Applied Science, Faculty of
Engineering, School of (Okanagan)
DSpace
UBCO
Basar, Ibrahim Alper
Doctor of Philosophy - PhD
University of British Columbia
In recent decades, resource recovery and waste reduction have become the prominent focus to adapt urban technologies to the circular economy concept. From this point of view, municipal wastewater has been considered a resource rather than a waste stream. A thermochemical conversion process known as hydrothermal liquefaction (HTL) has emerged as a promising candidate for the valorization of municipal sludge, owing to its high energy recovery efficiency and potential to produce value-added products. The HTL process converts organic feedstock into a liquid biofuel called bio-crude via a similar process to the natural occurrence of petroleum by using near-critical water temperatures (250-374°C) and pressures (52-220 bars). The HTL process alone has the potential to reach over 70% energy recovery from municipal sludge in the form of bio-crude, and over 80% of by-products of HTL aqueous phase and hydrochar can be valorized. While the utilization of the HTL process seems promising, there are several challenges that need to be addressed to enable its seamless integration into wastewater treatment plants (WWTP). The most critical challenge is the treatment or valorization of HTL process wastewater, named HTL aqueous phase. In this thesis, regression analysis was used to mathematically establish relationships between the HTL aqueous characteristics and the HTL reaction conditions. HTL aqueous obtained at temperatures ranging from 290 to 360°C and retention times between 0 to 30 minutes was found to contain chemical oxygen demand (COD) levels ranging from 87 to 103 g/L, total ammonia levels of 3.9 to 7.1 g/L, and total phenolics levels of 1.5 to 2.3 g gallic acid equivalent/L. The ultimate biochemical oxygen demand and biochemical methane potential values showed a positive correlation with the secondary sludge ratio in mixed sludge. Among the on-site biological treatment options, both aerobic and mesophilic anaerobic processes were identified as potential choices. In a continuous-flow aerobic treatment system, up to 79.2% of COD from HTL aqueous was possible to degrade. These findings suggest that on-site aerobic treatment is a feasible option for treating HTL aqueous in WWTPs, but a tertiary treatment, like advanced oxidation or adsorption, is necessary to eliminate the recalcitrant COD.
EVALUATION OF ON-SITE BIOLOGICAL TREATMENT OPTIONS FOR HYDROTHERMAL LIQUEFACTION AQUEOUS PHASE DERIVED FROM SLUDGE IN MUNICIPAL WASTEWATER TREATMENT PLANTS by Ibrahim Alper Basar M.Sc., Akdeniz University, 2018 B.Sc., Akdeniz University, 2012 A DISSERTATION SUBMITTED IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF DOCTOR OF PHILOSOPHY in THE COLLEGE OF GRADUATE STUDIES (Civil Engineering) THE UNIVERSITY OF BRITISH COLUMBIA (Okanagan) April 2024 © Ibrahim Alper Basar, 2024 ii The following individuals certify that they have read, and recommend to the College of Graduate Studies for acceptance, a dissertation entitled: EVALUATION OF ON-SITE BIOLOGICAL TREATMENT OPTIONS FOR HYDROTHERMAL LIQUEFACTION AQUEOUS PHASE DERIVED FROM SLUDGE IN MUNICIPAL WASTEWATER TREATMENT PLANTS submitted by Ibrahim Alper Basar in partial fulfillment of the requirements of the degree of Doctor of Philosophy. Dr. Cigdem Eskicioglu, School of Engineering Supervisor Dr. Abbas Milani, School of Engineering Supervisory Committee Member Dr. Nicolas Peleato, School of Engineering Supervisory Committee Member Dr. Sina Kheirkhah, School of Engineering University Examiner Dr. Yang Liu, Civil and Environmental Engineering, University of Alberta External Examiner iii Abstract In recent decades, resource recovery and waste reduction have become the prominent focus to adapt urban technologies to the circular economy concept. From this point of view, municipal wastewater has been considered a resource rather than a waste stream. A thermochemical conversion process known as hydrothermal liquefaction (HTL) has emerged as a promising candidate for the valorization of municipal sludge, owing to its high energy recovery efficiency and potential to produce value-added products. The HTL process converts organic feedstock into a liquid biofuel called bio-crude via a similar process to the natural occurrence of petroleum by using near-critical water temperatures (250-374°C) and pressures (52-220 bars). The HTL process alone has the potential to reach over 70% energy recovery from municipal sludge in the form of bio-crude, and over 80% of by-products of HTL aqueous phase and hydrochar can be valorized. While the utilization of the HTL process seems promising, there are several challenges that need to be addressed to enable its seamless integration into wastewater treatment plants (WWTP). The most critical challenge is the treatment or valorization of HTL process wastewater, named HTL aqueous phase. In this dissertation, regression analysis was used to mathematically establish relationships between the HTL aqueous characteristics and the HTL reaction conditions. HTL aqueous obtained at temperatures ranging from 290 to 360°C and retention times between 0 to 30 minutes was found to contain chemical oxygen demand (COD) levels ranging from 87 to 103 g/L, total ammonia levels of 3.9 to 7.1 g/L, and total phenolics levels of 1.5 to 2.3 g gallic acid equivalent/L. The ultimate biochemical oxygen demand and biochemical methane potential values showed a positive correlation with the secondary sludge ratio in mixed sludge. Among the on-site biological treatment options, both aerobic and mesophilic anaerobic processes were identified as potential choices. In a continuous-flow aerobic treatment system, up to 79.2% of COD from HTL aqueous was possible to degrade. These findings suggest that on-site aerobic treatment is a feasible option for treating HTL aqueous in WWTPs, but a tertiary treatment, like advanced oxidation or adsorption, is necessary to eliminate the recalcitrant COD. iv Lay summary Hydrothermal liquefaction is a novel method that converts organic material into a petroleum like biofuel called bio-crude. Implementing this technology in wastewater treatment plants has the potential to double the energy recovery from wastewater solids compared to our current technology. Along with bio-crude, hydrothermal liquefaction also produces a liquid waste stream that is rich in organic matter. This dissertation explores the on-site treatment of the liquid stream generated during hydrothermal liquefaction in wastewater treatment plants. Through experimental investigations the study examines the feasibility of hydrothermal liquefaction aqueous phase treatment in various bioreactor environments employing different beneficial microbial cultures. Despite facing challenges such as toxic compounds, the results suggest promising avenues for sustainable waste treatment and resource recovery. The dissertation emphasizes how integrating hydrothermal liquefaction byproduct treatment can contribute to the overall sustainability of wastewater treatment plants by harnessing energy and nutrients from waste streams. v Preface The development, design, research execution, analysis, and preparation of this dissertation were primarily carried out by the dissertation author, with guidance and input provided by the supervisor, Dr. Cigdem Eskicioglu. Assistance and feedback for the literature review section were provided by the collaborators (Dr. Hélène Carrère and Dr. Eric Trably). The majority of the laboratory analyses were carried out by the author in the Bioreactor Technology Group (BTG) laboratory. Support was received from another Ph.D. student, Huan Liu, for municipal sludge dewatering, conducting hydrothermal liquefaction runs, and sample characterization. Abigail Stokes, an undergraduate student in mechanical engineering, assisted with the biochemical methane potential tests in Chapter 4. Organic compound quantification by GC-MS analysis was performed by BTG technician Jacky Takeuchi. This research has led to the publication of three peer-reviewed journal articles, with two additional articles currently under review: â¢ Parts of Chapter 1 and 2 were published as a review article: Basar, I.A., Liu, H., Carrere, H., Trably, E., Eskicioglu, C., 2021. A review on key design and operational parameters to optimize and develop hydrothermal liquefaction of biomass for biorefinery applications. Green Chemistry 23, 1404â1446. https://doi.org/10.1039/D0GC04092D. â¢ Part 3 of the research was published as the journal article: Basar, I.A., Liu, H., Eskicioglu, C., 2023. Incorporating hydrothermal liquefaction into wastewater treatment â Part III: Aqueous phase characterization and evaluation of on-site treatment. Chemical Engineering Journal 467, 143422. https://doi.org/10.1016/j.cej.2023.143422. â¢ Part 4 of the research was published as the journal article: Basar, I.A., Stokes, A., Eskicioglu, C., 2024. Evaluation of on-site biological treatment options for hydrothermal liquefaction aqueous phase derived from sludge in municipal wastewater treatment plants. Water Research 121206. https://doi.org/10.1016/j.watres.2024.121206. â¢ Part 5 of the research was submitted to Bioresource Technology as a journal article: Basar, I.A., Liu, H., Eskicioglu, C. Effects of municipal sludge composition on hydrothermal liquefaction products: Aqueous phase characterization and biodegradability assessment. Bioresource Technology 130671. https://doi.org/10.1016/j.biortech.2024.130671. â¢ Part 6 of the research was submitted to Water Research as a journal article: Basar, I.A. and Eskicioglu, C. Continuous-flow aerobic treatment of municipal sludge derived hydrothermal liquefaction aqueous phase in wastewater treatment plants. vi In addition to the articles mentioned, out of the scope of this dissertation, one journal article and one conference article in the area of hydrothermal liquefaction of municipal sludge were collaboratively published with Huan Liu as co-equal authors: â¢ Liu, H., Basar, I.A., Eskicioglu, C., 2023. Hydrothermal liquefaction for sludge-to-energy conversion: