This study assessed the characteristics of biosolids of a pilot-scale anaerobic membrane bioreactor (AnMBR) treating municipal wastewater. The production of total solids (TS) and volatile solids (VS) was comparable to that reported for the extended aeration system at solids residence time (SRT) longer than 40 days. The yields of TS and VS were reduced as SRT increased from 40 to 100 days and increased with the addition of 26 mg/L of FeCl3. The AnMBR destroyed 60–82% of the VS loading in feed wastewater and hence it was concluded the biosolids met the requirements for vector attraction reduction for land application. The concentrations of volatile suspended solids and total suspended solids in the sludge were less than those reported after anaerobic digestion of conventional primary and secondary sludge mixtures, and hence dewatering of the waste stream may be required for some applications. The nutrient content in terms of total Kjeldahl nitrogen and total phosphorus was similar to that of anaerobically digested municipal sludges. The dewaterability of the biosolids was poorer than that reported for sludges from aerobic treatment and anaerobically digested sludges. Dewaterability was improved by addition of FeCl3 and reduced SRT. The biosolids met standards for land application with regards to the concentration of heavy metals but would need further treatment to meet Class B pathogen indicator criteria.

INTRODUCTION

Anaerobic membrane bioreactors (AnMBRs) are being recognized as a sustainable technology for wastewater treatment because the anaerobic process has low sludge production and low energy and nutrients requirements, and can generate methane as an alternative energy source (Metcalf & Eddy Inc. 2003; Smith et al. 2014). The integration of membrane modules into the anaerobic bioreactor (Smith et al. 2012) enables operation at extended solids residence times (SRTs), and hence good effluent quality can be maintained over a wide range of operating conditions (Huang et al. 2011).

The application of AnMBRs to authentic municipal wastewaters has been reported relatively recently (Lew et al. 2009; Gimenez et al. 2011; Huang et al. 2011; Martinez-Sosa et al. 2011; Smith et al.; 2012; Gao et al. 2014). In these studies, it has been found that extending the SRT resulted in improved removal of chemical oxygen demand (COD) and a concurrent increase in biogas production. Further, mixed liquor suspended solids (MLSS) concentrations increased. The improved removal of COD can be expected to result in a reduction in the generation of volatile suspended solids in the waste biosolids stream; however, the extent of reduction in biosolids production has not been quantified. Further, the increase in MLSS concentrations at extended SRTs could be expected to influence the dewatering of the produced biosolids but the extent of the impact has not been reported.

Other AnMBR operating strategies may also influence the properties of biosolids generated in AnMBRs. For example, the addition of FeCl3 to mitigate membrane fouling in an AnMBR treating municipal wastewater has been reported (Dong et al. 2015). The addition of FeCl3 enhanced the removal efficiencies of COD and biochemical oxygen demand, but showed no significant influence on the removal efficiencies of total Kjeldahl nitrogen (TKN) and total phosphorus (TP) and the methane yield. Superior membrane performance was attributed to the shift of particle size distribution to the particulate fraction and the reduced colloidal and soluble substances in the sludge. The addition of FeCl3 can be expected to increase the quantity of fixed suspended solids produced in the biosolids streams; however, the extent of the increase has not been quantified. The change in particle size distribution will also likely influence the dewaterability of the produced biosolids; however, this has also not been quantified.

These previous studies provide evidence that AnMBR operating conditions can influence bioreactor and membrane performance through modifying the biosolids characteristics. However, there is a lack of quantitative information in the literature that describes the characteristics of biosolids produced in AnMBRs in terms of subsequent biosolids handling and potential use in land application. In addition to understanding the quantity of biosolids that will be generated, an assessment of the feasibility for land application requires information on nutrient content as the concentration of TKN and TP, vector attraction reduction (VAR) and the concentrations of pathogens and heavy metals (McFarland 2001; Metcalf & Eddy Inc. 2003). Information on these properties of the biosolids generated in AnMBRs treating authentic municipal sewage is not available. This paper presents the quantity and quality of biosolids generated through treatment of authentic municipal wastewater in a pilot-scale AnMBR.

