha-nisms are discussed below.

4.1. Role of biodegradation

Operational parameters such as retention time and temperature(El-Hadj et al., 2006; Patureau et al., 2008) have been reported toinfluence biodegradation of DEHP and NPE during sludge AD, whilethey did not cause any effect on biodegradation of some pharma-ceuticals, synthetic musks and estrogens (Carballa et al., 2006,2007a).

Factors such as microbial population, target compounds’ bio-availability and co-metabolic phenomena have also been foundto affect biodegradation of some micropollutants. Biomass accli-matization favored biodegradation of diclofenac, diazepam andestrogens (Carballa et al., 2006, 2007a), indicating that the abun-dance of micropollutant degrading microorganisms is a crucial fac-tor for their removal in anaerobic digesters. According to Shin et al.(2010), poor bioavailability limited biodegradation of some poly-brominated diphenyl ethers (PBDEs) in anaerobic digesters, anobservation consistent with previous results noticing that thereis a threshold concentration, below of which dehalogenation pro-cess cannot take place (Cho et al., 2003). On the other hand, LASbiodegradation in anaerobic digesters was observed only at lowconcentrations, possibly due to the inhibition caused to activemicroorganisms at higher concentrations of these compounds(Angelidaki et al., 2004). The co-existence of a readily biodegrad-able carbon source enhanced anaerobic degradation of NP (Changet al., 2005) due to stimulation of overall metabolism. Similar re-sults have also reported for other groups of micropollutants (Barretet al., 2010a).

Finally, the physico-chemical properties of the target com-pounds as well as the sludge characteristics (Paterakis et al.,2012) seem also to play an import role on anaerobic biodegrada-tion, affecting bioavailability of micropollutants and accessibilityof readily biodegradable carbon to microorganisms (Barret et al.,2010b). Information about the biodegradation of different groupsof ECs on sludge anaerobic digestion is given below.

4.1.1. SurfactantsAnaerobic recalcitrance of LAS has been reported by Garcia et al.

(2005). On the other hand, other studies reported that LAS can bebiodegraded during sludge AD (Angelidaki et al., 2004; Carballaet al., 2007b). According to Sanz et al. (2003), LAS biodegradationwas higher in the absence of external co-substrates, indicating thatthese compounds can be partially used as carbon and energysource by anaerobic bacteria. No significant differences on their re-moval were observed in experiments performed under mesophilicand thermophilic conditions (Carballa et al., 2007b).

Regarding NPE, it is believed that under anaerobic conditionsbiotransformation commences at the hydrophilic part of the mole-cule and as a result ethoxylate chain of higher NPE is shortened un-til NP1EO, NP2EO and NP are formed (Lu et al., 2008). NP1EO andNP2EO are biodegraded to some extent during AD, producing NP(Patureau et al., 2008), while this compound does not undergo fur-ther transformation and accumulates on biosolids (Hernandez-Ra-quet et al., 2007; Janex-Habibi et al., 2009). On the other hand,other authors reported NP biotransformation during sludge AD(Chang et al., 2005; Patureau et al., 2008; Paterakis et al., 2012).According to Chang et al. (2005), NP biodegradation followedfirst-order kinetics, while different microbial strains were isolatedfrom sludge, being capable to use NP as carbon source. It is likelythat the anaerobic degradation of NP is slower compared to its for-mation from NPE. However, so far there are no studies reporting

Table 2Removal of emerging contaminants during sludge anaerobic digestion.

Compound Target compound initialconcentration (units)

Mesophilicanaerobic digestion

Thermophilicanaerobic digestion

Sludgeused

Reference

Removal(%)

SRT(days)

Removal(%)

SRT(days)

