WEF Residuals and Biosolids Conference 2017

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1 The role of soluble P & EPS on dewatering performance Ester Rus *, Paul Fountain *, Nick Mills *, Achame Shana *, Obinna Molokwu **, and Manocher Asaadi **. * Thames Water, ** AD Technologies Corresponding author: Ester.rusperez@thameswater.co.uk Abstract Crossness and Beckton wastewater treatment plants have experienced issues with dewatering of their final digested sludge, requiring double the polymer demand for floc formation than that expected and achieving up to 25% less dry solids in the final cake. This paper looks into two possible root causes for the poor dewatering performance observed onsite: (1) the impact that high concentrations of soluble phosphorous on the monovalent to divalent ions on the sludge and its effects on dewaterability, and (2) the effects of erratic feeding of the digesters and the bacterial enhanced extra-polymeric substances excretion due to shock loads. Analysis and lab scale dewatering done on digested sludge samples from other 9 THP sites across the UK showed a clear relationship between soluble P and monovalent to divalent ratio, which also showed a relationship with polymer demand and final cake dry solids. Nevertheless, both Crossness and Beckton have seen an improvement in dewaterability with no change or increasing monovalent to divalent ratios, which goes against the observed trend. Furthermore, a set of lab scale digesters seeded with Crossness and Beckton digested sludge and fed with hydrolysed sludge from the two sites under controlled conditions (constant feeding and similar iron dosing) show an improvement in dewaterability with time compared to site. This suggests other factors may be driving this improvement. Extra-polymeric substances were considered a possible second factor having a negative impact on dewaterability. Both bound and soluble protein related extra-polymeric substances showed a weak relationship with polymer demand. Beckton site showed higher concentrations with higher polymer demand. This was not the case with Crossness. The results seem to show that both factors explored here play a role in digested sludge dewaterability. Nevertheless, other factors are also most likely having an impact on polymer demand and final cake dry solids and further work to explore other causes is recommended. Introduction Thames Water has built a number of Thermal Hydrolysis Plants (THP) to meet new capacity demands, improve efficiency and reduce operating costs. Two of these plants, Beckton & Crossness, located in East London utilise a two stream Cambi B12 THP system, 6 x 4,000m 3 digesters, 3 x 2MW CHP units and 5 x Bucher press dewatering units. The guarantee for the final stage dewatering cake quality was 38% dry solids (DS). Initially the plant dewatering achieved 38%DS with a polymer consumption of kg/tds. Nevertheless, after 2 hydraulic retention times (HRT) the cake DS dropped below 20% and the polymer consumption increased dramatically, averaging 27kg/TDS. This poor dewatering timed with the loss or washout of the seed sludge which had come from a different STW. During this period volatile solids destruction (VSD) and biogas 1007