MATERIALS AND METHODS

A pilot-scale AnMBR that was fed with 3 mm screened sewage from the Burlington Skyway Wastewater Treatment Plant (ON, Canada) was employed in this study (Figure 1). A detailed description of the design and operating parameters of the system is presented in Table 1. The AnMBR consisted of a completely mixed anaerobic digester (AD) and a separate membrane tank (MT). The MT held a polyvinylidine fluoride hollow-fibre membrane module (GE ZeeWeed 500). The AD contents were mixed by recirculation with a positive displacement pump. In addition the AD contents were circulated through the MT using a centrifugal pump that withdrew mixed liquor from the bottom of the AD and pumped it to the bottom of the MT, after which it flowed to the top of the AD by overflow. This circulation mixed the MT contents and generated a cross-flow velocity that would enhance surface shear for membrane fouling control. Biogas produced in the AD was released from the head space of the AD and its production was measured by a gas flow meter (Aalborg GFM17). Biogas was recirculated through the MT with a blower to reduce membrane fouling.

Table 1

AnMBR operational parameters

Parameter Pilot AnMBR 
AD volume (L) 550 
MT volume (L) 80 
Membrane surface area (m25.4 
Membrane pore size (μm) 0.04 
Temperature (°C) 23 ± 1 
pH 6.7–6.8 
Mixing flow (L/h) 3,600 
Recirculation flow (L/h) 918 
Biogas sparging rate (m3/h at 20 °C and 1 atm) 0.786 
Hydraulic retention time (h) 8.5 
Parameter Pilot AnMBR 
AD volume (L) 550 
MT volume (L) 80 
Membrane surface area (m25.4 
Membrane pore size (μm) 0.04 
Temperature (°C) 23 ± 1 
pH 6.7–6.8 
Mixing flow (L/h) 3,600 
Recirculation flow (L/h) 918 
Biogas sparging rate (m3/h at 20 °C and 1 atm) 0.786 
Hydraulic retention time (h) 8.5 
Figure 1

Schematic of AnMBR system.

Figure 1

Schematic of AnMBR system.

Throughout the study the temperature of the AD was maintained using a heat tape that was controlled by a temperature controller which was informed by a temperature sensor in the digester. The pH of the AD was controlled through NaHCO3 addition, informed by a pH sensor in the digester. The operation and data acquisition were controlled using a programmable logic controller. The feeding of raw sewage and the wasting of anaerobic digester contents for SRT control were controlled on the basis of the weight of the pilot digester, which was monitored by load cells installed at the base of the pilot digester. The biosolids generated in this study were wasted directly from the AD and hence the concentrations of particulate species in the biosolids were the same as that of the AD.

The test plan facilitated an assessment of the impact of SRT and FeCl3 addition on biosolids production and characteristics. Prior to the study the pilot AnMBR was operated at a hydraulic retention time (HRT) of 8.5 hours and SRT of 70 days for 5 months without FeCl3 addition as the start-up phase. Subsequently the research was conducted in four phases at a constant HRT of 8.5 hours that corresponded to a membrane flux of 17 L/(m2·h). The pilot AnMBR was initially operated at an SRT of 70 days and fed with non-Fe-dosed sewage (Phase 1) to establish a base condition. Subsequent testing was conducted with the AnMBR operating at SRTs of 100, 70 and 40 days in Phases 2, 3 and 4 respectively and with addition of FeCl3 (26 mg/L) to the influent (Table 2). The FeCl3 dosage was based on previous work (Dong et al. 2015) that demonstrated this dosage could significantly mitigate membrane fouling while not affecting methane production. The impact of FeCl3 was assessed by comparing the data from Phases 1 and 3, and the impact of SRT was assessed by analyzing the data from Phases 2–4. Each phase was operated to achieve both biological steady state and to assess long-term membrane performance.

