Impact of Cationic Polyelectrolyte Addition on Mesophilic Anaerobic Digestion and Hydrocarbon Content of Sewage Sludge

Abstract

The agricultural spreading of treated sewage sludge is a valid strategy in terms of circular economy for the management of this nutrient-rich waste. Anaerobic digestion (AD) can be applied to stabilize and hygienize sewage sludge, making it suitable for agricultural reuse, while producing biogas to be utilized as an energy vector. However, the presence of contaminants, including petroleum hydrocarbons, could limit the widespread agricultural utilization of sewage sludge. In this context, the impact of dewatering agents, such as cationic polyelectrolytes, on AD efficiency and hydrocarbon biodegradation has been poorly investigated, although it represents a noteworthy aspect when conditioned sludge is digested for agricultural use in centralized biogas plants. This work aims to elucidate the effect of cationic polyelectrolyte addition on biomethanation as well as the degradation and extractability of C₁₀-C₄₀ hydrocarbons during mesophilic AD of sewage sludge. The addition of 26.7 g/kg_TS of cationic polyelectrolyte was observed to extend the AD lag phase, although similar methane yields (573–607 mL_CH4 per g of degraded volatile solids) were observed for both conditioned and raw sludge. Furthermore, a significant impact on hydrocarbon degradation was observed due to chemical conditioning. Indeed, this work reveals that cationic polyelectrolytes can affect hydrocarbon extractability and suggests moreover that the presence of natural interferents (e.g., biogenic waxes) in sewage sludge may lead to an overestimation of potentially toxic C₁₀-C₄₀ hydrocarbon concentrations, potentially limiting the application of sludge-derived digestates in agriculture.

1. Introduction

The disposal of sewage sludge represents one of the main issues for modern municipal wastewater treatment plants (WWTPs) due to the increased sewage sludge production at a global level [1], which is currently estimated at about 45 million tons per year [2]. Among the potential routes adopted for sewage sludge disposal, the agricultural application of sludge digestate as fertilizer/soil conditioner is encouraged by Sewage Sludge Directive 86/278/EEC as an effective alternative to chemical fertilizers, sewage sludge being rich in nitrogen, phosphorous, and valuable organic matter [3,4,5,6,7]. On the other hand, the attention being paid to the presence of potentially toxic contaminants in biosolids is increasing with an aim to prevent the pollution of soil and water, as highlighted in the Circular Economy Action Plan adopted in March 2020. Sewage sludge also contains natural organic matter (NOM), which has a binding effect towards substances such as hydrocarbons, as well as pathogens and heavy metals sometimes in potentially harmful concentrations [8,9,10]. Given the risks to the environment and human health due to the presence of pollutants and pathogens, proper sewage sludge treatment is necessary to stabilize NOM, prevent nutrient immobilization and phytotoxicity, and inactivate pathogens. NOM indeed has high complexation capacities, being described as a heterogeneous mixture of organic compounds and reactive functional groups with acid–base properties [11].

Total hydrocarbons (THs) are a class of recalcitrant organic compounds that can occur in sewage sludge from municipal WWTPs, being mainly collected from industrial areas and transported to sewers by rainwater. THs include both hydrocarbons of anthropogenic origin, mainly aliphatic and polyaromatic petroleum hydrocarbons (PHs), as well as hydrocarbons of biogenic origin, such as natural waxes secreted by plants [12,13]. Due to their negative impact on the environment, PHs are typically used as an indicator of stormwater quality [14]. The range for determining PHs in terms of the length of carbon atom chains is usually between C₁₀ and C₄₀ based on the composition of crude oil processing products [15]. Due to their high persistence in the environment, low biodegradability, and high lipophilicity, several aliphatic and polyaromatic PHs are highly toxic and have been categorized as persistent organic pollutants [16,17]. PHs’ concentrations in sewage sludge as reported in the literature range from tens to a few thousand mg per kg of dry matter [17,18]. The current legislation governing the landspreading of sewage sludge in many European countries has introduced limits in terms of PHs content [19]. Currently, the Italian legislation (L. 130/2018) has set the limit for C₁₀-C₄₀ hydrocarbons at 1000 mg/kg sludge. The presence of oil in the environment for millions of years has led to the evolution of microorganisms capable of degrading and using hydrocarbons as their main, sometimes exclusive, source of carbon and energy [20]. Several studies have focused on the degradation of hydrocarbons in contaminated soils and highlighted that microbial degradation follows a pattern in which simpler, less toxic, more bioavailable, and energy-efficient hydrocarbons are degraded earlier than their counterparts [21]. Moreover, it has been clarified that hydrocarbon biodegradation can occur under both aerobic and anaerobic conditions [20,22,23,24].