2 production remained stable and as expected (Rus et al., 2015). The dewatering issue seemed to affect the rate of release of water as well as polymer demand and cake DS. The nature of the dewatering problems is very typical of dewatering issues seen on other Biological Nutrient Removal (BNR) sites. Beckton and Crossness were not designed to operate in BNR mode; nevertheless it was subsequently found that both sites were operating effectively as BNR. However, initial lab investigations suggested that the poor dewatering performance cannot be explained by the BNR \ sludge chemistry issue alone, and suggested the biology could be having a direct impact on dewaterability. Both approaches are reviewed in this paper. Work at lab scale: Some research has been done on the effect of soluble phosphate on dewatering of digested sludge, finding that high concentrations of soluble P in the digester could indeed lead to poor dewatering. Higgins and Novak (1997b) and Higgins et al., (2014) concluded that the monovalent to divalent (M/D) ion ratio was a good indicator of dewaterability, being positively correlated with the specific resistance of sludge to filtration, and therefore, cakes with higher DS could be achieved with sludges showing lower M/D ratios. This was also found by Alm et al., (2015) and Park (2002). Higgins et al., (2014) suggested that biological phosphorous is key to understanding the dewaterability issues in digested sludge. According to this research, high concentrations of soluble P in the digester can result in a decrease in available magnesium and calcium ions, which play an important role in bio-floc formation prior to digested sludge dewatering. Furthermore, it can cause the release of monovalent ions such as potassium which are known to be detrimental to dewatering. Since both Crossness and Beckon sites showed signs of opportunistic BNR, this theory was initially explored and an experiment with different concentrations of ferric chloride and sulphide to lab scale digesters was performed. The results showed a drop in sol P and a better dewaterability, both in terms of polymer demand and final cake solids for those lab digesters or chemostats with higher ferric chloride dose (Rus et al., 2015). Furthermore, it is common knowledge in the wastewater industry that anaerobic digesters perform better under stable conditions. That includes constant temperatures, feed flows, working volumes, loading rates, hydraulic retention times, etc. Data shows that both Crossness and Beckton were seeing a variability in their digester feed of more than 30% during a one given day. Identifying a method to quantify the effects of erratic/ inconsistent feeding or shock loads of a digester has been the real challenge and would be the first attenpt for THP+AD sludges. What is proposed in this paper is that under sustained inconsistent feeding of the digesters, bacteria would have to addapt to new, sub-optimal conditions. The biogas generation, methane production, and volatile solids reduction (VSR) may be close to what would be expected from a THP site, nonetheless, bacteria would be using a defence methanism against the unpredictable environment and excreted higher quantities of extracelular polymeric substances (EPS). EPS is a mix of different classes of organic macromolecules excreted by bacteria as part of their metabolism and to protect themselves from the environment (Aquino and Stuckey, 2004; Aquino and Stuckey, 2007; and Shana, 2015). The main macromolecules present in EPS are polysaccharides and proteins, but there are other polymeric compounds that could be found in the intracellular space or outside the cell surface. In literature, EPS is commonly classified as: Soluble EPS - soluble macromolecules, slimes (Sheng et al., 2010). Bound EPS sheaths, capsular polymers, condensed gels, etc. (Sheng et al., 2020). o Loosely bound EPS (LB-EPS): Outer layer with no well-defined edge. o Tightly bound EPS (TB EPS): Inner layer that has a structure and it is stable around the cell surface. The bound EPS has been identified as a gel like material that is able to hold water inside (4-5 kg water/kg bound EPS), which keeps the environment around the cell hydrated, crucial for survival. This also means that releasing water from this gel is difficult and therefore higher content of bound EPS may result in higher polymer demand for dewatering and lower cake dry solids (DS). 1008

3 Soluble EPS on the other hand can also have a negative impact on dewatering since it is negatively charged and can act as a sink for polymer during conditioning (Higgins et al., 2016). Higgins et al., (2016) found a strong correlation between Sol EPS and optimum polymer dose (R 2 = 0.88) for different AD sites. Both bound and soluble EPS are explored here. Methodology 1. Lab experiment set up: Six laboratory or bench scale semi-continuous anaerobic digestion (AD) rigs were set up. Each unit consists of a 10L spherical glass chemostat with 8L working volume (WV), with overhead stirrers for continuous mixing, aspirators with graduated biogas collection bottles, and a common water bath (see figure 1). Figure 1: (a) schematic of semi-continuous AD rig and (b) picture of chemostat in a water bath (Shana, 2015). Chemostats 1, 2, and 3 were seeded with digested sludge from Beckton and chemostats 4, 5, and 6 with digested sludge from Crossness. Ferric chloride was also dosed at different % by weight in the chemostats(see table 1): Table 1: Sludge procedence, OLR and Fe dose for each chemostat: Chemostat Site Beckton Beckton Beckton Crossness Crossness Crossness OLR (kgvs/m3/d) Fe dose (%w/w) before Nov Fe dose (%w/w) after Nov The dosing was changed in November 2015 after the initial trial (Rus et al., 2015). After November 2015, ferric chloride was added at 0.4% w/w for the dosed chemostats as that was considered the maximum dosing possible onsite. Chemostats 1, 3,4, and 6 where set at the organic loading rate (OLR) that would be expected onsite, and chemostats 2 and 5 where set up at the actual OLR onsite. Every day, hydrolysed sludge from both sites was imported to TW Innovation facilities to feed the chemostats. The feeding was done once a day. Volatile fatty acids (VFA), alkalinity, ph, ammonia, and gas production were checked 3 times a week. 1009