Table 2

Test plan

Phase SRT (d) FeCl3 dosage (mg/L sewage) Duration (d) Number of samples 
70 90 26 
100 26 80 24 
70 26 93 28 
40 26 178 50 
Phase SRT (d) FeCl3 dosage (mg/L sewage) Duration (d) Number of samples 
70 90 26 
100 26 80 24 
70 26 93 28 
40 26 178 50 

The wastewater samples were 24-hour composite samples collected by an auto-sampler. The mixed liquor samples were discretely collected. Wastewater and mixed liquor samples were collected twice a week from the outlets of the feed pump (upstream of the FeCl3 dosing) and anaerobic reactor, respectively. The number of the samples is showed in Table 2. The samples were analyzed for concentrations of total COD, soluble COD (0.45 μm filter), total solids (TS), total suspended solids (TSS), volatile solids (VS), volatile suspended solids (VSS), TKN and TP. All analyses were conducted according to Standard Methods (APHA 2005). The dewaterability of the sludges was assessed using the capillary suction time (CST) test and was measured using a GENEQ model 304B device. The concentrations of pathogen indicators (fecal coliforms and Escherichia coli (E. coli)) were assessed by method 9222D (APHA 2005). The concentrations of heavy metals (arsenic, cadmium, copper, lead, mercury, molybdenum, nickel, selenium and zinc) were analyzed to further assess the biosolids with respect to qualities that might impact on subsequent handling. The heavy metals were measured by ICP-MS (inductively coupled plasma atomic emission spectroscopy) using an Agilent ICP-MS 7500ce that was equipped with a CETAC ASX510 Autosampler.

The data collected from steady state of each phase was employed to evaluate the influence of addition of FeCl3 and SRT on the responses. The influence was assessed statistically using analysis of variance tests. In the subsequent discussion of results the significance of the statistical analysis is presented in brackets (i.e. p < xxx) whenever a statistical assessment was conducted to determine if a comparison was statistically significant.

RESULTS AND DISCUSSION

The properties of biosolids are affected by the characteristics of the fed sewage (Metcalf & Eddy Inc. 2003). Hence, the sewage employed in this study was characterized with respect to total and soluble COD, TSS, VSS, TP and TKN. The average (± standard deviations) concentrations of these responses are presented in Table 3. Due to the seasonal variability, the average concentrations of total COD in the sewage in Phases 3 and 4 were generally higher than in Phases 1 and 2. However, no significant difference was found between Phases 3 and 4 (p < 0.67). Other responses such as soluble COD, TSS, VSS, TKN and TP followed the trend of total COD in the four phases. In the subsequent discussion, the yields of TS and VS in mixed liquor were based on the volume of the sewage. Therefore, the seasonal variation in sewage strength was considered in the analysis of the results.

Table 3

Sewage characteristics

Phase 
Total COD (mg/L) 251 ± 59 304 ± 45 388 ± 65 383 ± 48 
Soluble COD (mg/L) 35 ± 14 35 ± 15 62 ± 20 52 ± 10 
TSS (mg/L) 118 ± 42 163 ± 77 208 ± 63 189 ± 29 
VSS (mg/L) 95 ± 23 154 ± 68 172 ± 61 166 ± 26 
TKN (mg/L) 27.4 ± 5.9 31.6 ± 11.3 44.4 ± 9.6 49.2 ± 10.3 
TP (mg/L) 3.6 ± 0.9 4.5 ± 0.9 5.2 ± 1.0 4.9 ± 1.2 
Phase 
Total COD (mg/L) 251 ± 59 304 ± 45 388 ± 65 383 ± 48 
Soluble COD (mg/L) 35 ± 14 35 ± 15 62 ± 20 52 ± 10 
TSS (mg/L) 118 ± 42 163 ± 77 208 ± 63 189 ± 29 
VSS (mg/L) 95 ± 23 154 ± 68 172 ± 61 166 ± 26 
TKN (mg/L) 27.4 ± 5.9 31.6 ± 11.3 44.4 ± 9.6 49.2 ± 10.3 
TP (mg/L) 3.6 ± 0.9 4.5 ± 0.9 5.2 ± 1.0 4.9 ± 1.2 