In municipal WWTPs, sewage sludge stabilization can be carried out via anaerobic digestion (AD) in the sludge streamline. During this process, the hydrolysis of organic macromolecules has been identified as the limiting phase of the degradation process [25]. Previous studies revealed that indigenous hydrocarbons in sewage sludge are more difficult to biodegrade compared to spiked fractions, which are more accessible [26]. Moreover, it has been shown that AD of sewage sludge with hydraulic retention time (HRT) in the order of months can lead to a reduction in PH concentrations of about 15–20% [27,28]. Due to the high costs for energy consumption and process management that a biological stabilization treatment section entails for a medium–small WWTP, the AD process is often performed in centralized plants [1,29,30,31] collecting conditioned sewage sludge from different WWTPs. However, the sludge entering centralized plants typically contains cationic polyelectrolytes, i.e., polymers that improve dewatering performance by particle–particle bridging and surface charge neutralization mechanisms [32]. Moreover, polyelectrolytes might be added to the wastewater in the mainstream treatment, i.e., during the clarification phases, both primary, as in the case of chemically enhanced primary treatment (CEPT) [33,34,35,36,37,38], or secondary, as in the case of secondary sedimentation tank influent addition (SSTIA) [39,40] to improve the settling properties of the sludge. As a result, conditioned sludge settling in the clarifiers is conveyed to the AD phase of the sludge streamline, if present. In Figure 1, three cases, i.e., (a), (b), and (c), of AD of conditioned sludge are described:

(a)

in the WWTP, there is an AD phase before sludge conditioning and no polyelectrolytes are added before AD. In this case, the order of sludge treatments is:

anaerobic stabilization -> chemical conditioning -> dewatering -> SS disposal

(b)

in the WWTP, either there is no AD or it is out of service, and stabilization is performed outside the WWTP in a centralized plant [1,29,30,31]. In this case, the order of sludge treatments is:

chemical conditioning -> dewatering -> centralized AD (outside the WWTP) -> SS landspreading (or another
disposal method)

(c)

in the WWTP sewage line, polyelectrolytes are used in CEPT or SSTIA systems to improve the settleability of the flocs in settling phases [33,34,35,36,37,38,39,40]. In this case, the order of treatments is:

sludge settling with polyelectrolyte addition in the wastewater streamline ->sludge pre-thickening -> AD (within
or outside the WWTP) -> conditioning (if needed) -> dewatering -> disposal.

Recent reports have suggested that the presence of conditioners in sewage sludge may alter AD yields both in terms of methane production and volatile solid (VS) destruction [1,41]. Furthermore, sewage sludge pre-treated with biopolymers has been observed to release complex organic substances during AD [42]. In many cases, these conditioners contain significant concentrations of C₁₀-C₄₀ hydrocarbons, which can increase the THs content of sewage sludge [43,44]. However, there is a lack of data available in the literature on the influence of polyelectrolytes on AD of sewage sludge, particularly on the degradation and quantification of C₁₀-C₄₀ hydrocarbons. This topic deserves immediate attention due to the increasing agricultural valorization of SS in many European countries [45].

Notwithstanding that monitoring the presence of PHs in sewage sludge applied in agriculture is required by the law, it is an analytical challenge due to the variety of interactions between hydrocarbons and the NOM contained in sludge and soil [18]. Currently, the reference methods for the determination of PHs concentration are the ISO16703:2004 and UNI-EN14039:2005, which refer, respectively, to soils and solid waste, less complex matrices than sewage sludge [46,47]. The major issues which arise when measuring the concentration of C₁₀-C₄₀ hydrocarbons in sewage sludge relate to the finding of the most suitable methodologies for (1) the storage and conservation of sludge samples, (2) the extraction of the C₁₀-C₄₀ hydrocarbon fraction, being mainly saturated and non-polar, (3) the removal of all the undesirable substances, including polar, hydrolyzed and carboxylate compounds, that can be co-extracted due to their affinity with the extraction solvent, and (4) the isolation of the analyte to be measured. Further problems arise in the definition of the operating parameters to be adopted in the chromatographic analysis [14]. The high level of uncertainty, together with the complexity of the analytical procedures, makes PHs’ concentration very difficult to monitor and control in municipal WWTPs sludge [14,18]. The methodology usually applied, based on the analysis of C₁₀-C₄₀ organic extracts via gas chromatography with flame-ionization detection (GC-FID), but also with other detection methods, results very often in undefined chromatographic humps, which are reported in the literature as an unresolved complex mixture (UCM) [48,49,50]. The integration of the area underlying the UCM constitutes the concentration of the cumulative parameter C₁₀-C₄₀ hydrocarbons, a mixture of aliphatic hydrocarbons ranging from n-decane (C₁₀H₂₂) to n-tetracontane (C₄₀H₈₂) [51]. From such data, the quantification and speciation of C₁₀-C₄₀ hydrocarbons in sewage sludge are challenging and exhibit poor reproducibility; therefore, an optimization of the applied method would be required.