An evaluation of biocrude production and management of waste streams. Energy 281, 128268. https://doi.org/10.1016/j.energy.2023.128268. â¢ Liu, H., Basar, I., Lyczko, N., Eskicioglu, C., Liu, H., 2023. Mining Energy from Wastewater: Coupling Hydrothermal Liquefaction and Anaerobic Digestion for Municipal Sludge Treatment, in: Proceedings of the Water Environment Federation. WEFTEC. https://doi.org/10.2175/193864718825159036. Additionally, outside the scope of this dissertation, I collaborated with Huan Liu on a journal article focusing on optimizing the HTL process for bio-crude production: â¢ Liu, H., Basar, I.A., Lyczko, N., Nzihou, A., Eskicioglu, C., 2022. Incorporating hydrothermal liquefaction into wastewater treatment â Part I: Process optimization for energy recovery and evaluation of product distribution. Chemical Engineering Journal 449, 137838. https://doi.org/10.1016/j.cej.2022.137838. Findings from this study have been presented at three international conferences and have been submitted for presentation at two additional upcoming conferences: â¢ Basar, I.A., Liu, H., Eskicioglu, C., Kadota, P. (2021). Reinventing the solids train by incorporating hydrothermal liquefaction into a wastewater treatment plant, 4th International Water Association (IWA) Resource Recovery Conference, Istanbul/Turkiye. â¢ Basar, I.A., Liu, H., Eskicioglu, C. (2023). Municipal sludge valorization via hydrothermal liquefaction: On-site treatment of process wastewater, 2nd International Water Association (IWA) Canadian Young Water Professionals Conference, Vancouver/British Columbia/Canada. â¢ *Liu, H., *Basar, I.A., Eskicioglu, C. (2023). Mining Energy from Wastewater: Coupling Hydrothermal Liquefaction and Anaerobic Digestion for Municipal Sludge Treatment, WEFTEC, Chicago/Illinois/USA (*Equally contributed authors). â¢ Basar, I.A., Liu, H., Eskicioglu, C. (2024). Hydrothermal liquefaction: a promising solution for sustainable municipal sludge management, BCWWA 2024 Annual Conference & Trade Show, Whistler/British Columbia/Canada. vii â¢ Basar, I.A., Liu, H., Blair, D., Eskicioglu, C. (2024). Characterization and biodegradability assessment of hydrothermal liquefaction aqueous phase derived from different municipal sludge streams, WEFTEC, New Orleans/Louisiana/USA. viii Table of contents Abstract...................................................................................................................................... iii Lay summary ............................................................................................................................. iv Preface ....................................................................................................................................... v Table of contents ...................................................................................................................... viii List of tables .............................................................................................................................. xii List of figures ........................................................................................................................... xiii Acknowledgements .................................................................................................................. xix Dedication ................................................................................................................................. xx Chapter 1: Introduction ............................................................................................................... 1 1.1 Background ......................................................................................................... 1 1.1.1 Existing sludge treatment techniques .................................................................. 1 1.1.2 Hydrothermal liquefaction .................................................................................... 2 1.1.3 Hydrothermal liquefaction technology to date ...................................................... 3 1.2 Objectives and hypotheses ................................................................................. 4 1.3 Novelty ................................................................................................................ 5 1.4 Structure of the dissertation ................................................................................. 6 Chapter 2: Literature Review ...................................................................................................... 8 2.1 Hydrothermal liquefaction chemistry .................................................................... 8 2.1.1 Depolymerization ...............................................................................................10 2.1.2 Decomposition ................................................................................................... 11 2.1.3 Recombination ...................................................................................................14 2.2 Hydrothermal liquefaction products ....................................................................16 2.2.1 Bio-crude ............................................................................................................16 2.2.2 Hydrochar ..........................................................................................................19 2.2.3 Hydrothermal liquefaction gaseous phase ..........................................................20 2.2.4 HTL aqueous ......................................................................................................21 2.3 Impact of hydrothermal liquefaction conditions and feedstock on products .........26 2.4 Use of catalyst in the hydrothermal liquefaction process ....................................30 2.5 Techno-economic analysis of the hydrothermal liquefaction process ..................32 2.6 Summary of findings from the literature review ...................................................34 Chapter 3: Characterization of Municipal Sludge-Derived Hydrothermal Liquefaction Aqueous Phase Under Various Reaction Conditions ................................................................................36 3.1 Introduction ........................................................................................................36 3.2 Materials and methods .......................................................................................37 ix 3.2.1 Sample preparation ............................................................................................37 3.2.2 Experimental design ...........................................................................................37 3.2.3 Bench-scale HTL operation ................................................................................38 3.2.4 Characterization of hydrothermal liquefaction aqueous stream...........................40 3.3 Results and discussion .......................................................................................45 3.3.1 Chemical oxygen demand (COD) .......................................................................45 3.3.2 Dissolved proteins and total ammonia ................................................................47 3.3.3 Volatile fatty acids (VFAs) ...................................................................................48 3.3.4 Total alkalinity and pH ........................................................................................50 3.3.5 Phosphorus ........................................................................................................51 3.3.6 Total carbohydrates and total phenolic compounds ............................................52 3.3.7 Furanic compounds, levulinic acid, ethanol, and glycerol ...................................53 3.3.8 Gas chromatography-mass spectrometry results ...............................................54 3.3.9 Ultimate analysis and inorganic elements ...........................................................58 3.3.10 ANOVA test and regression analysis ..................................................................59 3.3.11 Evaluation of on-site biological treatment options for HTL aqueous in a WWTP .63 3.4 Summary ............................................................................................................66 Chapter 4: Analyzing four biological treatment methods for treating hydrothermal liquefaction aqueous phase derived from sludge in municipal wastewater treatment plants .........................68 4.1 Introduction ........................................................................................................68 4.2 Materials and methods .......................................................................................68 4.2.1 Experimental design and sample preparation .....................................................68 4.2.2 Sample and inoculum characterization ...............................................................70 4.2.3 Aerobic biodegradability test (Scenario I) ...........................................................70 4.2.4 Anaerobic biodegradability test (Scenario II and III)............................................72 4.2.5 Dark fermentation test (Scenario IV) ..................................................................74 4.2.6 Data analysis ......................................................................................................75 4.3 Results and discussion .......................................................................................76 4.3.1 Characterization of HTL aqueous samples .........................................................76 4.3.2 Scenario I â Aerobic treatment ...........................................................................77 4.3.3 Scenario II â Mesophilic anaerobic treatment .....................................................80 4.3.4 Scenario III â Thermophilic anaerobic treatment ................................................86 4.3.5 Scenario IV â Dark fermentation .........................................................................89 4.3.6 ANOVA test and regression analysis ..................................................................90 