The production of biosolids in the AnMBR was evaluated in terms of the yields of VS and TS and was normalized on the basis of the volume of treated sewage (Figure 2). From Figure 2 it can be seen that the generation of VS in the AnMBR ranged between 30.0 and 90.8 g/m3 sewage and differed between the phases of the study. The values were lower than the typical VS yield that would be expected from a combination of primary and secondary aerobic treatment, which has been reported to range from 138 to 190 g/m3 sewage (Metcalf & Eddy Inc. 2003). Hence, it was concluded that direct treatment of municipal sewage in an AnMBR can produce approximately 25–50% of the VS that are generated by traditional wastewater treatment sequences.

Figure 2

Yields of TS and VS.

Figure 2

Yields of TS and VS.

As expressed in this study, the yield of VS in the AnMBR was expected to vary with the concentration of COD in the raw sewage, the SRT of the AnMBR and potentially the use of FeCl3 as a flux enhancer. As previously described in Table 3 the COD of the raw sewage differed substantially between the phases due to seasonal variability. To remove the impact of feed variability from the analysis, the VS yield was calculated on the basis of the mass loading of COD to the AnMBR (Table 4). From Table 4 it can be seen that normalization of VS yield on the basis of fed COD substantially reduced but did not eliminate the variability in the yields. Hence, the impacts of SRT and FeCl3 addition were further explored.

Table 4

VS yield based on COD loading

Phase 
VS yield (g VS/g CODfed0.12 ± 0.01 0.16 ± 0.04 0.18 ± 0.03 0.22 ± 0.01 
Phase 
VS yield (g VS/g CODfed0.12 ± 0.01 0.16 ± 0.04 0.18 ± 0.03 0.22 ± 0.01 

It was anticipated that increasing the SRT of the AnMBR would reduce the VS yield as extended SRTs would allow more time in the reactor for hydrolysis of the fed VS. Further, reduced yields of biomass would also result from increased endogenous decay of biomass at the extended SRTs. The impact of SRT on VS yields was assessed by examining the values estimated for Phases 2–4, which all had FeCl3 dosed into the wastewater. From Table 4 it can be seen that the average yield in Phase 4 was 22.2% (p < 1.7*10−5) and 37.5% (p < 4.3*10−7) higher than in Phase 3 and Phase 2 respectively as the SRT was increased from a value of 40 days (Phase 4) to 100 days (Phase 2). Hence, it was apparent that operation at extended SRTs could substantially reduce the production of VS in the AnMBR and this was consistent with increased hydrolysis and endogenous processes. This observation is consistent with a thorough COD mass balance analysis including COD destruction and biogas production that was presented by Dong et al. (in press).

It was hypothesized that the addition of FeCl3 to the AnMBR feed may increase the VS yield as it would act as a coagulant which would capture colloidal organic matter into floc. In this study, the impact of FeCl3 addition was assessed by comparing the yields observed in Phases 1 and 3, which both had the same SRT but were operated without and with FeCl3 addition respectively. From Table 4 it can be seen that the average yield of VS increased from 0.12 to 0.18 g VS/g CODfed upon FeCl3 addition (p < 5.8*10−8). Hence, the addition of FeCl3 substantially increased the production of VS in the AnMBR. This was attributed to the coagulation of colloidal and soluble organics into a particulate form (Dong et al. 2015).

The production of TS was deemed to be an important characteristic of biosolids production in AnMBRs as it reflects the quantity of solids that may require further handling. From Figure 2 it can be seen that the yield of TS ranged from 41.1 to 139.9 g/m3 and varied substantially between the phases of the study. Despite the variability, the TS yields in the current study were considerably less than the yields that have been reported from the combination of primary and secondary aerobic treatment (180–270 g/m3 sewage) (Metcalf & Eddy Inc. 2003). With the exception of Phase 4, the TS yields were lower or comparable to those reported for the extended aeration process without primary treatment (80–120 g/m3 sewage). It was concluded that treatment with the AnMBR could substantially reduce TS production when compared to traditional wastewater treatment when only the wastewater treatment train is considered.