The present study investigates the impact of cationic polyelectrolyte addition to sewage sludge on biomethane production, solid reduction, and degradation of C₁₀-C₄₀ hydrocarbons during mesophilic AD and provides new insights on the extractability and bioavailability of hydrocarbons in chemically conditioned sludge and digestate. Particular attention has been paid to clarifying the speciation of the extracted hydrocarbons. Moreover, an optimized method based on gravimetric quantification supported by the GC-MS analysis was developed for determining the hydrocarbon concentration in sewage sludge.

2. Materials and Methods

2.1. Origin, Characteristics, and AD of Raw and Conditioned Sludge

Two sewage sludge samples were collected from the sludge streamline of a municipal WWTP located in southern Italy, one, i.e., the raw sludge (RS), consisting of a pre-thickened mixture of primary and secondary sludge having a total solids (TS) content of 3.86(±0.15)%, the other, i.e., conditioned sludge (CS), being pre-thickened, chemically conditioned, and dewatered. The WWTP sludge streamline consists of thickening by gravimetric sedimentation, chemical conditioning by mixing a 22 m³/h sludge flow at 3% TS content with a 2.2 m³/h cationic polyelectrolyte (Dryfloc^® EM77A, SNF, France) flow at a concentration of 8 g/L, therefore with a final concentration of 26,667 mg/kg_TS, and dewatering by industrial centrifugation to reach a TS content of 27.25(±0.42)%. No AD stage was present within the WWTP. RS was collected after pre-thickening upstream of the chemical conditioning stage, while CS was taken from a holding tank at the sludge streamline outlet (after centrifugation). Both sludge types (2 L each) were stored at 4 °C for the analytical determination of the hydrocarbon content.

The AD of the two sludge types was conducted in lab-scale batch reactors consisting of 1 L borosilicate glass bottles (SIMAX, Czech Republic) filled with 630 mL of each sludge. The collected sludge used for the experiments was subjected within the WWTP to a pre-thickening process ensuring short-term anaerobic conditions triggering the activation of methanogens and immediate methane production under anaerobic conditions, as observed in preliminary AD tests. In our laboratory, RS was used without any dilution, while CS was diluted with tap water to reach a TS content of 4.47(±0.09)%, in order to have the same initial VS concentration of about 23 g/L in all reactors, both for RS and CS. Reaching similar VS contents was necessary to correctly compare the biomethanation and hydrocarbon removal potential of the two sludges, as the solid content may influence AD performance [1]. The experiments were conducted in triplicate at mesophilic temperature (35.0 ± 0.5 °C) by placing the bottles in a thermostatic bath for a period of 80 days. Such a long digestion period was chosen to target significant degradation efficiencies that could be also compared to the data available in the literature obtained with HRTs in the order of several months [27,28]. Mixing was performed once a day by manual shaking before sampling.

The two sludge types were characterized in terms of TS, VS, THs and C₁₀-C₄₀ concentrations, before AD as well as after 80-day AD, during which the biomethane volume produced by each digester was measured daily. The volumetric biomethane production was measured by the passage of the biogas through a double-column system consisting of a first absorption column filled with a 17% (w/w) NaOH solution (CAS n. 1310-73-2, Carlo Erba Reagents, Italy) for CO₂ removal, and a second column containing distilled water which allowed for the produced biomethane volume to be quantified through the liquid displacement method. The specific biomethane production (SBP) of mesophilic AD tests was calculated as the normalized volume of biomethane produced over the total mass of VS degraded during AD.

2.2. Determination of TS and VS

The TS and VS contents of the two sludge types before and after AD were determined by following the APHA-AWWA-WEF Standard Method 2450 [52] through the use of a laboratory oven (Argolab TCN 115, Italy), a muffle furnace (Asal ZB/1, Italy), and a precision scale (Sartorius 1801, Germany).