4.3.7 Advantages and disadvantages of treatment options .........................................93 x 4.4 Summary ............................................................................................................93 Chapter 5: Characterization and biodegradability assessment of hydrothermal liquefaction aqueous phase derived from various sludge streams of different treatment plants ....................95 5.1 Introduction ........................................................................................................95 5.2 Materials and methods .......................................................................................95 5.2.1 HTL feedstock sample preparation and characterization ....................................95 5.2.2 Bench-scale HTL operation ................................................................................96 5.2.3 Chemical characterization of HTL aqueous ........................................................96 5.2.4 Aerobic biodegradability tests .............................................................................97 5.2.5 Anaerobic biodegradability tests .........................................................................97 5.2.6 Data analysis ......................................................................................................97 5.3 Results and discussion .......................................................................................98 5.3.1 Sludge characterization ......................................................................................98 5.3.2 Chemical characterization of HTL aqueous ........................................................99 5.3.3 Biochemical oxygen demand ............................................................................ 104 5.3.4 Biochemical methane potential ......................................................................... 106 5.3.5 Kinetic modeling ............................................................................................... 109 5.3.6 Biodegradability and principal component analysis........................................... 110 5.4 Summary .......................................................................................................... 113 Chapter 6: Treatment of municipal sludge derived hydrothermal liquefaction aqueous phase in continuous-flow activated sludge systems of wastewater treatment plants .............................. 115 6.1 Introduction ...................................................................................................... 115 6.2 Materials and methods ..................................................................................... 116 6.2.1 HTL feedstock preparation and operation ......................................................... 116 6.2.2 HTL aqueous characterization .......................................................................... 116 6.2.3 Continuous-flow activated sludge process design ............................................ 116 6.2.4 Feed preparation for activated sludge process testing ...................................... 118 6.2.5 Activated sludge process start-up, operation and monitoring ............................ 119 6.2.6 Effluent characterization ................................................................................... 120 6.2.7 Data analysis .................................................................................................... 121 6.3 Results and discussion ..................................................................................... 121 6.3.1 HTL aqueous characteristics ............................................................................ 121 6.3.2 Activated sludge system treatment performance monitoring ............................. 123 6.3.3 Chemical oxygen demand ................................................................................ 126 6.3.4 Biochemical oxygen demand ............................................................................ 128 xi 6.3.5 Total ammonium nitrogen (NH4+-N) .................................................................. 129 6.3.6 Orthophosphate phosphorus (PO4-3-P) ............................................................. 130 6.3.7 Total phenolic compounds ................................................................................ 132 6.3.8 Settling and sludge volume index ..................................................................... 133 6.3.9 UV absorbance ................................................................................................ 135 6.4 Summary .......................................................................................................... 136 Chapter 7: Conclusion .......................................................................................................... 138 7.1 Overall summary and conclusions .................................................................... 138 7.2 Limitations and recommendations for future research ...................................... 140 Bibliography .......................................................................................................................... 142 Appendices ............................................................................................................................. 178 Appendix A: Calibration curves for the colorimetric HTL aqueous characterization tests and one factor model graphs of HTL aqueous characteristics ................................................. 178 Appendix B: Organic compounds detected in HTL aqueous samples obtained at various reaction temperatures and normal distribution test of characterization data. .................... 182 Appendix C: Positive control test results and characterization of the inoculum used in biochemical assays, contents of biochemical assays, one factor model graphs and normality test results of HTL aqueous biodegradability presented in Chapter 5. .............................. 193 Appendix D: Characterization of aerobic and anaerobic inoculum used in biochemical tests, pH levels during biochemical methane potential tests, feedstock total solids and volatile solids levels, and HTL aqueous contents per volatile solids in Chapter 5. ........................ 199 Appendix E. Two-sample t-test results of Chapter 6 ......................................................... 203 xii List of tables Table 2.1 Fuel quality comparison of HTL bio-crude and pyrolysis oil. .......................................19 Table 2.2. Hydrothermal liquefaction aqueous phase valorization and treatment studies in the literature. ...................................................................................................................................23 Table 2.3. Overview of reaction conditions and product yields in batch hydrothermal liquefaction research. ...................................................................................................................................26 Table 3.1. ANOVA test results of HTL aqueous characterization data. .......................................60 Table 3.2 Regression equations of HTL aqueous characterization parameters. ........................61 Table 3.3. Potential effect of HTL aqueous characteristics on on-site biological treatment options. .................................................................................................................................................64 Table 4.1. Buffer, nutrient, and inhibitor solutions added to biochemical oxygen demand tests. 71 Table 4.2. Characterization data of mixed municipal sludge used as feedstock in hydrothermal liquefaction process (n=3). ........................................................................................................76 Table 4.3. Characterization results of HTL aqueous samples obtained from mixed sludge at various conditions. ....................................................................................................................77 Table 4.4. Modified Gompertz model kinetic parameters of five-day biochemical oxygen demand of HTL aqueous samples. .........................................................................................................79 Table 4.5. Modified Gompertz model kinetic parameters of mesophilic biochemical methane potential results of HTL aqueous samples. ................................................................................84 Table 4.6. Modified Gompertz model kinetic parameters of thermophilic biochemical methane potential results of HTL aqueous samples. ................................................................................87 Table 4.7 ANOVA test results for aerobic/anaerobic biodegradation assessment for HTL aqueous samples. ...................................................................................................................................90 Table 4.8. Regression equations of aerobic/anaerobic biodegradation assessment for HTL aqueous samples. .....................................................................................................................91 Table 5.1. Composition of sludge feedstocks used in hydrothermal liquefaction runs (n=3). ......99 Table 5.2. Kinetic parameters of biochemical tests performed on HTL aqueous samples obtained from HTL of various sludge types. ........................................................................................... 110 Table 6.1. Synthetic domestic wastewater formulation used in this study. ............................... 119 Table 6.2. Characterization results of hydrothermal liquefaction aqueous phase utilized for testing continuous aerobic treatment by spiking into synthetic wastewater in this study. ..................... 122 xiii List of figures Figure 1.1. Flow diagram of a typical municipal wastewater treatment plant. ............................. 1 Figure 1.2. A typical full-scale continuous-flow hydrothermal liquefaction reactor design. ........... 3 Figure 1.3. Research trend of hydrothermal liquefaction technology on various feedstocks. ...... 5 Figure 1.4. Experimental flow chart of the research chapters in the dissertation. ....................... 