4 2. THP sites sampling: Digested sludge was imported in 1 L bottles from different THP sites in the UK: Crossness, Beckton, Oxford, Crawley, Chertsey, Riverside, Cardiff, Brand Sands, Cotton Valley, Afan and Seafield. Crossness and Beckton were sampled regularly to monitor both the dewaterability of the digested sludge at lab scale and in some occasions, digester stability. Chemical analysis was also done. 3. Chemical analysis: Chemical analysis was done by Thames Water accredited labs every 2 to 3 weeks on all chemostats. The analysis included DS, VS, soluble phosphorous, soluble magnesium, soluble calcium, soluble potassium, sulphate, and total iron. The same chemical analysis done for the chemostats was done for the site samples. 4. Dewatering of samples: Dewatering trials with a bench scale piston press and a lab scale Bucher sock were done every 2 to 3 weeks. In some occasions, the frequency of dewatering tests was higher (5 times a week) to observe the effect of reseeding on dewatering. The bench scale piston press is able to mimic a full scale belt press and the Bucher socks are able to give an estimate of full scale Bucher Press performance, remaining a couple of perceptual points below the actual full scale results. Polymer preparation: The chosen powder polymer (Flopan 4440 CT from SNF) was prepared with water and mixed with a magnetic stirrer for 45 minutes. Using a 500ml beaker, a 200ml sample of digested sludge was collected. The application of polymer to the digested sludge was done with a syringe, adding 5-10ml of polymer at a time and stirring the mix gently until stable flocs formed. Depending on the sludge quality more or less polymer was be added. It has been observed that over dosing polymer does not improve floc quality. Figure 2 shows a sample with a stable floc ready to be dewatered in the Bucher sock. Figure 2: Flocculated sludge ready to be dewatered. The flocculated sludge was then dewatered with two different methodologies to mimic the Bucher press and belt press technologies. (a) Bucher Press: the 200ml sample is poured into the Bucher socks with no previous decantation to replicate the Bucher press as closely as possible. The methodology used to dewater in the Bucher sock was provided by Bucher. Appropriate training was provided. Figure 2 shows an example of a Bucher sock: 1010

5 Figure 2: Lab scale Bucher Sock: (left) flocculated sludge added inside the sock, (right) redistribution of sludge in between twisting cycles to simulate full scale Bucher torsion. (b) Piston press: the 200ml sample flocculated sample is added into the piston chamber. The device (seen in figure 3) is then pressurised with an air compressor in 4 different steps: (1) 5 min under 2 barg, (2) 5 min under 5 barg, (3) 10 min under 7.5barg, (4) 10 min under 10barg, and (5) min under 11 barg. During that period water release is recorded every minute. Figure 3: Piston press. 5. Cation analysis and M/D Ratio: The calculation for the monovalent to divalent ratio was done using the methodology followed in Higgins and Novak (1997) and Higgins et al., (2014). Na +, K +, Mg 2+, and Ca 2+ where analysed using ion chromatography and the M/D ratio was calculated on an eq/eq basis: M/D Ratio (eq/eq) = Na + + K + Mg 2+ + Ca