2.3. Hydrocarbon Analysis in Sludge

2.3.1. Gravimetric Determination of TH Concentration

The measurement of the TH content in the two types of sewage sludge collected was carried out by applying an optimized methodology based on the gravimetric determination of the oil and grease content in sludge, which was developed about 40 years ago by the Italian Research Council [53]. In this study, a series of functional changes were made to optimize both the alignment of the method with a GC-MS method (see Section 2.3.2) and the repeatability and reproducibility of the data. Moreover, the single components of the analyzed mixture could be identified through their mass spectra based on the MS NIST98 database. The developed procedure consisted of a series of sample treatment steps which are described below.

Sample Pretreatment for the TH Extraction

At the end of the 80-day AD, the stored sludge and the digestate collected from the reactors were centrifuged at 4600 rpm with a Thermo IEC Centra GP8R (Artisan Technology Group, USA) at ambient temperature for 10 minutes. Subsequently, the supernatant fraction was removed, and the remaining part was dried at 35.0(±1.0) °C for 7 days. After drying, the sludge samples reached a TS content of 93.64 (±0.64)% and were ground in a mortar.

Soxhlet Extraction with Dichloromethane

Three pretreated samples of 10 g for each sludge type were loaded into a cellulose fiber thimble (30 × 80 mm, G.E. Whatman, USA) and inserted in the 100 mL extraction chamber of a Soxhlet apparatus. A battery of six Soxhlet extractors, equipped with a six-seat electric mantle (G. Vittadini, Italy) and Allihn condensers, was used to carry out parallel extractions. In the boiling flasks, 200 mL of dichloromethane (CH₂Cl₂, CAS n. 75-09-2, Carlo Erba Reagents, Italy) was loaded in each extractor. After an extraction time of 11 h, the enriched solvent was removed and purified.

Clean-Up and Filtration of the Enriched Solvent

The first clean-up operation was aimed to remove water from the enriched solvent by mixing it with anhydrous sodium sulphate (Na₂SO₄, CAS n. 7757-82-6, Carlo Erba Reagents, Italy) to obtain Na₂SO₄·10H₂O. The next cleaning procedure was aimed at removing the polar interference by eluting the solvent through a 400 mm Pyrex^® glass burette, filled with 4 g of Florisil^® (30-60 U.S. Mesh, VWR Chemicals, Belgium) under a layer composed of 2 g of Na₂SO₄. The final clean-up operation consisted of filtration through 0.45 μm polypropylene filters (CAT n. 6878-1304, G.E. Whatman, USA).

Sample Concentration

A concentration step was carried out to extract the analyte, i.e., the TH fraction, from the solvent, through distillation. The low boiling point of the CH₂Cl₂ (39.6 °C at 1 atm) guarantees that this operation takes place without damaging the molecular structure of the analyte components. The operation was stopped when the volume of solvent was carefully reduced to about 2 mL.

Gravimetric Measurement

Weighing of the obtained analyte was carried out with a high-precision (0.001 mg) scale (Mettler Toledo MT5, USA). The sample was loaded on a tared weighing boat in borosilicate glass. After overnighting in a thermostat (Genlab MC200, UK) at 45 °C to guarantee the volatilization of residual solvent, the weighing boat containing the analyte was weighed, and the TH content in the initial 10 g of pretreated sludge sample was calculated.

2.3.2. Chromatographic-Spectrometric Evaluation of C₁₀-C₄₀ Hydrocarbons

C₁₀-C₄₀ hydrocarbons in sewage sludge were analyzed through a GC-based method. For this purpose, the ISO 16703:2011 method, conceived for the analysis of hydrocarbons in soils through FID [47], was adapted to MS detection, being much more sensitive and selective than FID [18]. A 5-point calibration curve was built with different concentrations of C₁₀-C₄₀ hydrocarbons obtained diluting an 8 g/L mineral oil standard mixture (Type A and B Standard mixture—69246, Sigma Aldrich, Switzerland) with retention time window standard solution (RTW Standard Solution—67583, Sigma Aldrich, Switzerland) on an Agilent GC 6850 gas chromatograph coupled with an Agilent MS 5973 mass spectrometer. The thermal curve was built by maintaining a temperature of 60 °C for 10 min, followed by an increase with a slope (dT/dt) of 11 °C/min up to 90 °C and of 4 °C/min up to 320 °C. The last temperature was held for 20 min, resulting in a total elution time of 90.22 min. The column was a Zebron^TM ZB-Semivolatiles Guardian capillary column (Phenomenex, USA) with a length of 30 m, an ID of 0.25 mm, and a thickness of 0.25 μm. The single quadrupole mass was maintained at 280 °C and fed with an electron multiplier voltage of 2200 V. The sample was recovered from the weighing boat downstream of the gravimetric measurement, diluted in acetone (CH₃COCH₃, CAS n. 67-64-1, Carlo Erba Reagents, Italy) up to 10 mL within a flask and then injected in the GC, which was run at 180 °C and with an injection flow of 1 mL/min. According to the classification of the Method for the Determination of Extractable Petroleum Hydrocarbons (EPHs), known as MADEP speciation [54], and exploiting the high selectivity of MS determination, two EPH fractions, i.e., L_EPH (lighter EPH, corresponding to the C₁₀-C₁₈ hydrocarbon fraction) and H_EPH (heavier EPH, corresponding to the C₁₉-C₄₀ hydrocarbon fraction), were identified and compared with each other to draw out the behavior of aliphatic hydrocarbons during AD of the two sludge types tested. Using the NIST98 database, the GC-MS analysis allowed for the identification of some of the species giving resolution peaks. The individual peaks for each sample of sludge were validated only if present on the chromatograms of each replicate and gave the same spectral identification.