Galaxolide, HHCB 4–400 (lg L�1) 65 ± 15 20, 10 67 ± 16 10, 6 MixSa Carballa et al. (2006)Tonalide, AHTN 4–400 (lg L�1) 60 ± 8 20, 10 67 ± 15 10, 6 MixS Carballa et al. (2006)Carbamazepine, CBZ 4–400 (lg L�1) 0 20, 10 0 10, 6 MixS Carballa et al. (2006)Diazepam, DZM 4–400 (lg L�1) 60 ± 18 20, 10 38 ± 21 10, 6 MixS Carballa et al. (2006)Ibuprofen, IBF 4–400 (lg L�1) 40 ± 15 20, 10 47 ± 10 10, 6 MixS Carballa et al. (2006)Naproxen, NPX 4–400 (lg L�1) 87 ± 5 20, 10 91 ± 5 10, 6 MixS Carballa et al. (2006)Diclofenac, DCF 4–400 (lg L�1) 60 ± 18 20, 10 73 ± 9 10, 6 MixS Carballa et al. (2006)Iopromide, IPM 4–400 (lg L�1) 23 ± 15 20, 10 23 ± 11 10, 6 MixS Carballa et al. (2006)Sulfomethoxazole, SMX 4–400 (lg L�1) 99 ± 1 20, 10 99 ± 1 10, 6 MixS Carballa et al. (2006)Roxithromycin, ROX 4–400 (lg L�1) 85 ± 15 20, 10 95 ± 5 10, 6 MixS Carballa et al. (2006)Estrone + 17b-Estradiol, E1 + E2 4–400 (lg L�1) 85 ± 10 20, 10 85 ± 5 10, 6 MixS Carballa et al. (2006)17a-Ethinylestradiol, EE2 4–400 (lg L�1) 85 ± 5 20, 10 75 ± 15 10, 6 MixS Carballa et al. (2006)17a-Ethinylestradiol, EE2 18 ± 4d/10 ± 2e (lg kg�1 dw) 34/4 30 43/14 15 PSb/MixS Paterakis et al. (2012)

9 ± 1f/10 ± 2g (lg kg�1 dw)17b-Estradiol, E2 9 ± 1d/6 ± 1e (lg kg�1 dw) �324/�325 30 �367/�621 15 PS/MixS Paterakis et al. (2012)

6 ± 3f/3 ± 2g (lg kg�1 dw)Estrone, E1 158 ± 14d/90 ± 21e (lg kg�1 dw) 79/70 30 96/68 15 PS/MixS Paterakis et al. (2012)

64.3 ± 2.5f/32.3 ± 2g (lg kg�1 dw)Estriol, E3 9 ± 1d/8 ± 1e (lg kg�1 dw) 45/43 30 17/4 15 PS/MixS Paterakis et al. (2012)

6 ± 1.5f/5 ± 1g (lg kg�1 dw)Estrone sulfate conjugate,

E1–3S7.6 ± 1.5d/7 ± 1.5e (lg kg�1 dw) 36/21 30 30/28 15 PS/MixS Paterakis et al. (2012)4 ± 1f/4 ± 1g (lg kg�1 dw)

Nonylphenol, NP 0.3 ± 0.1d/0.23 ± 0.1e (mg kg�1 dw) 0/100 30 50/100 15 PS/MixS Paterakis et al. (2012)0.23 ± 0.1f/0.1 ± 0.1g (mg kg�1 dw)

Nonylphenol carboxylates, NPEC 26.5 ± 0.1d/242 ± 0.1e (mg kg�1 dw) – 30 >�1000/�5800 15 PS/MixS Paterakis et al. (2012)215/0.00350.1 ± 0.1f/0.08 ± 0.1g (mg kg�1 dw)

Nonylphenol mono- anddiethoxylate, NP1–2EO

2.1 ± 0.5d/1.7 ± 0.5e (mg kg�1 dw) 88/�0.0274 30 2.5/100 15 PS/MixS Paterakis et al. (2012)15 ± 0.1f/90 ± 0.1g (mg kg�1 dw)

Nonylphenol polyethoxylates,NP3–12EO

1.5 ± 0.4d/0.7 ± 0.4e (mg kg�1 dw) 66/67 30 73/83 15 PS/MixS Paterakis et al. (2012)1.3 ± 0.25f/0.7 ± 0.3g (mg kg�1 dw)

Nonylphenol diethoxylate,NP2EO

0.45 ± 0.08 (lM L�1) 2.63 20 – – ASc Hernandez-Raquet et al.(2007)

Nonylphenol monoethoxylate,NP1EO

1.48 ± 0.2 (lM L�1) 3.76 20 – – AS Hernandez-Raquet et al.(2007)

Nonylphenol, NP 16.56 ± 1.62 (lM L�1) 0.4 20 – – AS Hernandez-Raquet et al.(2007)

Diethyl-phthalate, DEP 11069 ± 3571 (lM L�1) 92.8 26.5 – – MixS Parker et al. (1994)Di-n-butyl phthalate, DBP 12400 ± 4101 (lg L�1) 93.8 26.5 – – MixS Parker et al. (1994)Butyl benzene phthalate, BBP 9431 ± 2700 (lg L�1) 92.8 26.5 – – MixS Parker et al. (1994)Bis(2-ethylhexyl) phthalate, DEHP 10100 ± 3260 (lg L�1) 61.0 26.5 – – MixS Parker et al. (1994)

a MixS: mixture of primary + excess activated sludge.b PS: primary sludge.c AS: excess activated sludge.d In PS feeding mesophilic reactor.e In MixS feeding mesophilic reactor.f In PS feeding thermophilic reactor.g In MixS feeding thermophilic reactor.