As previously discussed, variability in wastewater properties, SRT and FeCl3 was found to impact on the yield of VS observed in the different phases of this study. It was expected that the first two factors would similarly impact upon TS yields and this was reflected by the 38% increase (p < 1.8*10−5) in TS yield that was observed as SRT was reduced from 70 days to 40 days in Phases 3-4 when the feed concentrations were similar. Hence SRT was found to substantially impact on TS yield.

The addition of FeCl3 as a flux enhancer was expected to have a greater impact on the yield of TS in the AnMBR than on VS production as it was expected to result in the production of inorganic precipitates. The impact of FeCl3 addition on TS production was assessed by comparing the yields in Phases 1 and 3, which had the same SRT but were operated without and with FeCl3 addition respectively (Figure 2). From Figure 2 it can be seen that the average TS yield in Phase 3 was 165% higher than Phase 1 (p < 2.6*10−22). The enhancement in TS yield was attributed to the higher sewage COD loading (Table 3) and the addition of FeCl3.The relative impact of FeCl3 addition on TS and VS production was assessed by examining the volatile fraction of the produced solids in Phases 1 and 2–4 (Figure 2). From Figure 2 it can be seen that the volatile fraction was approximately 92% in Phase 1 and decreased to 72–76% in Phases 2–4. Hence, it was apparent that the addition of FeCl3 resulted in the production of more inorganic solids than organic solids, thereby increasing the fixed fraction of the produced sludge.

Anaerobic digestion of the sludges produced in conventional wastewater treatment is often employed to reduce VS production. Hence, the previously referenced combination of yields of primary and secondary treatment sludges were combined with typical values for their anaerobic digestibility to provide a point of comparison with biosolids production from direct AnMBR treatment of wastewater. Assuming equivalent yields of primary and secondary VS and typical VS destruction efficiencies of 60% for primary sludges and 40% for secondary sludges (Jones et al. 2007), a combined system could be expected to generate 70–95 g VS/m3 wastewater. The VS production observed in this study was less than these typical values for three of the four scenarios tested.

VAR is a characteristic of biosolids treatment that needs to be satisfied if land application is considered as a reuse option. This property is reflective of the biological stability of the organic matter in the biosolids. In traditional digestion the requirements for VAR are considered to be satisfied if 38% of the VS mass is reduced in the treatment (US EPA 1999). In the current study, it was assumed that VAR requirements would be met if 38% of the VS loading in the feed wastewater to the AnMBR was destroyed. Hence, VS destruction was calculated on the basis of a mass balance of VS around the AnMBR. The VS mass flows decreased from 185 ± 55 to 31 ± 13 g/d in Phase 1, from 300 ± 104 to 75 ± 18 g/d in Phase 2, 326 ± 97 to 114 ± 29 g/d in Phase 3 and 317 ± 55 to 127 ± 17 g/d in Phase 4. Correspondingly, from Figure 3 it can be seen that the VS destruction through the AnMBR ranged from 60 to 83%. Hence it was concluded that the biosolids produced in the AnMBR would satisfy VAR requirements for land application.

Figure 3

VS mass reduction.

Figure 3

VS mass reduction.

The concentrations of TSS and VSS in the waste stream from the AnMBR (Figure 4) were characterized, as these values may influence decisions on downstream processing (storage, dewatering, etc.). From Figure 4 it can be seen that the mean concentration of TSS ranged from 5.8 g/L in Phase 1 to 17.5 g/L in Phase 3. The corresponding range of VSS concentrations was 5.5 g/L to 12.0 g/L. Hence, the concentrations of suspended solids in the waste stream were somewhat less than that which would typically be observed with anaerobic digestion of conventional primary and secondary sludge mixtures, which range from 1.5 to 4% TSS (Metcalf & Eddy Inc. 2003). It is therefore likely that thickening or dewatering of the waste stream may be required prior to offsite transport.

Figure 4

Concentrations of TSS and VSS in mixed liquor.