2.4. Calculations

VS reduction in sewage sludge after AD was evaluated according to Koch [55] as follows:

$$\mathrm{VS} \mathrm{reduction} \mathrm{efficiency} (\%)=1 - \frac{\left({1-\mathrm{VS}}_{\mathrm{IN}}\right)}{{\mathrm{VS}}_{\mathrm{IN}}}\cdot \frac{{\mathrm{VS}}_{\mathrm{OUT}}}{\left({1-\mathrm{VS}}_{\mathrm{OUT}}\right)}$$

(1)

where VS_IN and VS_OUT are the VS contents of sewage sludge at the beginning and end of the AD process, respectively, expressed as a percentual fraction of the TS concentrations. With a similar approach, the removal efficiency of both THs and C₁₀-C₄₀ hydrocarbon fractions were calculated as follows:

$$\mathrm{Hydrocarbons} \mathrm{removal} \mathrm{efficiency} (\%)=1 - \frac{{\mathrm{C}}_{\mathrm{OUT}}}{{\mathrm{C}}_{\mathrm{IN}}}\cdot \left(\frac{{\mathrm{TS}}_{\mathrm{OUT}}}{{\mathrm{TS}}_{\mathrm{IN}}}\right)$$

(2)

where:

• C_IN and C_OUT are the hydrocarbons (THs or C₁₀-C₄₀) concentrations, respectively, before and after AD expressed as mg per kg of TS;
• TS_IN and TS_OUT represent the TS concentrations expressed as g/L.

3. Results and Discussion

3.1. Methane Productivity of Raw and Conditioned Sewage Sludge

Table 1. Parameters of RS and CS measured before (IN) and after (OUT) AD.

Parameter	Unit of Measure	RS	CS
TS IN	g/L	38.65 ± 1.52	44.71 ± 0.89
TS OUT	g/L	24.14 ± 0.24	32.97 ± 1.87
VS IN	g/L	23.72 ± 0.76	22.89 ± 0.46
VS OUT	g/L	11.23 ± 0.23	13.95 ± 0.12
VS IN	%TS	61.41 ± 0.43	51.20 ± 0.09
VS OUT	%TS	46.51 ± 0.48	44.00 ± 0.65
SBP80	NmLCH4/gVSremoved	607 ± 41	573 ± 9
TH IN	mg/kgTS	22,235 ± 1087	11,533 ± 549
TH OUT	mg/kgTS	15,579 ± 292	13,029 ± 138
TH IN	mg a	541.36 ± 26.46	324.82 ± 15.47
TH OUT	mg a	236.87 ± 2.09	270.74 ± 17.84
C10-C40 IN	mg/kgTS	11,706 ± 1432	5918 ± 430
C10-C40 OUT	mg/kgTS	6417 ± 955	6715 ± 489
C10-C40 IN	mg a	285.01 ± 34.87	166.66 ± 12.11
C10-C40 OUT	mg a	97.72 ± 15.49	139.00 ± 4.46
C10-C40/TH IN	%	52.65 ± 6.96	51.31 ± 4.47
C10-C40/TH OUT	%	41.19 ± 6.18	51.55 ± 3.79

Note: ^a mg per reactor.