436 A.S. Stasinakis / Bioresource Technology 121 (2012) 432–440

biodegradation kinetics of different nonylphenols, as experimentalproof of this hypothesis.

QAC exhibit limited or no biodegradation in anaerobic digesters(Tezel et al., 2006; Tezel and Pavlostathis, 2009). Moreover, underanaerobic conditions, no evidence of mineralization have beennoticed for QAC containing alkyl or benzyl groups, possibly dueto the highly reduced nature of these substituted groups (Garciaet al., 2000). On the other hand, in a recent study, some esterQAC were transformed to methane, when concentrations lowerthat those causing inhibition to methanogens were used (Watsonet al., 2012). Abiotic hydrolysis of target compounds was initiallyobserved, following by biodegradation of hydrolysis products tomethane, carbon dioxide and ammonia (Watson et al., 2012).

4.1.2. Personal care productsDespite the frequent detection of these emerging compounds in

sludge samples, their fate during sludge AD has been studied inlimited papers. Regarding TCS and TCC, using data from full-scaleSTPs, some authors reported that AD is expected to have little orno impact on their removal (Heidler and Halden, 2009). To the bestof our knowledge, there are no results for the fate of these com-

pounds in lab-scale sludge anaerobic experiments. On the otherhand, in a recent study, significant biodegradation of TCS was re-ported under methanogenic conditions when acetate was used asco-substrate. The major biotransformation products were phenol,catechol and 2,4-dichlorophenol; whereas different bacteria ableto utilize this compound were isolated (Gangadharan PuthiyaVeetil et al., in press).

Kupper et al. (2006), applying mass balances in a Swiss STP, re-ported that galaxolide, tonalide, celestolide, phantolide and cashm-eran were removed by 50% during sludge AD. Carballa et al. (2006)observed average removal of galaxolide and tonalide during meso-philic and thermophilic digestion, ranging between 60% and 70%(Table 2). On the contrary, Clara et al. (2011) reported no or onlyslight removal of galaxolide, tonalide, cashmerane, celestolide,phantolide and traesolide during sludge AD.

4.1.3. PharmaceuticalsGolet et al. (2003) reported no significant removal of fluoro-

quinolones under methanogenic conditions of the sludge digesters.Carballa et al. (2006, 2007a) studied the removal of several phar-maceuticals during mesophilic and thermophilic AD. Excepting

A.S. Stasinakis / Bioresource Technology 121 (2012) 432–440 437

carbamazepine, all the other nine pharmaceuticals were removedto a significant extent (Table 2). Similar results for carbamazepinerecalcitrance during sludge AD have been reported in an earlierstudy (Stamatelatou et al., 2003). Gartiser et al. (2007) studiedthe anaerobic biodegradability of nine antibiotics using ISO11734 test. According to the results, only benzylpenicillin showedcertain ultimate biodegradation after 60 days. In a recent study,Lahti and Oikari (2011) reported that diclofenac was recalcitrantduring experiments with anaerobically digested sludge. On theother hand, bisoprol removed by 14% in a period of 161 days, whilenaproxen was completely removed in 14 days. Formation of 6-O-desmethylnaproxen was observed during anaerobic biodegrada-tion of naproxen.