Figure 4

Concentrations of TSS and VSS in mixed liquor.

Figure 4 reveals substantial variations in the TSS and VSS concentrations in the mixed liquor between the different phases. It was anticipated that these concentrations would increase with SRT due to the accumulation of non-biodegradable components. Further, it was anticipated that addition of FeCl3 and increases in the feed sewage strength, would also increase the concentrations in the waste stream. In the current study the feed sewage concentrations differed between phases and hence the impact of SRT on the mixed liquor TSS and VSS concentrations could not be isolated in all phases. However, Phases 3 and 4 had similar feed concentrations and SRTs of 70 days and 40 days respectively. Figure 4 reveals that the TSS and VSS decreased by 21.1% (p < 1.7*10−6) and 17.7% (p < 2.1*10−5) respectively when SRT was reduced. Further, the impact of FeCl3 addition and feed concentration variation on mixed liquor concentrations could not be isolated in this study. A comparison of Phase 1 versus Phase 3 in Figure 4, where both of these varied while the SRT remained constant, reveals that the concentrations of TSS and VSS increased by 201% (p < 4.3*10−11) and 118% (p < 3.4*10−8). The greater increase in TSS relative to VSS was attributed to the formation of inorganic precipitates due to FeCl3 addition. Hence, the trends in the MLSS concentrations followed those that were anticipated based upon SRT, feed concentration and FeCl3 addition.

The presence of TKN and TP in biosolids is an important property when considering application to agricultural land as a soil amendment. In the current study the presence of these components was normalized on the basis of TS concentrations and Figure 5 shows that there was relatively little variation in their fractions between the study phases. Further, the concentrations of TKN and TP were comparable to those reported for anaerobically digested municipal biosolids, which were in the range of 0.5–17.6% and 0.5–14.3% for TKN and TP respectively (US EPA 1984). The AnMBR biosolids were deemed to be equivalent to anaerobically digested municipal biosolids in terms of TKN and TP content for land application purposes.

Figure 5

Concentrations of TKN and TP in biosolids.

Figure 5

Concentrations of TKN and TP in biosolids.

The dewaterability of the waste stream was considered to be important when designing downstream dewatering processes (e.g., centrifugation, belt filter press and filter press). In the current study the dewaterability of the waste stream was evaluated using CST test (Figure 6). From Figure 6 it can be seen that the CST values ranged from 345 s to 2,265 s through the various phases of the study. It was anticipated that the CST values would be affected by SRT and the addition of FeCl3. Figure 6 shows that the CST generally decreased from Phase 2 to Phase 4 as the SRT decreased and hence the suspended solids concentrations in the waste stream decreased. By contrast the CST values in Phase 3 were substantially less than those observed in Phase 1 despite the significantly higher TSS concentrations. The reduced CST values in Phase 3 were attributed to the addition of FeCl3 that would have coagulated and precipitated soluble and/or colloidal matter to form larger flocs. Overall it was concluded that reduced SRT and the addition of FeCl3 enhanced the dewaterability of the waste stream from the AnMBR.

Figure 6

CST values for AnMBR waste stream.

Figure 6

CST values for AnMBR waste stream.

To further assess the dewaterability of the mixed liquor from the AnMBR, the CST values in this study were compared with those reported for aerobic processes and anaerobically digested biosolids. The CST values in all phases in this study were significantly higher than the values reported for aerobic sludges (5–13 s) (Smollen 1986; Ge et al. 2011). In addition, the CST values in Phases 1–3 were generally higher than those reported for anaerobically digested biosolids (200–800 s) (Smollen 1986; Vesilind 1988; Krishnamurthy & Viraraghavan 2005). The results suggest that, while the AnMBR process operating conditions may be manipulated to enhance dewaterability, special attention may be required in the design of downstream dewatering processes.