reports the values of the relevant parameters measured on raw and conditioned sludge before and at the end of the AD process, including hydrocarbon concentrations. VS percentages expressed in terms of TS for the two sludges were slightly different from each other (61.41 ± 0.43%_TS and 51.20 ± 0.09%_TS for RS and CS, respectively), which was likely due to the inorganic fraction of the polyelectrolyte mixture. However, these values began remarkably close at the end of AD (46.51 ± 0.48%_TS and 44.00 ± 0.65%_TS for RS and CS, respectively), demonstrating that the more biodegradable fraction of the two sludges was effectively degraded in 80 days with and without the addition of polyelectrolytes. Therefore, the presence of cationic polyelectrolytes was not detrimental in terms of VS degradation by mesophilic AD. Indeed, Figure 2 shows that the SBP at the end of the 80-day monitoring period (SBP₈₀) of RS (607 ± 41 NmL_CH4/g_VSremoved) was very close to that of CS (573 ± 9 NmL_CH4/g_VSremoved). It should be highlighted that the production of biomethane at 30 days compared to the initial mass of VS was equal to 218.85(±7.68) NL_CH4/kg_VSin and 144.51(±2.11) NL_CH4/kg_VSin for RS and CS, respectively, perfectly in line with the values reported in the literature for AD of sewage sludge [1,56]. Therefore, sludge conditioning did not affect the biomethane yield of sewage sludge significantly. However, an impact on the kinetics of mesophilic AD was observed. CS reactors showed a 13-day lasting lag phase in biomethane production which could be attributed to the presence of polyelectrolytes in the sludge matrix. The longer lag phase observed for CS suggests that the presence of cationic polyelectrolytes in the sludge would result in a lower energy recovery as biomethane compared to RS at similar retention time in the digester. Sludge conditioning aims to favor the aggregation of sewage sludge particles to enhance dewatering [57,58]. On the other hand, increasing particle size can slow down the kinetics of biomethane production by prolonging the disintegration phase, which becomes a rate-limiting step of the AD process [59]. The increase of sludge particle size can also hinder the bioavailability of substrates to the hydrolytic biomass consortia and slows down hydrolysis [60]. Moreover, cationic polyelectrolytes could also exhibit some bacteriostatic activity, thus requiring a longer time (lag phase) for the biomass to adapt to the substrate [61]. Different starting conditions for the AD of the two sludges also resulted in lower solid depletion for CS compared to RS. Based on data reported in Table 1, TS reduction at the end of the AD period was 37.54(±2.54)% and 26.26(±4.43)% for RS and CS, respectively (Figure 3a), whereas VS reduction, calculated according to Equation (1), was 45.34(±1.46)% and 25.08(±1.98)% for RS and CS, respectively (Figure 3b).

3.2. Effect of Sludge Conditioning on Hydrocarbon Degradation and Extractability during AD

The effect of AD on both THs and C₁₀-C₄₀ hydrocarbons is outlined by their outlet mass and concentration (Table 1). AD could significantly abate the mass of THs and C₁₀-C₄₀ hydrocarbons in digested sludges. Similar removal efficiencies were observed for the two hydrocarbon fractions, calculated according to Equation (2) and shown in Figure 3c,d. However, at the end of the digestion process, the two sludges showed similar contents of both THs (15,579 ± 292 mg/kg_TS for RS and 13,029 ± 138 mg/kg_TS for CS) and C₁₀-C₄₀ hydrocarbons (6417 ± 955 mg/kg_TS for RS and 6715 ± 489 mg/kg_TS for CS). The concentration of C₁₀-C₄₀ hydrocarbons was always below the regulatory limit of 1000 mg/kg of sludge imposed by Italian legislation, before and after the digestion process. Estimation of such concentrations was made by dividing the mass of C₁₀-C₄₀ hydrocarbons (Table 1) by the mass of sludge (~0.6 kg) in each reactor at the end of the digestion process. Based on this calculation, RS and CS have respectively about 475 and 278 mg/kg sludge before AD, and respectively 163 and 232 mg/kg sludge after AD. The remaining hydrocarbons likely correspond to the more recalcitrant fraction that could not be biodegraded during the 80-day mesophilic AD process.

The most evident effect of polyelectrolyte addition on the hydrocarbon fraction is related to their change in extractability from CS after AD. Indeed, notwithstanding the similar VS concentration (and mass) in each reactor before and after the AD process (Table 1), a large difference in the mass of extracted TH per reactor was observed before AD (541.36 ± 26.46 mg for RS and 324.82 ± 15.47 mg for CS), while more similar values were observed after AD (236.87 ± 2.09 for RS and 270.74 ± 17.84 for CS). These results indicate that polyelectrolyte addition can modify hydrocarbons’ extractability and affect their quantification. NOM, indeed, has been widely described for its surfactant activity and capability of forming hydrophobic niches able to sequestrate hydrophobic compounds, which were proven to be exceedingly difficult to access even by the action of organic solvents [62]. From this point of view, possible interferences of the described interactions with the analytical determinations of both THs and C₁₀-C₄₀ hydrocarbons cannot be ruled out, and future analytical insights into the macrostructures of hydrophobic niches are needed.