4.1.4. EstrogensSo far contradictory results have been presented for the fate of

natural or synthetic estrogens during sludge AD. This is probablydue to the difficulties often observed during analysis of these com-pounds in sludge matrix (Esperanza et al., 2007; Citulski andFarahbakhsh, 2010), as well as to the unclear yet behavior of estro-gens conjugates during wastewater and sludge treatment (Des Meset al., 2008). Andersen et al. (2003) observed similar inlet and out-let loads of E1 and E2 in a mesophilic anaerobic digester, suggest-ing that estrogens were not degraded under methanogenicconditions. Esperanza et al. (2007) reported mass removal efficien-cies for E1, E2, E3, EE2, testosterone, androstenedione and proges-terone, ranging from 60% to 77% across the anaerobic digester.However due to some analytical problems observed, strict conclu-sions could not be drawn from that study. In lab-scale experiments,Carballa et al. (2006) observed significant degradation of estrogensunder mesophilic and thermophilic conditions and for a widerange of SRT (Table 2). Des Mes et al. (2008) observed the transfor-mation of E1 to E2 in anaerobic batch studies using different typesof sludge. However, no significant loss of the sum of E1 and E2 wasobserved, indicating no further breakdown of E2. Additionally, noEE2 removal was observed for any sludge type. Muller et al.(2010), using mass balances, reported that E1, E2 and E3 removalduring sludge AD was not effective (<40%), whereas the outletmass flow for EE2 was higher comparing to the inlet. In a recentstudy, Paterakis et al. (2012) reported that the biodegradation oftotal steroid estrogens (sum of E1, E2, E3, E1-3S and EE2) was high-er than 50% during primary sludge mesophilic and thermophicdigestion, while lower removal was observed for mixed sludge un-der mesophilic and thermophilic conditions (Table 2). E1 was re-duced to E2, while the de-conjugation of E1-3S to E1 appeared tobe minor, not exceeding 7% for all conditions used.

4.1.5. Phthalate acid estersSome data is available for the fate of phthalates during AD.

According to Shelton et al. (1984) degradation of these compoundsis related to the size of alkyl side chain. Their anaerobic degrada-tion involves initially sequential hydrolysis of the two ester side-chains, while afterwards phthalic acid and alkyl alcohols are con-verted to methane and carbon dioxide. Having in mind that DEHPrepresents 90% of total phthalate production (El-Hadj et al., 2006),its fate during sludge treatment has taken greater attention com-paring to the other compounds during the last 20 years.

Shelton et al. (1984) reported that dimethyl (DMP), diethyl(DEP) and di-n-butyl (DBP) phthalate esters were degraded withina week in AD batch experiments, butyl benzene phthalate (BBP)decreased by 75% in two weeks, whereas di-n-octyl phthalate(DOP) and DEHP were persistent. Gavala et al. (2003) reported low-er removal rates of DEHP comparing to other phthalates (Table 2).Marttinen et al. (2003) performed mass balances in a full scale STPand reported that DEHP was removed at a percentage of 32% dur-ing sludge AD, while Angelidaki et al. (2000) reported that its

degradation was dependent on the inoculum used. El-Hadj et al.(2006) studied the fate of DEHP during sludge AD under differentoperating conditions and reported higher removal rates underthermophilic conditions and higher SRT values. According toGavala et al. (2003), degradation of DEP, DBP and DEHP is de-scribed by first-order kinetics.

4.1.6. Brominated flame retardantsGerecke et al. (2006) reported half-lives lower than one day for

tetrabromobisphenol A (TBBPA) and hexabromocyclododecane(HBCD) in experiments with anaerobically digested sludge, whilea value of 700 days was calculated for decabromodiphenyl ether(DecaBDE). Pseudo first-order biodegradation constants of TBBPAand HBCD were not dependent on the occurrence of primers oradditional nutrients. Slow degradation of DecaBDE by mesophilicanaerobic sludge has also been reported by Gerecke et al. (2005),while its reductive dehalogenation was indicated by the formationof octa- and nonabromodiphenyl ether congeners. Shin et al.(2010) reported that the concentrations of BDE 47, 99, 100 and209 decreased by a percentage of 22–40% in a period of 238 daysin batch experiments performed under mesophilic conditions.Among the aforementioned compounds, the degradation of BDE209 was more rapid; reaching almost 40% in a period for 28 days.Microbial reductive debromination was considered as possiblecause of target compounds’ loss; however no simultaneous identi-fication of debrominated products was performed in this study. Onthe other hand, the concentrations of BDE 138, 153, 154 and 183,remained stable during this experiment.

4.1.7. Other emerging contaminantsThe removal of tributyltin (TBT) during sludge AD has been

studied using mass balances in full-scale plants (Fent, 1996) andin laboratory experiments (Voulvoulis and Lester, 2006). In allcases, its removal was minimal.