6 The ammonium ion was not included at this point, since most of the digesters have similar concentrations of 2.8-3gNH 4 -N/L. 6. EPS and SMP: The soluble fraction of EPS is obtained by centrifuging the sample for 15 minutes at 3000 rpm at 5 C. The supernatant is then double filtered through 1.5µm filters. To measure the bound part of EPS, distilled water is added to the pellet after centrifugation and the suspension is transferred into a flask. This is heated for an hour at 80 C and left to cool. The suspension is then centrifuged at 3000rpm for 30 minutes and the supernatant is subjected to double filtration using a 1.5 µm filters. Results and discussion 1. Stability of chemostats, specific gas production and volatile solid reduction: The average and standard deviation (SD) for alkalinity, VFA, ph, NH 4 -N, methane, specific gas production (SGP), and VSR across all six chemostats are shown in table 2. Table 2: Average and SD for alkalinity, VFA, ph, NH 4 -N, methane, SGP, and VSR. Chemostat Alkalinity VFA NH ph (-) 4 -N SGP VSR - CH (mg/l) (mg/l) (mg/l) 4 (%) (m 3 /TDS) VK (%) 1 - Beckton 7943± ±73 8± ±476 64±9 418±40 56± Beckton 7943± ±96 8± ±420 63±5 439±54 57±0.1* 3 - Beckton 7800± ±127 8± ±367 65±5 408±39 58±0.1* 4- Crossness 8188± ± ± ±317 65±4 414±47 55± Crossness 7600± ± ± ±342 62±5 403±43 56±0.1* 6- Crossness 7615± ± ± ±309 63±5 408±56 55±0.1* *Corrected for iron addition. Some variability on the alkalinity and VFA values was observed due to instrumentation issues. Nevertheless, both parameters remained stable for most of the recorded period. Ammonium levels stayed below 3,000mg/L and the methane content in the biogas remained between 62-65% for all chemostats. The SGP and VSR with the Van Kleeck method were very similar to those observed onsite. Therefore the overall performance of AD seen at lab scale was comparable to that seen at full scale. In terms of gas production and VSR, what is observed at both sites (both at lab and full scale) is what would be expected from a THP +AD site. 2. Dewaterability of lab scale and full scale: Figure 4 shows that eventhought the overall performance of the digesters onsite in terms of gas yield and VSR is similar to that observed at lab scale, the dewaterability is very different, particulartly in the polymer demand. 1012

7 Figure 4: Evolution of full scale dewaterability compared to lab scale. All samples dewatered with a lab Bucher sock. 3. Soluble Phosphate and the monovalent to divalent ratio: Figure 5 shows the relationship between the M/D ratio done with four main ions: Mg 2+, Ca 2+, Na +, and K

8 Figure 5: Relationship between sol-p and M/D Ratio (no ammonium) for all THP sites. A strong correlation was found between both, with an R 2 of 0.7. Looking at the relationship between soluble P and the cations involved in the M/D ratio (figure 6 (a) and (b)), a downward trend was observed for the two divalent ions (Mg 2+ and Ca 2+ ). Magnesium seems to show a stronger relationship than calcium. Figure 6: Relationship between (a) magnesium ion, (b) calcium ion, (c) potassium ion, and (d) sodium ion and soluble P. 1014

9 According to the divalent ion bridging theory, high concentrations of soluble P are accompanied by a decrease in both magnesium and calcium ions, which are associated with bio-floc formation (Higgins and Novak, 1997). Phosphorous seems to be able to form complexes and precipitates with these two cations as observed by Higgins et al., (2014). These results seem to indicate that there is a better linear fit with magnesium. In terms of soluble calcium, only Oxford sludge seemed to not follow the linear trend, with very low levels of soluble P (between 4-10mg/L) and corresponding soluble calcium concentrations between 30-50mg/L. Figure 6 (c) and (d) show the relationship of the two monovalent ions potassium and sodium. Potassium follows an upward trend with increasing soluble P which was also found by Higgins et al., The relationship of soluble sodium with soluble P is not clear at the observed concentrations. Data seems to show slightly higher concentrations of soluble Na after soluble P reached 200mg/L. Figure 7 shows the results on the M/D ratio of the listed sites and their polymer demand during the dewatering trials in the lab. All sites depicted are THP-AD sites. Seafield is clearly driving the trend line with most of the sites remaining at an M/D ratio of This may mean that even for these sites, other factors aside from the soluble P concentrations may be influencing the polymer demand for floc formation, including shock loads or sudden changes in temperature, in the digester or right before dewatering. It may also mean that our methodology (manual polymer injection and mixing + lab scale Bucher sock test) is not sensitive enough to give us significant differences across sites with similar dewatering performance. Figure 7: Relationship between polymer demand (kg/tds) and M/D ratio (All sites apart from Crossness and Beckton). 1015