It was anticipated that the operational conditions in Phase 4 may result in the least membrane fouling as the mixed liquor exhibited the highest dewaterability, thus these conditions may hold the most potential for full-scale implementation. Therefore, the pathogen indicator and heavy metal concentrations were tested in Phase 4 to evaluate the feasibility of biosolids for agricultural application. The potential presence of pathogens in municipal biosolids is an important property when considering land application of biosolids. Hence in this study the presence of fecal coliforms and E. coli. was assessed in Phase 4 to assess the feasibility of direct agricultural application of the biosolids (Table 5). From Table 5 it can be seen that the concentrations of fecal coliforms in the AnMBR biosolids were less than that reported for typical untreated sludges (1010 CFU/L)and were comparable to that reported for anaerobically digested biosolids (3*105–6*107 CFU/L) (US EPA 1979). However, when normalized on the basis of TS concentrations the AnMBR biosolids failed to meet Class B pathogen-reduction criteria (2*106 CFU/g TS) (US EPA 1999). The E. coli. concentrations (1.7–8.0*106 CFU/g TS) in the AnMBR waste stream also failed to meet reported standards for agricultural land application in Ontario (2*106 CFU/g TS) (Nutrient Management Act 2002). The results suggest that additional treatment of the biosolids would be required to satisfy requirements for direct land application of biosolids.

Table 5

Concentration of metals and pathogens in AnMBR biosolids

 As Cd Cu Hg Mo Ni Pb Se Zn Fecal coliforms E. coli 
Concentration in this study (mg/kg)a 8.4–10.2 1.5–1.9 310–320 0.65–0.91 35–37 73–76 28–30 2.8–3.4 739–843 2.2–9.8*106 CFU/g TS
or
4.4–16.6*107 CFU/L 
1.7–8.0*106 CFU/g TS
or
3.4–13.6*107 CFU/L 
Standards for land application(mg/kg)a 170 34 1,700 11 94 420 1,100 34 4,200 
 As Cd Cu Hg Mo Ni Pb Se Zn Fecal coliforms E. coli 
Concentration in this study (mg/kg)a 8.4–10.2 1.5–1.9 310–320 0.65–0.91 35–37 73–76 28–30 2.8–3.4 739–843 2.2–9.8*106 CFU/g TS
or
4.4–16.6*107 CFU/L 
1.7–8.0*106 CFU/g TS
or
3.4–13.6*107 CFU/L 
Standards for land application(mg/kg)a 170 34 1,700 11 94 420 1,100 34 4,200 

aDry weight basis.

The concentration of heavy metals (arsenic, cadmium, copper, lead, mercury, molybdenum, nickel, selenium and zinc) was evaluated to further assess the feasibility of biosolids for agricultural application. From Table 5 it can be seen that all the measured metal concentrations met standards for land application in Ontario (Nutrient Management Act 2002). It was concluded that heavy metal concentrations were not a concern for land application of biosolids from the AnMBR treating municipal wastewater.

CONCLUSIONS

The results obtained in this study provided insight into the biosolids characteristics of AnMBRs when operated over a range of SRTs and with/without addition of FeCl3 for treating municipal wastewater. The results indicated that the direct treatment of municipal sewage in an AnMBR produced less TS and VS than traditional wastewater treatment sequences. The yields of VS and TS were reduced with extended SRTs and increased with the addition of FeCl3. The biosolids produced in the AnMBR satisfied VAR requirements for land application and were deemed to be equivalent to anaerobically digested municipal biosolids in terms of TKN and TP content for land application purposes. The dewaterability of the waste stream, as measured by the CST test, was poorer than that reported for both aerobic and anaerobically digested sludges. The addition of FeCl3 and operation at reduced SRT improved dewaterability. The biosolids met Ontario standards for land application with respect to heavy metals, but needed additional treatment to satisfy pathogen requirements (fecal coliforms and E. coli).

ACKNOWLEDGEMENTS

This research project was financially supported by the National Sciences and Engineering Research Council (NSERC). The authors want to thank Dr Peter Seto from Environment Canada for providing space for experimental set-up and access to municipal wastewater in Wastewater Technology Centre and GE Analytical Lab for the assistance in sample analysis. Qirong Dong wants to thank the China Scholarship Council for the PhD Fellowship.

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