Regarding perfluorinated compounds, use of mass balances in afull-scale STP showed no change of perfluorodecane sulfonate (PFDS),perfluorodecanoate (PFDA) and 2-(N-ethylperfluorooctanesulfonam-ido) acetic acid (N-EtFOSAA) mass during sludge AD (Schultz et al.,2006). In the same study, mass flows of perfluorooctane sulfonate(PFOS), 2-(N-methylperfluorooctanesulfonamido) acetic acid (N-MeFOSAA) and perfluorononanoate (PFNA) were increased at the out-let of digester, indicating the possible degradation of their precursorsduring the process. The non-biodegradability of PFCs has been ob-served in several studies with different microorganisms (Fromeland Knepper, 2010). There was only one study presenting PFOS andperfluorooctanoic acid (PFOA) anaerobically biodegradation, how-ever no clear conclusions could be excluded as no increase in fluorideconcentrations was observed (Schroder, 2003). Despite the fact thatdefluorination is a thermodynamically favored reaction and microor-ganisms could obtain energy for growth from reductive defluorina-tion; it seems that this mechanism is hindered by the kineticstability of the C–F bond (Parsons et al., 2008).

Experiments with mesophilic anaerobic biomass showed thatMeBTr was not degraded (Gruden et al., 2001). On the other hand,benzotriazole (BTr), MeBTr and 5-chlorobenzotriazole (CBTr) weredegraded slowly under anaerobic conditions in experiments withdigested sludge (Liu et al., 2011). Half-life values, ranging from 44(CBTr) to 144 days (BTr), were calculated for the target compounds,while dechlorination was the main biotransformation pathway forCBTr. To the best of our knowledge, no results are available for thefate of benzothiazoles and nanoparticles during sludge AD.

4.2. Role of sorption

Apart from contributing to the removal of ECs from the waterphase, sorption may also influence their biodegradation rates,

Table 3Distribution coefficient Kd values of emerging contaminants in anaerobically digested sludge.

Compound Mesophilic Kd (L kg�1) Thermophilic Kd (L kg�1) Reference

Galaxolide, HHCB 13300 ± 5500 6709 ± 2980 Carballa et al. (2008)Tonalide, AHTN 15200 ± 7800 10019 ± 3343 Carballa et al. (2008)Carbamazepine, CBZ 35.4 20.2 Carballa et al. (2008)Ibuprofen, IBF 37.8 ± 14.2 21.0 ± 11.7 Carballa et al. (2008)Naproxen, NPX 10.8 12.4 Carballa et al. (2008)Diclofenac, DCF 65.5 ± 23.1 62.0 ± 52.0 Carballa et al. (2008)Iopromide, IPM 6.9 2.9 Carballa et al. (2008)Sulfamethoxazole, SMX 22.9 15.1 Carballa et al. (2008)Roxithromycin, ROX 83.3 13.7 Carballa et al. (2008)Estrone, E1 303 ± 59 151 Carballa et al. (2008)17b-Estradiol, E2 461 ± 212 – Carballa et al. (2008)17a-Ethinylestradiol, EE2 432 ± 168 135 ± 24 Carballa et al. (2008)C10LAS 343 – Garcia et al. (2005)C12LAS 4119 – Garcia et al. (2005)C14LAS 32577 – Garcia et al. (2005)Dodecyl trimethyl ammonium chloride, C12TMAa 3.14 ± 1.29 2.86 ± 1.49 Ismail et al. (2010)Hexadecyl trimethyl ammonium chloride, C16TMAa 48.36 ± 4.21 49.88 ± 6.24 Ismail et al. (2010)Dodecyl benzyl dimethyl ammonium chloride, C12BDMAa 10.76 ± 2.09 9.76 ± 1.44 Ismail et al. (2010)Hexadecyl benzyl dimethyl ammonium chloride, C16BDMAa 82.03 ± 24.13 99.66 ± 20.85 Ismail et al. (2010)

a Freundlich equation was applied, KF has been expressed as (mg/g VS)(L/mg)n.