Besides the differences between RS and CS in terms of mass removal efficiencies, the two sludge types have also shown different behavior in terms of distribution of the hydrocarbon fractions. Indeed, while a comparable C₁₀-C₄₀/TH ratio was measured for both sludges before AD, about 52%, this ratio became very different after AD, with 41.19(±6.18)% for RS and an almost unchanged 51.55(±3.79)% for CS. Such a decrease indicates that in RS AD affected the C₁₀-C₄₀ fraction more than THs. This can be explained by considering that TH concentration expresses the whole aliphatic hydrocarbon content in the sludge, while the C₁₀-C₄₀ fraction includes alkane compounds from n-decane to n-tetracontane. According to studies on hydrocarbon degradation during fermenting processes, degradative pathways with electron acceptors alternative to oxygen lead to the formation of aliphatic compounds of smaller size as well as substances such as fatty acids [20,63,64]. These compounds are not detected within the retention time window of the GC-MS identifying the C₁₀-C₄₀ fraction, although they are still detected with the gravimetric measurement, explaining the observed decrease of the C₁₀-C₄₀/TH ratio for RS. In CS, instead, the effect of chemical conditioning in aggregating and compacting the NOM might affect the biological degradation of the hydrocarbons without selectivity among C₁₀-C₄₀ and other aliphatic substances.

3.3. Impact of AD on Hydrocarbon Fractionation and Contaminant Degradation in Raw and Conditioned Sewage Sludge

The GC-MS chromatograms illustrated in Figure 4 point out a greater content of L_EPH for the CS than for the RS both before and after AD, which might be due to the presence of low molecular weight (MW) hydrocarbons in the flocculants added during the conditioning phase. By comparing Figure 4a,b, it appears that during AD there is a shift in the nature of the hydrocarbons, with an increase of the L_EPH fraction at the expense of H_EPH. This is in agreement with several studies reporting how the degradative pathways of the hydrocarbons lead to the reduction of molecular complexity by increasing the amount of smaller molecular groups [20,63].

An interesting comparison between the two sludge types is provided in

Table 2. EPH fractions UCM area ratios for RS and CS before (IN) and after (OUT) AD.

(a)	LEPH/HEPH	(b)	RS	CS	(c)	LEPH	HEPH
RS IN	0.38	LEPH OUT/IN	1.67	0.99	RS/CS IN	0.63	1.91
RS OUT	0.71	HEPH OUT/IN	0.89	0.85	RS/CS OUT	0.86	1.63
CS IN	1.14
CS OUT	1.34

a by calculating the ratios between L_EPH and H_EPH for RS and CS before and after the AD process. CS presents a L_EPH/H_EPH greater than 1, both before and after the AD process, which is higher than the L_EPH/H_EPH values (<1) observed for RS.

It can be observed that the area of the L_EPH UCM hump is lower for RS when compared to CS (Figure 4, Table 2c) and vice versa for the H_EPH area (RS/CS < 1 and RS/CS > 1, both IN and OUT). The H_EPH fraction was successfully degraded in both RS and CS, being H_EPH OUT/IN 0.89 and 0.85 for RS and CS, respectively (Table 2b). H_EPH in RS was more abundant than in CS and was partially converted into L_EPH during AD. Indeed, L_EPH OUT/IN was higher for RS (1.67) than for CS, which practically does not change (0.99).

An interesting consideration could be made on the nature of some substances identified through GC-MS analysis and the NIST98 database (Figure 4) and listed in

Table 3. Substances identified with GC-MS for RS and CS before (IN) and after (OUT) AD.

Peak	Sample	MW	Substance
1	CS OUT	202	methane, benzo, decahydro, cyclodecene
2	CS OUT	258	cyclopenta, hexahydro, hexamethyl, benzopyran
3	RS OUT	228	penta decanol
4	RS OUT	256	hexadecanoic acid
5	RS OUT	284	octadecanoic acid
6	RS, CS IN	355	dihydro, methyl papaverine
7	RS, CS IN	398	octadecanoic acid, dimethyl, dioxolanyl, methyl ester
8	RS IN	424	dodecanoic acid, hexadecyl ester
9	RS, CS IN	452	tetradecanoic acid, hexadecyl ester
10	RS, CS IN	480	hexadecanoic acid, hexadecyl ester
11	CS IN	514	propanoic acid, thio-bis, dodecyl ester
12	CS IN	553	hexadecanoic acid, pentadecyl, dioxanyl ester (trans-)