7 Figure 2.1. Potential hydrothermal liquefaction reactions and pathways..................................... 8 Figure 2.2. Biochemical composition distributions of various HTL substrates. ...........................10 Figure 2.3. Hydrothermal decomposition of carbohydrates........................................................12 Figure 2.4. Hydrothermal decomposition pathways of amino acids. ..........................................13 Figure 2.5. Lipid hydrothermal decomposition reactions. ...........................................................14 Figure 2.6. Recombination reaction examples in HTL. ..............................................................15 Figure 2.7. Maillard reaction pathway and products. .................................................................16 Figure 2.8 Organic compound groups that form bio-crude. .......................................................17 Figure 2.9. Typical composition of HTL gaseous phase at 350°C. .............................................20 Figure 2.10. Distribution of organic compounds in HTL aqueous phase of various biomass. .....21 Figure 2.11. van Krevelen diagram of bio-crudes produced from different substrates in the literature. ...................................................................................................................................30 Figure 2.12. Overall catalyst performance on different hydrothermal liquefaction substrates. ....31 Figure 2.13. Summary of economic analysis in the hydrothermal liquefaction process. .............32 Figure 3.1. Experimental design for HTL operating conditions for dewatered municipal mixed sludge (Severity is defined by Eq. 3.1 in Section 3.2.3). ............................................................38 Figure 3.2. (a) Parr high temperature/high pressure reactor during operation and (b) internal components of the reactor system, which include a magnetic drive mixing system, a cooling coil, and a sampling tube. .................................................................................................................39 Figure 3.3. Hydrothermal liquefaction of mixed sludge and separation of HTL aqueous ............40 Figure 3.4. (a) GENESYSâ¢ 10S UV-VIS spectrophotometer used in the colorimetric tests and (b) test tubes used for colorimetric characterization analyses. ........................................................41 Figure 3.5. (a) Gas chromatograph equipped with a flame ionization detector (GC-FID) and (b) high-performance liquid chromatograph with a refractive index detector (HPLC-RID) systems located at the Bioreactor Technology Group laboratory. ............................................................44 Figure 3.6. Dewatered mixed sludge derived HTL aqueous characterization results of (a) COD (n=6), (b) dissolved protein and total ammonia (n=6), (c) VFAs (n=6), (d) total alkalinity and pH xiv (n=4), (e) phosphorus (PO4-3-P) (n=6), (f) total carbohydrates and total phenolic compounds (n=6) at various reaction temperatures and retention times (GAE: gallic acid equivalent). ..................46 Figure 3.7. Glycerol concentrations of HTL aqueous samples (n=6) derived from dewatered municipal mixed sludge at various reaction temperatures and retention times. .........................54 Figure 3.8 Major compound groups detected by GC-MS in HTL aqueous samples derived from dewatered municipal mixed sludge under conditions of (a) 290°C-15 min, (b) 325°C-15 min, and (c) 360°C-15 min. ......................................................................................................................56 Figure 3.9. Hydrothermal formation pathways of compounds and groups detected in GC-MS analysis. ....................................................................................................................................57 Figure 3.10. (a) Ultimate analysis (n=2), and (b), (c) trace element analysis (n=4) results of HTL aqueous samples derived from dewatered municipal mixed sludge (MS) at various reaction temperature and retention times. ..............................................................................................59 Figure 3.11 Predicted values of (a) COD, (b) dissolved protein, (c) total ammonia, (d) total alkalinity, (e) pH, (f) phosphorus (PO4-3-P), (g) total carbohydrates, (h) total phenolics, and (i) glycerol by the regression analysis for HTL aqueous derived from dewatered municipal mixed sludge at various reaction temperatures and retention times (GAE: gallic acid equivalent). ......62 Figure 3.12. Predicted values of (a) total VFAs, (b) acetic acid, (c) propionic acid, (d) butyric acid, (e) valeric acid, and (f) isovaleric acid obtained by the regression analysis for HTL aqueous derived from dewatered municipal mixed sludge at various reaction temperatures and retention times. ........................................................................................................................................63 Figure 4.1. On-site HTL aqueous treatment pathways of (a) scenario I: aerobic treatment, (b) scenario II: mesophilic anaerobic digestion, scenario III: thermophilic anaerobic digestion, and (c) scenario IV: dark fermentation. ..................................................................................................69 Figure 4.2. (a) Image of a 500 mL borosilicate respirometer reactor with 30% KOH filled CO2 absorption tube in the headspace and (b) RSA PF-8000 respirometer during measurement. ....72 Figure 4.3. Images of (a) Agilent 7820A gas chromatography - thermal conductivity detector, (b) New Brunswick Innova®44 shaking incubators, and (c) batch of BMP assay bottles. ...............73 Figure 4.4. Biochemical hydrogen potential assays. ..................................................................75 Figure 4.5. Respirometric oxygen uptake and uptake rate results of municipal sludge derived HTL aqueous samples obtained under various temperature and retention times (shaded regions indicate error bands with standard deviation). ...........................................................................78 Figure 4.6. Mesophilic biochemical methane potential results of various HTL aqueous samples (n=6). ........................................................................................................................................80 xv Figure 4.7. Volatile fatty acid concentrations of mesophilic biochemical methane potential reactors throughout the test (n=2). ..........................................................................................................82 Figure 4.8. pH of mesophilic biochemical methane potential reactors throughout the test. ........83 Figure 4.9. Thermophilic biochemical methane potential test results of HTL aqueous samples (n=6). ........................................................................................................................................86 Figure 4.10. Volatile fatty acid concentrations of thermophilic biochemical methane potential reactors throughout the test (n=2). ............................................................................................88 Figure 4.11. pH of thermophilic biochemical methane potential reactors throughout the test. ....89 Figure 4.12. Predicted values of (a) five-day biochemical oxygen demand, (b) aerobic biodegradability index (BOD5/COD), (c) mesophilic methane potential, (d) mesophilic anaerobic biodegradability, (e) thermophilic methane potential, and (f) thermophilic anaerobic biodegradability by the regression analysis for HTL aqueous derived from dewatered municipal mixed sludge at various reaction temperatures and retention times. .........................................92 Figure 5.1. Chemical characteristics of (a) chemical oxygen demand, (b) protein, total ammonia, (c) fatty acids, (d) orthophosphate-P, total sugar, (e) total alkalinity, pH, and (f) total phenolic compounds of HTL aqueous derived from various sludge types (n=6). ................................... 100 Figure 5.2. Quantitative analysis of selected N-heterocycles compounds and ketones in HTL aqueous samples using gas chromatography-mass spectrometry (GC-MS). .......................... 104 Figure 5.3. Specific oxygen uptake and oxygen uptake rate of HTL aqueous samples measured in respirometric biochemical oxygen demand tests. (n=2) ....................................................... 105 Figure 5.4. Cumulative methane production curves, methane production rates, and volatile fatty acids concentrations of mesophilic biochemical methane potential tests conducted on HTL aqueous samples derived from various sludge types (n=6). .................................................... 107 Figure 5.5. Cumulative methane production curves, methane production rates, and volatile fatty acids concentrations of thermophilic biochemical methane potential tests conducted on HTL aqueous samples derived from various sludge types (n=6). .................................................... 108 Figure 5.6. (a) Biodegradability levels of HTL aqueous samples obtained from HTL of various sludge types and (b) Pearson correlation plot of feedstock characteristics, HTL aqueous composition, and HTL aqueous biodegradability. .................................................................... 112 Figure 6.1. (a) Ender 3-V2 3D printer and (b) designed lab-scale treatment units in 3D CAD software. ................................................................................................................................. 117 Figure 6.2. Photos of (a) 3D printed, two parallel activated sludge treatment systems during the testing period, (b) top view of an empty aeration reactor with a 10 cm diameter diffuser at the xvi bottom, and (c) HTL aqueous spiked reactor in operation, showing clear effluent and settled sludge. .................................................................................................................................... 118 Figure 6.3. (a) pH, (b) temperature, and (c) dissolved oxygen levels of aeration tanks in control and test systems treating synthetic wastewater and HTL aqueous spiked synthetic wastewater, respectively. ............................................................................................................................ 