10 Figure 8: Relationship between cake DS (%) and M/D ratio (All sites apart from Crossness and Beckton) Figure 8 shows the final DS of the digested sludge from the different sites and their M/D Ratio. The scatter is significant with an R 2 of 0.48 indicating a high percentage of variation in the data. This could be due to other variables affecting the dewaterability of the sample. For instance, Cardiff did not achieve more than 40% DS even though a high M/D ratio is observed. This may be related to the surplus activated sludge (SAS) content being very high (85%) compared to the other sites plotted with a similar M/D ratio. In the case of Oxford, the samples corresponding to a final cake DS of circa 40% where taken during a period where the site was feeding very inconsistently, with constant start and stops which would shock load the digesters. 1016

11 (a) Beckton M/D Ratio and final cake DS (%) (b) Crossness M/D ratio and Final cake DS (%) Figure 9: M/D ratio and final cake DS (%) for (a) Beckton and (b) Crossness compared to other THP sites (shown separately on figure x). Figure 9 shows the evolution of the M/D ratio and final cake DS compared to the rest of the THP sites. Where most sites achieved DS between 35 to 46%, both Crossness and Beckton struggled to achieve 30%DS back in 1017

12 August At that time both sites were seeing extreme erratic feeding, with more than 60% variability in the feed observed in the 15 min interval data collected. In terms of soluble P, only Beckton had a significant concentration of 149mg/L whereas Crossness only had 56mg/L. The higher M/D ratio for Beckton is therefore expected according to the divalent ion bridging theory (Higgins et al., 2014). This difference in sol P values across sites could be due to the higher amount of SAS treated at Beckton (80% SAS compared to 40% in Crossness). In late November 2015, soluble P levels in Beckton went up to 247mg/l which corresponds to the increase in the M/D ratio. Nonetheless, we see higher final cake DS. Looking at the feeding regimes for that month, the feed flows into the digesters seem to be more stable, with a variability of 40%. This improvement may partly explain the better dewaterability obtained even though the sol P increased. Other improvements onsite may have contributed to it. Crossness saw a similar improvement in DS with not much change in sol P in the month of November. In late January 2016, both sites continued to follow an upward trend in final cake DS observed at lab scale. At that point, Beckton had much more work done around the feeding system and therefore the variability of digester feed was lower than at Crossness. This may explain why even though Crossness only had 102mg/l of Sol P compared to the 246mg/l seen at Beckton, the dewatering was very similar at both sites. The Beckton chemostats showed a significantly lower M/D ratio than that observed onsite. This also correlates to the lower soluble P seen in the chemostats (149 mg/l compared to 250mg/l seen onsite). The possibility of some mechanism changing the chemistry in the full scale digesters compared to the chemostats is being studied. An initial theory is the possibility of higher amounts of struvite formation at full scale. Magnesium ion seemed to be slightly lower for sites samples (3.4mg/l compared to 9mg/l in the chemostats). Nevertheless, whether or not this difference is significant is to be determined. The concentration of calcium ions was found to be much lower onsite as well, with values close to 40mg/l compared to 70mg/l in the chemostats. In terms of dewatering, from November to January there was a slight improvement that could be associated with a long stable period with optimum AD conditions. The Crossness chemostats showed a similar M/D ratio to that seen onsite. The DS obtained where in all cases slightly higher than those seen for site samples in the lab scale dewatering trials also showing an improvement with time. 1018