438 A.S. Stasinakis / Bioresource Technology 121 (2012) 432–440

affecting bioavailability to microorganisms. Sorption depends onthe physicochemical characteristics of the suspended solids andthe chemicals involved, as well as on ambient conditions such astemperature, pH, ion strength, or presence of complexing agents.Anaerobic sludge is significantly different comparing to primaryand excess activated sludge. It contains lower percentage oforganic matter and lower carbohydrate and lipid content due tothe use of these materials as substrates supporting anaerobicmicrobial growth (Barret et al., 2010c). Experiments withhydrophobic micropollutants revealed that anaerobic treatmentenhanced sorption of these compounds to suspended solids andto dissolved and colloidal matter, compared to primary and sec-ondary sludge (Barret et al., 2010c). However, such an effect wasnot observed for more hydrophilic micropollutants in anotherstudy. Specifically, Carballa et al. (2008) reported that distributioncoefficient (Kd) values of some pharmaceuticals, estrogens andmusk fragrances were similar between anaerobic, primary andsecondary sludge. Moreover, no significant influence of AD opera-tional conditions (mesophilic or thermophilic) and sludge pretreat-ment (alkaline pretreatment, thermal pretreatment, ozonation)was observed for the sorption potential of these compounds(Carballa et al., 2008).

Despite the existence of a great number of data for Kd values ofECs in activated sludge, similar data have been reported for a lim-ited number of ECs in anaerobic sludge (Table 3). Moreover, thereare few studies reporting data on the fraction of ECs that is sorbedin an anaerobic digester. Due to the above, the role of sorption onthe total fate of ECs during sludge AD is not yet clear. Gruden et al.(2001) reported that between 10% and 30% of MeBTr was sorbed todigester sludge, while sorption was followed Freundlich model.Garcia et al. (2005) reported that LAS sorption affinity on anaerobi-cally digested sludge increased with LAS chain length. Carballaet al. (2008) reported that musk fragrances and estrogens weresorbed to 80–99% on particles in the anaerobic digester, whilesorption on digested sludge was of minor importance for pharma-ceuticals. Ismail et al. (2010) investigated the sorption capacity offour QAC on mesophilic and thermophilic anaerobic sludge. Sorp-tion was increased with increasing alkyl chain length. Moreover,benzyl containing homologues had higher Freundlich constant val-ues (KF) than non-benzyl containing QAC (Table 3). Finally, Yanget al. (2012) investigated the fate of silver nanoparticles on sludgeAD. According to the results, more than 90% of these nanoparticleswere removed from the liquid phase and associated with thesludge.

4.3. Use of other processes for enhancing emerging contaminantsremoval during sludge anaerobic digestion

During the recent years, sludge pretreatment methods andcombination of AD with other biological or chemical processeshave been used for improving ECs removal during sludge treat-ment. Regarding sludge pretreatment, the results that have beenpresented for the use of thermal hydrolysis (Carballa et al., 2006;Barret et al., 2010b), thermal pretreatment (Gavala et al., 2004)and ozone pre-treatment (Carballa et al., 2007b) were not encour-aging for micropollutants’ removal. This was probably due to thefact that some pretreatment processes decreased target com-pounds’ bioavailability (e.g., thermal processes) or in other casestarget compounds could not be attacked by the oxidizing agent(e.g., ozone) as they were strongly sorbed onto sludge. On the otherhand, combination of AD with other biological (e.g., post-aeration;thermophilic aerobic treatment; composting) or chemical pro-cesses (post-ozonation) enhanced ECs’ removal during sludgetreatment (Sanz et al., 2006; Hernandez-Raquet et al., 2007; Patu-reau et al., 2008).

5. Conclusions

Concentrations of ECs in sludge range between few lg kg�1

(estrogens, some pharmaceuticals, PFCs) to some g kg�1 (LAS).The inhibitory concentrations and the way of inhibition vary be-tween different compounds. ECs fate is often affected by SRT, tem-perature, sludge characteristics, adaptation of biomass andcompounds’ bioavailability. NPE, some pharmaceuticals and mostphthalates are biotransformed to some extent during sludge AD,while contradictory results have been published for LAS, NP, syn-thetic musks, estrogens and DEHP. No significant removal has beenobserved for TCS, TCC, fluoroquinolones, carbamazepine, TBT andPFCs; however the available data for these compounds is yet verylimited.

Acknowledgements

This study was implemented under the Operational Program‘‘Education and Lifelong Learning’’ and funded by the European Un-ion and Greek National Resources–THALIS: WATERMICROPOL(www.aegean.gr/environment/watermicropol). Author would liketo thank Eawag/ETH (Process Engineering Department) for hostinghim during the period that this paper was written.

A.S. Stasinakis / Bioresource Technology 121 (2012) 432–440 439

Appendix A. Supplementary data

Supplementary data associated with this article can be found, inthe online version, at http://dx.doi.org/10.1016/j.biortech.2012.06.074.

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