124 Figure 6.4. Mixed liquor volatile suspended solids levels of aeration tanks in control and test systems treating synthetic wastewater and HTL aqueous spiked synthetic wastewater, respectively. ............................................................................................................................ 125 Figure 6.5. (a) Chemical oxygen demand levels of aeration tank influents and effluents, (b) chemical oxygen demand removal performances of aeration tank, and (c) removal performance of HTL aqueous originated chemical oxygen demand in aeration tanks working at 4.5-h and 6-h hydraulic retention times (n=3). ............................................................................................... 127 Figure 6.6. Influent and effluent five-day biochemical oxygen demand levels (BOD5) of aeration tanks in control and test systems treating synthetic wastewater and HTL aqueous spiked synthetic wastewater, respectively.......................................................................................................... 129 Figure 6.7. Influent and effluent total ammonium nitrogen concentrations (NH3-N) of aeration tanks in control and test systems treating synthetic wastewater and HTL aqueous spiked synthetic wastewater, respectively.......................................................................................................... 130 Figure 6.8. Influent and effluent phosphorus concentrations (PO4-3-P) of aeration tanks in control and test systems treating synthetic wastewater and HTL aqueous spiked synthetic wastewater, respectively. ............................................................................................................................ 131 Figure 6.9. Influent and effluent total phenolic compound concentrations of aeration tanks in control and test systems treating synthetic wastewater and HTL aqueous spiked synthetic wastewater, respectively.......................................................................................................... 133 Figure 6.10. (a) Total suspended solids levels reached in settling tank underflows of test and control systems treating synthetic wastewater and HTL aqueous spiked synthetic wastewater and (b) sludge volume index levels of mixed liquors from aeration tanks. ...................................... 134 Figure 6.11. Influent and effluent UV absorbance levels (at 254 nm) of aeration tanks in control and test systems treating synthetic wastewater and HTL aqueous spiked synthetic wastewater, respectively. ............................................................................................................................ 136 xvii List of abbreviations AD: Anaerobic Digestion ANOVA: Analysis of Variance AP: Apatite Phosphate BHP: Biochemical Hydrogen Potential BMP: Biochemical Methane Potential BNR: Biological Nutrient Removal BOD: Biochemical Oxygen Demand BOD5: Five-day Biochemical Oxygen Demand CCC: Circumscribed Central Composite Design CH4: Methane CHG: Catalytic Hydrothermal Gasification CO2: Carbon Dioxide COD: Chemical Oxygen Demand DCM: Dichloromethane DIET: Direct Interspecies Electron Transfer DO: Dissolved Oxygen DS: Digested Sludge ER: Energy Recovery F/M: Food to Microorganism Ratio GAE: Gallic Acid Equivalent GC-FID: Gas Chromatography-Flame Ionization Detector GC-MS: Gas Chromatography-Mass Spectrometry GC-TCD: Gas Chromatography-Thermal Conductivity Detector HHV: Higher Heating Value HMF: Hydroxymethylfurfural HPLC: High-Performance Liquid Chromatography HRT: Hydraulic Retention Time HT/HP: High Temperature â High Pressure HTL: Hydrothermal Liquefaction ICP-MS: Inductively Coupled Plasma Mass Spectrometry xviii LBL: Lawrence Berkeley Laboratory MCFA: Medium-Chain Fatty Acids MLSS: Mixed Liquor Suspended Solids MLVSS: Mixed Liquor Volatile Suspended Solids NAIP: Non-Apatite Inorganic Phosphate NO2: Nitrogen Dioxide OUR: Oxygen Uptake Rate PERC: Pittsburgh Energy Research Center PETG: Polyethylene Terephthalate Glycol PNNL: Pacific Northwest National Laboratory PS: Primary Sludge RID: Refractive Index Detector SCFA: Short-Chain Fatty Acids SRT: Sludge Retention Time SS: Secondary Sludge STP: Standard Temperature and Pressure SVI: Sludge Volume Index TAN: Total Ammonium Nitrogen TPU: Thermoplastic Polyurethane TS: Total Solids TSPS: Thickened Screened Primary Sludge TSS: Total Suspended Solids TWSS: Thickened Waste Secondary Sludge UASB: Up-flow Anaerobic Sludge Blanket UBOD: Ultimate Biochemical Oxygen Demand v/v: Ratio by volume VFA: Volatile Fatty Acids VS: Volatile Solids VSS: Volatile Suspended Solids wt.%: By weight percentage WWTP: Wastewater Treatment Plant xix Acknowledgements Under the mentorship of Dr. Cigdem Eskicioglu, my Ph.D. journey was enriched by her patience and invaluable guidance. The outstanding lab environment she meticulously developed over the years significantly enhanced my research experience. I am profoundly grateful for her mentorship and the substantial impact it has had on my academic and personal growth. I would also like to thank my Ph.D. committee members, Dr. Abbas Milani and Dr. Nicholás Peleato, for investing their valuable time and offering crucial insights that have greatly contributed to the depth and quality of my research. I would like to thank my colleagues in the UBC Bioreactor Technology Group, starting with Dr. Huan Liu and Dr. Timothy Abbott, and extending my gratitude to Dr. Resty Nabaterega, Shiva Khoei, Dr. Vikas Kumar, Abigail Stokes, Dr. Gokce Kor Bicakci, Ronald Kizza, Milad Goodarzi, and Farid Alizad Oghyanous, for their friendship, support, and valuable contributions over the many years of laboratory work we have undertaken. Special thanks go to Jacky Takeuchi, the technician of the UBC Bioreactor Technology Group, and Dr. Parmila Devi, the research engineer, for their exceptional support and contributions. Furthermore, I am grateful to the technicians at the UBC Okanagan School of Engineering for their invaluable assistance and expertise. Lastly, I would like to express my profound gratitude to our industrial partner, Metro Vancouver - Collaborative Innovations, for generously providing samples, extensive data, insightful knowledge, and continuous support throughout this research. This work was funded by the Natural Sciences and Engineering Research Council of Canada (NSERC) and Metro Vancouver Industrial Research Chair Program (IRCPJ 548816â18) in Advanced Resource Recovery from Wastewater. xx Dedication 1 Chapter 1: Introduction This chapter describes the engineering problem that this study was conducted to address by explaining present municipal wastewater treatment technology and approaches to municipal sludge treatment. 1.1 Background In a typical wastewater treatment plant (WWTP) organic material is removed through two main process groups, primary and secondary treatments, as depicted in Figure 1.1. In the primary treatment phase, inorganic suspended particles are removed through screening and grit removal processes, followed by the separation of suspended organics via settling in primary clarifiers, resulting in the formation of primary sludge. Primary sludge (PS) retains approximately 50% of the carbon present in the raw wastewater. During secondary treatment, the remaining 50% of carbon, which is in dissolved form, is targeted through aerobic treatment processes. Aerobic treatment sequesters 20% of the carbon in the form of excess microorganism biomass, called secondary sludge (SS), converts another 20% into CO2, and leaves the remaining 10% in the treated wastewater. In total, mixed sludge (a mixture of PS and SS) comprises approximately 70% of the carbon that enters the WWTP. Therefore, efficient valorization of municipal sludge is essential for enhancing the sustainability of overall wastewater treatment processes. Figure 1.1. Flow diagram of a typical municipal wastewater treatment plant. 1.1.1 Existing sludge treatment techniques Globally, it is estimated that 380 billion m3 of wastewater is produced each year, with an expected increase of 51% by 2050 [1] due to increasing population and urbanization. As a result of this wastewater production, approximately 45 million dry tons of municipal sludge are generated 2 worldwide annually [2]. The widespread methods for valorization or disposal of municipal sludge currently include anaerobic digestion (AD), composting, incineration, land application (for agricultural purposes), and landfilling. AD is an efficient sludge treatment method that produces a biofuel, methane, while reducing the mass of the sludge and stabilizing it biologically [3]. However, this technology is constrained by biological limitations and can achieve only 50% organic matter removal from sludge [4]. Composting of municipal sludge has been reported to result in greenhouse gas emissions of up to 216 g/kgdry sludge CO2, 118 g/kgdry sludge NO2, and 10 g/ kgdry sludge CH4 [5]. Incineration of municipal sludge is an energy consuming process due to the moisture content of the feedstock. Along with incineration some thermochemical conversion methods, such as pyrolysis and torrefaction, have been investigated for municipal sludge treatment. Moreover, these methods also require drying, and they generate greenhouse gases during the conversion process [6]. Lastly, untreated municipal sludge is not suitable for land application because it contains heavy metals and pathogens [7]. For a similar reason, landfilling municipal sludge has become increasingly challenging due to evolving regulations that restrict this practice. While AD is the current best sludge treatment method, only a minority of WWTPs are equipped with an anaerobic digester. In the US, this figure is below 10% [15], whereas in Europe, less than 12% of municipal sludge undergoes AD [16] meaning that many WWTPs are not equipped with digesters. Furthermore, a recent study revealed that even in WWTP with anaerobic digesters, a significant amount of CH4 is still released into the atmosphere [8]. Thus, there is a pressing need for fundamental changes and significant advancements in current municipal sludge valorization technology to incorporate WWTPs into the circular economy. 