. It seems clear that both sludges before digestion (as can be seen for GC-MS peaks 6–12 in Figure 4 and Table 3) held a series of compounds, such as wax esters, probably of natural origin [12]. As shown by the chromatograms obtained before and after AD, a high portion of these natural compounds was decomposed by AD, leading to the formation of lighter molecules (indicated by peaks 1–5) such as organic molecules of lesser complexity and fatty acids. Based on the comparison between the L_EPH UCM areas of the two sludge types before AD, it can be hypothesized that part of the light hydrocarbons found in CS originated from the conditioning agents. The GC-MS profiles of CS before and after AD (Figure 4b) suggest that such light hydrocarbons feature an extremely low anaerobic biodegradability and/or bioavailability. Indeed, the L_EPH UCM area downstream of AD in CS remained unaltered (Figure 4b and L_EPH OUT/IN = 0.99 in Table 2b), and the lighter degradation products showed low identifiability (Table 3). This might be linked to the organization of the NOM of CS in compact flocs due to chemical conditioning, which could give rise, even downstream of the AD, to an entrapment effect on the compounds, making them less bioavailable [62].

Among the organic compounds listed in Table 3, several compounds disappeared or were reduced in concentration after AD. The compounds identified by peaks 6–12 are related to compounds belonging to the family of wax esters. Such compounds, initially present in the two sludge types, were partially degraded by AD. On the other hand, the substances indicated by peaks 3–5 for RS and 1–2 for CS were produced during AD. Comparing RS and CS after AD, quite different degradation products could be found. In detail, the alcohol and acids identified by peaks 4 and 5 were likely generated by the de-esterification of the hexadecanoic acid-hexadecyl-ester (peak 10) and the octadecanoic acid-dimethyl-dioxolanyl-methyl-ester (peak 7), respectively. Moreover, aromatic contaminants (peaks 1 and 2) were only identified in CS samples collected after AD. This is explainable by considering the effect of conditioning on contaminant retention by NOM. Such compounds, indeed, were released only after the degradation occurred on the NOM during AD. Aromatic compounds are indeed retained within the hydrophobic inner core of NOM through π-stacking interactions, among others [11,65]. It should be highlighted that the compounds indicated by peak 2 belong to the polycyclic musks family, which includes ubiquitous, persistent, and bioaccumulative pollutants suspected to cause lethal and sublethal effects on the exposed biota [66]. It is noticeable that, in both RS and CS, a sludge metabolite of papaverine (peak 6) was detected, probably due to the release in wastewater of anthropogenic drugs. However, AD showed a marked capability of reducing the concentration of this compound in both sludges, regardless of the sludge conditioning effect. The regulation of C₁₀-C₄₀ hydrocarbons in sewage sludge intended for agricultural reuse has as its primary scope the control of the release of potentially toxic PHs of anthropogenic origin. For this purpose, it is common to estimate the concentration of PHs in sludge as well as in other solid matrixes by measuring the C₁₀-C₄₀ UCM chromatographic area. The presence of biogenic waxes and their anaerobic degradation products could lead to a significant overestimation of the concentrations of anthropogenic C₁₀-C₄₀ hydrocarbons in sewage sludge. It is recommended to carry out toxicological studies to assess the potential toxicity of the molecules produced from the anaerobic degradation of the organic compounds contained in digestate produced from both raw and conditioned sewage sludge.

4. Conclusions

Our results show that the addition of cationic polyelectrolytes to sewage sludge has no significant detrimental impacts upon the solid reduction, biomethane yield, and degradation of THs and C₁₀-C₄₀ hydrocarbons by mesophilic AD. Indeed, AD of both RS and CS led to similar VS and hydrocarbon contents and SBPs in the digestates. In contrast, polyelectrolyte addition was observed to slow down the biomethanation kinetics of mesophilic AD by prolonging the initial lag phase. Also, the presence of cationic polyelectrolytes in the sludge was observed to affect hydrocarbon extraction and bioavailability in sewage sludge, especially of the lighter fraction, as clearly shown by the MADEP speciation results. Moreover, our results show that the presence of natural interferents, such as biogenic waxes, might affect the determination of the C₁₀-C₄₀ hydrocarbons, leading to an overestimation of the concentrations of potentially toxic C₁₀-C₄₀ hydrocarbons in sewage sludge. More research is recommended to investigate the impact of cationic polyelectrolytes on AD kinetics in continuous bioreactors and optimize the analytical method to specifically target anthropogenic C₁₀-C₄₀ hydrocarbons in sewage sludge and avoid their overestimation.