13 (a) Beckton M/D Ratio and polymer demand (kg/tds) (a) Crossness M/D ratio and polymer demand (kg/tds) Figure 10: M/D ratio and polymer demand (kg/tds) for (a) Beckton and (b) Crossness compared to other sites (shown separately on figure x). 1019

14 Figure 10 shows the relationship of the M/D ratio with the polymer demand of Crossness and Beckton site and chemostats compared to other THP sites. Beckton site digested samples had a significantly higher polymer demand compared to the chemostats. The M/D ratio observed onsite was higher than in the chemostats. Crossness sees a similar difference in polymer demand between the chemostats and site, but in this case the M/D ratios are very similar, staying circa Effects of reseeding chemostats 3 (Beckton) and 6 (Crossness): Two chemostats (Chemostat 3 from Beckton and Chemostat 6 from Crossness) were reseeded and dewatering tests with the piston press where done twice a week for a period of 4 weeks. Both DS and polymer dose was recorded. The results can be seen in figure 11: Figure 11: Reseeding of chemostats (a) 3 with digested sludge from Beckton STW and (b) 6 with digested sludge from Crossness STW and monitoring of dewaterability and polymer demand. The iron dose remained constant at 0.2% (w/w) to match what was then dosed onsite. Therefore, under same iron conditions, a clear improvement in dewatering as well as a gradual decrease in polymer demand was observed in both cases. This would indicate that something else apart from the iron addition was affecting the dewatering performance. 1020

15 5. Extra-Polymeric Substances: Figure 12 shows the relationship between polymer demand (kg/tds) and soluble protein related EPS (mg/l) for all the THP sites sampled including Crossness and Beckton. Chemostats 2 (Beckton) and 5 (Crossness) are also plotted for being the most representative of both sites. There seems to be a relationship between these two parameters (R 2 = ). Some scatter may be expected since the presence of soluble P seems to be another variable affecting dewatering. Figure 12: Relationship between soluble protein related EPS (mg/l) and polymer demand (kg/tds). In terms of the relationship between chemostats and sites, only Beckton followed the general trend of higher polymer demand with higher soluble EPS. The higher EPS onsite may be related to the more erratic or irregular feeding regime observed there compared to the consistent feeding regime of the chemostats. This is not observed for Crossness, where the EPS is similar in both the chemostat and full scale. The variance in sludge feeding flows into the digesters for a given day is higher at Crossness and therefore higher EPS was expected at full scale. 1021

16 Figure 13: Relationship between bound protein related EPS (mg/l) and polymer demand (kg/tds). Figure 13 shows the relationship between the bound EPS and polymer demand of THP sites and chemostats 2 (Beckton) and 5 (Crossness). The relationship is weaker for this parameter, with an R 2 = Beckton chemostat 2 shows a lower concentration in bound EPS and polymer demand before dewatering compared to site as observed in figure 12 for soluble EPS and polymer demand. Nevertheless, Crossness does not follow the linear trend, with higher polymer demand onsite compared to chemostat 5 even though the bound EPS is similar in both cases. This is also the case for soluble EPS and polymer demand depicted in figure 12. It is worth noting that there is currently no standard methodology to follow to be able to compare our own findings with other papers. Other variables that have not been assessed in this paper may also be having an impact on polymer demand and final cake DS which would increase the scatter for both the EPS and M/D ratio relationships with polymer demand. Conclusions Poor digested sludge dewaterability was observed at Crossness and Beckton sites with no other signs of digester underperformance. Two possible causes for poor dewaterability were identified for Crossness and Beckton sewage treatment plants. One theory proposed by Higgins et al., (2014) suggest that soluble P is an indirect driver for dewaterability issues, both in terms of polymer demand for flocculation and final cake DS. Results from 9 THP sites showed a strong correlation between soluble P and the M/D ratio. This ratio also correlated to dewaterability, particularly for polymer demand. Plants with higher M/D ratios showed poorer sludge dewaterability, confirming previous work (Higgins and Novak, 1997b; and Higgins et al., 2014). Nevertheless, both Beckton and Crossness saw an 1022