1.1.2 Hydrothermal liquefaction Hydrothermal liquefaction (HTL) is a thermochemical biomass conversion method that could be a good candidate for municipal sludge valorization in the near future [9]. The HTL process works similarly to the natural formation of petroleum from biomass. In a sealed reactor system at 300-350°C temperature and 85-165 bar pressure, biomass is hydrolyzed, decomposed, and recombined into a petroleum-like biofuel called bio-crude. In previous studies, it was found that in optimum HTL conditions, 69% carbon recovery - equivalent to 70.8% energy recovery - can be achieved by HTL in a WWTP processing municipal sludge [10]. Bio-crude produced through HTL can be further processed in preexisting oil refineries to produce sustainable aviation [11] and marine fuels [12]. The HTL process produces two additional streams known as hydrochar and HTL aqueous. The hydrochar stream is the carbonaceous solid residue of the process containing 7.2-8.5 wt.% P, which allows for approximately 90% P recovery from municipal sludge [13]. The 3 HTL aqueous phase, on the other hand, is the liquid medium in which the HTL reactions take place, and it makes up 80-81% of the reactor effluent when municipal sludge with 20 wt.% total solids (TS) is used as the feedstock. In WWTPs, building continuous-flow HTL systems is a promising option for pilot- and full-scale applications. Although a high-pressure pump, up to 220 bars, is required for feeding the reactor, the higher heat recovery rate makes these systems effective. Anastasakis et al. [14] reported heat recovery rates between 66.5-79.9% in continuous-flow HTL system. Another major advantage of these systems is the high heating (ramping) rate, which was reported in the literature as 100°C/min [14,15] or even 200-400°C/min [16]. HTL conditions with lower ramping times lead to a higher bio-crude yield [17â19] and decrease the required reactor volume significantly. HTL reactors need to be built from expensive, heat and corrosion-resistant alloys due to the high temperatures that they are operated on. Therefore, smaller reactor volumes can lower the capital costs drastically and make continuous-flow HTL systems more affordable. A flow diagram of a continuous-flow HTL reactor, that can be integrated into WWTPs for municipal sludge valorization, is presented in Figure 1.2. Figure 1.2. A typical full-scale continuous-flow hydrothermal liquefaction reactor design. 1.1.3 Hydrothermal liquefaction technology to date The first HTL research was performed by Professor Ernst Berl in Darmstadt, Germany. His results in the 1920s indicated that bio-crude production from biomass in hot water with alkali catalyst was possible [20]. In 1933, he continued his HTL research at Carnegie Institute of 4 Technology in Pittsburgh, Pennsylvania. In the 1940s, his research attracted significant attention due to his publications and magazine articles [21â25]. However, no significant progress was achieved until the petroleum fuel shortage in the late 1960s [26] when the Arab oil embargo in the mid-1970s [27] revealed the necessity of developing a sustainable fuel production process. In the 1970s and 1980s, some of the most important studies were performed by Pittsburgh Energy Research Center (PERC) [28], Lawrence Berkeley Laboratory (LBL) [29], Rust International Corporation [26] in the USA, and Shell Research Institute in the Netherlands. Pacific Northwest National Laboratory (PNNL) in Washington, USA, has been the pioneering research institute in the HTL field since 1990. In recent years, Metro Vancouver has been planning to build a ten wet ton per day capacity demonstration HTL facility at its Annacis Island WWTP (Delta, BC, Canada) using this technology developed by PNNL and licensed to Genifuel Corporation (Utah, U.S.). 1.2 Objectives and hypotheses The primary objective of this research was to determine the effect of varying HTL conditions and the ratio of primary to secondary sludge in the feedstock on the characteristics and biodegradability of the HTL aqueous phase (Hypotheses 1,2, and 3). This objective involved identifying the most suitable on-site biological treatment options, including aerobic, mesophilic anaerobic, thermophilic anaerobic, and dark fermentation processes, for HTL aqueous treatment. The second objective of this study is to examine how blending HTL aqueous phase with incoming wastewater impacts aerobic treatment, effluent properties, and subsequent settling and UV disinfection processes (Hypothesis 4). To meet these objectives, the following hypotheses were tested: Hypothesis 1: Within the HTL condition ranges of 290-360°C and 0-30 minutes, significant variations in the characteristics of the HTL aqueous phase will be observed. Hypothesis 2: As HTL intensity (temperature and holding time) increases, the activity of both aerobic and anaerobic cultures may decrease due to an increased inhibitory effect. Hypothesis 3: Variations in the ratio of primary to secondary sludge in the feedstock will cause significant changes in the characteristics of the aqueous phase produced by hydrothermal liquefaction. Hypothesis 4: HTL aqueous can be returned to the headworks of the wastewater treatment plant without causing a significant disturbance on the downstream processes of aerobic treatment, secondary settling, and UV disinfection processes. 5 1.3 Novelty The HTL process is a fairly new thermochemical biomass conversion technology for compared to hydrothermal gasification, hydrothermal carbonization, and pyrolysis. Although the HTL process is a century old, recent technological developments, such as enhanced heat exchangers that allow 80% of the heat to be recycled [14], made it efficient and applicable. Therefore, most of the HTL research has been conducted in the last decade. Many of these research projects and journal articles focused on using lignocellulosic biomass, microalgae, macroalgae, cyanobacteria, and food waste as feedstock in the HTL process. However, literature on applications of HTL aqueous in the wastewater treatment field is quite limited. The number of HTL studies performed on various biomass sources from 2000 to 2019 is given in Figure 1.3. Figure 1.3. Research trend of hydrothermal liquefaction technology on various feedstocks. Sludge management is one of the major daily operating costs in the WWTPs. Although anaerobic digestion can decrease the volume of municipal sludge by 50%, the residue, digested municipal sludge, still must be composted, combusted, or landfilled. The biggest cost that is caused by digested sludge is its transportation. In the HTL process, the only solid byproduct is hydrochar, which has an insignificant volume compared to the digested sludge. Furthermore, hydrochar can be valorized in nutrient recovery, combustion, gasification, adsorption and it can be used as a heterogenous catalyst [30]. The most significant unknown about the application of HTL process in wastewater treatment is the treatment pathway of HTL aqueous which consists of 81 wt.% of HTL reactor effluent. This research will be the first in literature investigating the on-site 6 treatment of HTL aqueous in a WWTP and its effects on the existing treatment units, bringing the technology one step closer to a full-scale application. To the best of our knowledge, there is limited information in the literature about on-site biological treatment of municipal sludge derived HTL aqueous phase. Particularly, the literature lacks information on the continuous-flow aerobic treatment of HTL aqueous, with only limited details available regarding batch aerobic treatment. Additionally, no studies have been conducted on the characterization of the HTL aqueous phase derived from municipal sludge that investigate how its characteristics vary with HTL reaction conditions and the composition of the feedstock sludge. 1.4 Structure of the dissertation The experimental flow chart of the dissertation is presented in Figure 1.4. This dissertation is prepared in article-based format with the following chapters: â¢ Chapter 1: Background information, description of the engineering problem, objectives and hypotheses, tasks, and novelty. â¢ Chapter 2: Literature review about the HTL process, hydrothermal chemistry, products of HTL, and HTL aqueous treatment. â¢ Chapter 3: Detailed characterization of aqueous phase produced under various HTL operating conditions, along with mathematical analysis to correlate its characteristics with the specific reaction conditions. â¢ Chapter 4: Evaluating municipal sludge-derived HTL aqueous phase treatment through on-site biological methodsâsuch as aerobic, mesophilic anaerobic, thermophilic anaerobic, and dark fermentationâand analyzing how treatment performance correlates with HTL conditions. â¢ Chapter 5: Detailed characterization and biodegradability assessment of HTL aqueous phases derived from various sludge types collected from multiple wastewater treatment plants. â¢ Chapter 6: Assessing the treatment of HTL aqueous in a continuous-flow activated sludge system, focusing on its impact on the aerobic process, downstream processes, and the effluent discharge quality. â¢ Chapter 7: Summarizing the conclusions drawn from chapters and discuss the limitations of the study. 7 Figure 1.4. Experimental flow chart of the research chapters in the dissertation. a: Optimized HTL condition for maximum biocrude production [10]. 8 Chapter 2: Literature Review This section provides a summary of existing literature on HTL, covering aspects such as HTL process mechanisms, HTL products, the influence of feedstock on product yield and quality, and studies on HTL aqueous phase treatment. 2.1 Hydrothermal liquefaction chemistry Reaction pathways in HTL can be separated into three main groups: depolymerization (hydrolysis), decomposition, and recombination (repolymerization). Although many of the exact HTL reactions remain unclear due to the large number of organic compound types, studies in the literature were able to reveal some of the major pathways. Potential reaction pathways of HTL are summarized in Figure 2.1. The selectivity of these reactions can vary depending on pH, severity of HTL process (temperature, pressure, ramping and retention time), type and concentration of solvent, and type and concentration of catalyst. Figure 2.1. Potential hydrothermal liquefaction reactions and pathways [9]. 9 In order to investigate the effect of HTL on model (feedstock) compounds, Lu et al. (2018) studied HTL of soybean oil, soy protein, cellulose, xylose and ligninâs binary, ternary, quaternary and quinary mixtures, and revealed the effect of substrate composition on HTL products at 350°C for 30 minutes. For individual model compounds, they reported bio-crude yields of 82 wt.% for lipids, 21.1 wt.% for proteins, 4.6 wt.% for cellulose, 6.6 wt.% for xylose, and 1.4 wt.% for lignin. For the binary mixtures, they reported yield increases for protein-carbohydrate and protein-lignin mixtures; and cellulose-xylose mixture was found to be ineffective for bio-crude production. A mixture of lipid and lignin showed an antagonistic interaction. The highest energy recovery was observed for the quinary mixture, underlining the importance of a balanced substrate in the HTL process to benefit from synergistic effects. Biller and Ross [32] used representative model compounds in HTL processes, such as albumin, asparagine, glutamine, soya protein for proteins, glucose and starch for carbohydrates, and sunflower for lipids. Similar to the other studies in the literature, they reported bio-crude yields of 86% for lipids, 23% for proteins, and 10% for carbohydrates. According to the average biochemical compositions presented in Figure 2.2, microalgae and municipal sludge are the most suitable substrate types for taking advantage of Maillard reactions. In these reactions, reducing sugarâs aldehyde group reacts with amino groups via the formation of Schiffâs base and Amadori rearrangement [33]. Microalgae composition can vary depending on the species. Therefore, some species might not be suitable for the HTL process. On the other hand, municipal sludge has fewer varying characteristics, and its composition can be slightly adjusted by changing the primary and secondary sludge blending ratio. There are two main composition differences between the two sludge types. The first is the high protein content of secondary sludge that can be used for balancing the substrate composition to take advantage of Maillard reactions. The second is the high lipid content of primary sludge [34,35], which can be directly converted into bio-crude in the HTL process. Figure 2.2 also shows that macroalgae and lignocellulosic biomasses are carbohydrate-rich due to their cellulose and hemicellulose contents. This causes them to have a lower bio-crude yield compared to the other types of biomasses. Excessive organic acids form during the process and decrease the pH as a result of the HTL pathway of carbohydrates. Acidic conditions cause HTL reactions to favor hydrochar production rather than bio-crude. Therefore, to balance the pH, an alkaline catalyst needs to be used in HTL of macroalgae and lignocellulosic biomasses. Like microalgae, cyanobacteria biochemical composition also differs depending on the species. On 10 average, cyanobacteria have high protein content and less lipid content compared to microalgae. This causes bio-crude production from cyanobacteria to mostly rely on Maillard reactions. Figure 2.2. Biochemical composition distributions of various HTL substrates [9]. In summary, from the results of HTL studies on model compounds, lipids can be converted into bio-crude in high yields. Carbohydrates and proteins, on the other hand, need to be in a specific ratio to benefit from Maillard reactions. There are varying results in the HTL literature for the optimum ratio for carbohydrates and proteins. Therefore, revealing the optimum ratio is required to fill that gap in the literature either by using model compounds or co-liquefaction of substrates. 2.1.1 Depolymerization Depolymerization (hydrolysis) reaction is the first step in the HTL process. In these reactions, lipids are broken down into fatty acids and glycerol, proteins are broken down into amino acids, and carbohydrates are broken down into monosaccharides [36]. No significant bio-crude forming occurs at this stage. If the severity of the HTL process is adjusted only for the depolymerization step, effluent can be used in fermentation (biological) processes to produce biofuels (i.e., methane, hydrogen, and ethanol). Depending on the substrateâs characteristics, the level of HTL temperature required for depolymerization varies between 150-250°C. 020406080100% of Volatile Solids Carbohydrate Protein Lipid LigninMixed Municipal SludgePrimaryMunicipal SludgeSecondaryMunicipal SludgeMicroalgaeMacroalgaeCyanobacteriaHerbaceous Lignocellulosic Biomass 11 Hydrothermal depolymerization is a widely investigated technology under the name of hydrothermal pretreatment. The technology is mostly used to increase the biodegradability of municipal sludge or enhance the hydrolysis of lignocellulosic biomass [37â39]. Most of the candidate substrates, such as woody and herbaceous plants, microalgae, macroalgae, cyanobacteria, and manure, contain lignocellulose in different ratios. All the monosaccharides obtained from the hydrothermal depolymerization of lignocellulosic biomass are valuable platform chemicals [40] that have great potential in developing sustainable chemical and biofuel industries. In the HTL, these monosaccharides can contribute to bio-crude formation by reacting with amino acids. Lignin depolymerization products, monolignols, on the other hand, remain in the aqueous phase mostly in phenol and catechol forms [41]. Monolignols were found the contribute to bio-crude in HTL if a water â phenol mixture aqueous phase is used [42,43]. 2.1.2 Decomposition Hydrolyzed smaller molecules go through several thermal decomposition reactions such as dehydration, decarboxylation, decarbonylation, deamination, dehydrogenation, and some bond cleavages due to high temperatures. The decomposition of different compound groups in HTL occurs in the 180-340°C range [44]. In the HTL process, carbohydrates are the first group to start decomposing at 180°C. Over 200°C, proteins and lipids begin to decompose. For the total decomposition of proteins and lipids, temperatures of 300°C and 640°C are required, respectively [45]. Since some lipids, such as long-chain fatty acids, are directly included in bio-crude, total decomposition of lipids is not required in the HTL process. Obtaining the highest efficiency in the HTL process is directly related to the balance of decomposition reactions. The intensity of HTL operating conditions should be high enough to allow organic compounds to decompose into bio-crude forming compounds, but at the same time, it should not be too severe to decompose bio-crude into gases. For wastewater treatment applications, decomposition reactions in the HTL process allow for destroying most pharmaceuticals [46] and converting microplastics into bio-crude [47]. The first degrading group, carbohydrates, have several decomposition pathways ending with short-chain (volatile) fatty acids, basic alcohols, ketones, furan compounds, furanic acids, and furanic alcohols. Except for large furanic molecules, decomposition of carbohydrates does not contribute to bio-crude formation significantly. However, products of carbohydrate decomposition can form larger molecules in recombination reactions. Important carbohydrate decomposition pathways are summarized in Figure 2.3. Due to the formation of furanic 12 compounds in the HTL, carbohydrate-rich substrates, such as lignocellulose and macroalgae, can form bio-crude and aqueous phase with higher oxygen percentages than the other substrates. Figure 2.3. Hydrothermal decomposition of carbohydrates [9]. The second decomposing group in HTL is proteins. Protein decomposition pathways mostly consist of deamination and decarboxylation reactions of amino acids. Depending on the type of amino acid, decomposition temperature, reaction pathways, and products differ [48]. Amino acid deamination reaction is the primary source of ammonia in the HTL aqueous. Therefore, HTL aqueous phase of protein-rich substrates can have up to 16 g/L ammonia, causing pH levels over 8 [49]. Protein decomposition end products are mostly simple amines, amides, aldehydes, short-chain fatty acids, and N-heterocycles. Like O-heterocyclic compounds (furan derivatives) in carbohydrate decomposition, large N-heterocyclic molecules are the only significant direct bio-crude contribution of protein decomposition. Decomposition reaction examples of some amino acids (indicated in red) are presented in Figure 2.4. 13 Figure 2.4. Hydrothermal decomposition pathways of amino acids [9]. In terms of pH, substrateâs carbohydrate and protein amount should be in a specific ratio in the HTL so that decomposition products, carboxylic acids and ammonia, would balance each other. If the substrate is protein-rich, alkaline conditions can occur in the HTL, which is reported to be beneficial for bio-crude production [50]. On the other hand, in HTL of carbohydrate-rich substrates, such as lignocellulosic biomass, acidic conditions can promote the formation of hydrochar and lower the bio-crude yield [51]. The last decomposing group in HTL process is lipids. Lipids have the highest contribution to bio-crude production among biomass fractions due to their larger molecule sizes. This contribution occurs by the decomposition products of long chain fatty acids, alkanes, and alkenes. Due to the formation of phosphoric acid and long chain fatty acids, lipid decomposition can cause 14 a pH drop. The majority of sterols also decompose thermally [52] into various smaller molecules, such as ketones and aldehydes, that can later go into recombination reactions. Lipid decomposition pathways are presented in Figure 2.5. Figure 2.5. Lipid hydrothermal decomposition reactions [9]. 