17 improvement in dewaterability with increasing final cake DS and decreasing polymer demand with time even though the M/D ratio did not change for Crossness and saw an increase for Beckton. This may indicate some other factor is influencing dewaterability. Additionally, chemostats 3 and 6 were re-seeded with digested sludge from Crossness and Beckton respectively to observe the evolution from full scale to lab scale of dewaterability with time. The results for the first 30 days after seeding showed a steady improvement of dewatering with time with no change in M/D ratio. This also suggested that soluble P may not be only factor having an impact on dewaterability. The second factor considered was EPS. According to the data from both Crossness and Beckton, the digesters from both sites were being fed erratically. The hypothesis proposed in this paper suggests that the shock loads observed at both sites could trigger a defence mechanism in the bacterial population in the form of higher EPS excretion. Since EPS is known to be difficult to dewater, higher EPS concentrations would likely translate into higher polymer demand for bio-floc formation. Initial results showed a weak relationship between the protein related EPS (both soluble and bound parts) and polymer demand which indicates EPS is likely partly driving digested sludge dewaterability. When comparing chemostats to site, Beckton showed higher soluble and bound protein related EPS onsite compared to the chemostat. This was expected given the better feeding conditions provided at lab scale compared to that at full scale. Nevertheless, this result was not found for Crossness, with both bound and soluble EPS figures remaining very similar at lab and full scale. This may suggest some other factor driving the poor dewatering at this particular site. Challenges around the methodology for EPS extraction were found since there is no one particular standard method to extract the polymeric substance across the business. This means no comparison can be done with other research. Therefore the next steps to take would be expand our understanding on EPS extraction methodologies by comparing results with different methods available in the literature. Furthermore, work around the observation of EPS under the microscope is already in place with encouraging initial results. Other factors that may be having an impact on dewaterability will also be looked into. Some preliminary work on particle size distribution suggests that site and chemostats may show differences in particle sizes. Sludge viscosity and bacterial population are two other parameters to investigate. References: Alm, R., Sealock, A.W., Koo, A., and Sprouse, G. (2015) Investigations into improving dewaterability at a bio- P/anaerobic digestion plant. Residuals and Biosolids Conference 2015, Washington DC, US. Aquino, S.F. and Stuckey, D.C. (2004) The effect of organic and hydraulic shock loads on the production of soluble microbial products (SMP) in anaerobic digesters. Water Environ. Res. 76 (6), Aquino, S.F., and Stuckey, D.C. (2007) Integrated model of the production of soluble microbial products (SMP) and extracellular polymeric substances (EPS) in anaerobic chemostats during transient conditions. Biochemical Engineering Journal. 38, Higgins, M. and Novak, J. (1997) Characterization of Exocellular Protein and Its Role in Bioflocculation. J. Environ. Eng. 123:5(479), Higgins, M., Bott, C., Schauer, P., and Beightol, S. (2014) Does bio-p impact dewatering after anaerobic digestion? Yes, and not in a good way! WEF Residuals and Biosolids Conference 2014, Austin TX. US. Higgins, M., Murthy, S., Bott, C., and Rajagopalan, G. (2016) Bioflocculation Mechanisms and Implications for Understanding the Effects of Co-Digestion and BioP on Dewatering of Anaerobically Digested Biosolids. WEF Residuals and Biosolids Conference, Milwaukee. US. 1023

18 Park, C. (2002) Cations and activated floc structure. Master Thesis. Virginia Polytechnic Institute and State University. Rus, E., Perrault, A., Fountain, P., and Shana, A. (2015) East London THP commissioning and laboratory investigations. European Biosolids Conference Manchester, UK. Shana, A. (2015) Application of an innovative process for improving mesophilic anaerobic digestion of sewage sludge. PhD Thesis. University of Surrey. Sheng, G-P., Yu, H-Q., and Li, X-Y. (2010) Extracellular polymeric substances (EPS) of microbial aggregates in biological wastewater treatment systems: A review. Biotechnology advances. 28 (6),