2.1.3 Recombination Various reactive fragments formed by the decomposition reactions start to recombine (repolymerize) above 300°C into bio-crude compounds [53]. Large molecules formed in these reactions contribute to bio-crude formation. Basic molecules, on the other hand, stay in the aqueous phase. Most of the organic molecule groups in bio-crude, such as aromatics, ketones, amides, amines, and esters, form in the recombination reactions of long-chain fatty acids. Also, complex alcohol molecules of bio-crude come from the hydration and cyclization reactions of alkenes. Examples of recombination pathways and products are given in Figure 2.6. 15 Figure 2.6. Recombination reaction examples in HTL [9]. One of the most important recombination reactions is the Maillard reaction taking place between reducing sugars and amino acids. While the thermochemical changes are simplified as the Maillard reaction, it is actually a combination of reactions with complex pathways [54]. Although Maillard products, O and N heterocyclic compounds, decrease the quality of bio-crude due to the oxygen and nitrogen in their structures, the reaction is essential in terms of achieving higher carbon recovery in the bio-crude phase. The impurity caused by the Maillard reaction products can be treated in refineries prior to the distillation process. Studies performed on model compounds revealed that if Maillard reaction is not occurring in an HTL process, bio-crude yields obtained from carbohydrates and proteins were in the 7-18% range, while bio-crude yields up to 40% were reported when carbohydrates and proteins are used together in the HTL process [55,56]. Maillard reaction pathways are summarized in Figure 2.7. 16 Figure 2.7. Maillard reaction pathway and products [9]. Recombining fragments also include some of the bio-crude compounds which form coke and contribute to hydrochar [57]. Coke formation can be minimized with the application of high pressurized H2 during the recombination step [58]. At this point, severe HTL conditions can start secondary cracking of bio-crude and decrease the process efficiency. 2.2 Hydrothermal liquefaction products 2.2.1 Bio-crude Bio-crude is a dark brown, free-flowing, hydrocarbon-rich liquid obtained in the HTL process as the primary fuel product. When produced from renewable or waste biomass sources, it is accepted as a suitable and sustainable energy source replacement for fossil fuels [59]. Bio-crude contains saturated fatty acids, alkanes, alkenes, aromatic hydrocarbons, N and O heterocyclic compounds, ketones, alcohols, esters, amines, and amides. Concentrations of these organic compound groups in the bio-crude are highly dependent on carbohydrate, protein, and lipid composition of the HTL substrate, HTL process conditions, and bio-crude separation method. Examples for the compound groups that form bio-crude are presented in Figure 2.8. 17 Bio-crudeOrganic acidsPalmitic acidPentadecanoic acidMyristic acidBenzenepropanoic acidDifluorophosphoric acidn-Hexadecanoic acidOleic acidOctadecanoic acidDodecanoic acidTetradecanoic acidAlkanesDodecaneTetradecaneHexadecaneHeptadecaneOctadecanePentadecaneHentriacontaneHeptacosaneEicosanePentacosaneAlkenes2,5 Octadiene1,4-Pentadiene8-Heptadecene5-OctadeceneCyclooctene1-Heptadecene3-EicosenePhytenePentadeceneHexadeceneAromatic Hydrocarbonsβ-Ylangeneβ-PhellandreneTrans-Z-alpha-Bisabolene epoxideHydroquinoneMethylphenanthreneBenzo[a]pyreneTrimethyl benzene<1,2,4->Benzene, 4-(2-butenyl)-1,2-dimethyl-, (E)-2,4-CyclohexadieneCyclopentane, 1,1,3-trimethylKetones1,2 cyclopentanedioneCholestan-3-one, (5-alpha)Dihydroactinidiolide3-Penten-2-one, 4-methyl-2-Pentanone, 4-hydroxy-4-methyl-1,5-Heptadien-4-one,3,3,6-trimethyl-2,5-Heptadien-4-one, 2,6-dimethyl-cis-2-Propanone2-Nonanone, 9-hydroxy2-NonadecanoneAlcoholsMenthol, cis-1,3,trans-1,4-DihydrocholestrolStigmastenol2-Methyl-2-cyclopenten-1-ol1-TridecanolEugenol1,4-Benzene diol, 2-methoxyHomovanillyl alcoholMethyl syringolMethyl guaiacolO-HeterocyclicFuran, 2,5-dihydro-2,5-dimethyl-5-Methyloxazole3-Acetyl-2,5-dimethyl furanTetrahydropyran Z-10-dodecenoate2-FurancarboxaldehydeEthanone, 1-(2-furanyl)-Benzofuran2(3H)-Furanone, 5-ethyldi2-Oxo-3-methyl-cis-perhydro-1,3-benzoxazine5-Methyl furfuralN-Heterocyclic1-ethyl-2-pyrrolidinone1-(2-Phenylethyl)-piperidine6-Oxo-2,3,5,6-tetrahydro-1H-pyrrolizine4-Pyrimidinamine, 2,6-dimethyl-5-Methyloxazole1H-Pyrazole, 4-methyl-1H-indole, 3-(3-methyl-2-butenyl)-MethylphenidateHydantoin, 1-butylPyrrolidine, 1(1-oxo-9-octadecenyl)EstersDioctyl phthalateGeranyl acetateMethyl palmitateDimethyl 2-ethyl-3-methylmaleateOxalic acid, butyl cyclohexylmethyl esterHexadecanoic acid, methyl esterOxalic acid, cyclobutyl tridecyl esterOleic acid, methyl esterOctadecanoic acid, ethyl esterOleic acid, ethyl esterAminesTrimethylamineDiphenylamineMethanamine, N-hydroxy-N-methyl-Ethenamine, N-methyl-N-nitrosoTryptaminePhenylethylamine2-Propen-1-amine1-Heptadecanamine, N,N-dimethyl-HexadecylamineOctadecylamineAmidesOleic diethanolamideTetradecanamideHexadecanamideStearamide, N.N-dimethyl-(9Z)-Octadecenamide13-Docosenamide, (Z)-N-MethyldodecanamideAcetamide, N-(2-phenylethyl)-Octanamide, N,N-dimethyl-Propanamide, 3-cyclopentyl-N-methyl Figure 2.8 Organic compound groups that form bio-crude [9]. Bio-crude is a renewable fuel that can be upgraded and used as a replacement for crude oil. It can be distilled into gasoline, jet fuel, diesel, naphtha [60], fuel oil, and a heavy fraction that can be used in engine lubricant production [61]. Contrary to fossil fuels, bio-crude based fuels lower the greenhouse gas emissions [62] owing to the carbon fixation occurring during the growth of biomass. Therefore, it can be said that fuels obtained from bio-crude are environmentally friendly, whereas they are also renewable due to their sustainable source, biomass. The efficiency of HTL process can be calculated by the amount of bio-crude obtained at the end of the process by Eq 2.1. Although biomass to bio-crude conversion percentage is mostly used as an efficiency measurement in the literature, in some cases, this parameter can be misleading. Bio-crude yield may decrease as a result of recombination reactions above 300°C; however, reactions can lower the oxygen percentage of bio-crude and increase its higher heating value (HHV). On the other hand, energy recovery (ER) parameter, which can be calculated using 18 Eq 2.2, represents both quality and quantity of bio-crude, which makes it more representative for the process efficiency measurement. ER should be used as the response of the process, especially for HTL modeling studies. Bio-crude yield (%) =Dry mass of bio-crudeDry mass of substrate usedà 100 (2.1) Energy recovery (%) =Dry mass of bio-crude Ã HHV of dry bio-crudeDry mass of substrate Ã HHV of dry substrateà 100 (2.2) Bio-crude is physically quite similar to crude oil; however, chemically it has some impurities which decrease its fuel quality, such as O-heterocyclic and N-heterocyclic compounds. In the HTL process, to achieve the best fuel quality possible, carbon and hydrogen are aimed to be maximized, whereas oxygen and nitrogen are to be minimized. The typical elemental composition for bio-crude is 60-78% carbon, 6-12% hydrogen, 2-6% nitrogen, 0-2% sulfur and 8-20% oxygen. There are five major substrate groups that were extensively studied as an HTL substrate in the literature: microalgae, macroalgae, cyanobacteria, lignocellulosic biomass, and municipal sludge. When the elemental compositions of bio-crudes produced from these groups in the literature are evaluated in terms of carbon and hydrogen, it can be seen that they all have similar characteristics, which makes them valuable fuels with high HHV. According to nitrogen (N) content, cyanobacteria bio-crudes visibly have a higher N profile, and lignocellulosic bio-crudes have a lower N profile. For oxygen, on the other hand, lignocellulosic bio-crudes have significantly high levels compared to other bio-crudes. Those nitrogen and oxygen content variations are directly related to the biochemical and elemental compositions of substrates. Most of the impurities in bio-crude caused by N and O are due to the heteroatoms being formed during HTL [63]. Although sulfur contents were found to be low for all bio-crudes, municipal sludge bio-crudes were found to have higher levels and large standard deviation due to the varying characteristics of feedstock (substrate) in the studies. This sulfur content may cause catalyst fouling in downstream processes, such as hydrodeoxygenation bio-crude upgrading [64]. The average HHV values indicate that microalgae is the most suitable HTL substrate in terms of bio-crude quality [9]. In contrast, lignocellulosic bio-crudes were found to have a vast diversity of HHV and, on average, the lowest fuel quality. Macroalgae, cyanobacteria, and municipal sludge bio-crudes were reported to have similar HHV. However, difficulties in growing and harvesting of microalgae and cyanobacteria [65] along with macroalgae being a food source 19 make municipal sludge the best substrate for HTL. This is the main driver for the increasing research and application of municipal sludge HTL in recent years. In the literature, pyrolysis oil is also mentioned as bio-crude in some studies. Pyrolysis oil is another thermochemically produced renewable fuel that can replace fossil fuels [66]. However, compared with HTL bio-crude, pyrolysis oil has higher water content (15-30%), higher oxygen percentage, and ignition complications [67]. Typical properties of pyrolysis oil and HTL bio-crude are given in Table 2.1. Table 2.1 Fuel quality comparison of HTL bio-crude and pyrolysis oil. Parameters HTL bio-crude Pyrolysis oil [68,69] Drying Not required Required C% 46.1-84.2 32-49 H% 2.7-14.2 6.9-8.6 O% 1.7-40.2 44-60 N% 0-10.5 0-0.2 S% 0-5.1 0-0.05 HHV (MJ/kg) 16.5-42.5 16-19 After the HTL process, bio-crude can be separated either by physical methods or solvents. The most utilized method in the literature (mostly in batch-fed HTL studies) is to dissolve bio-crude in dichloromethane (DCM), separate it by centrifugation, and then evaporate the solvent for reuse. Bio-crude that is dissolved and separated from the liquid fraction, and hydrochar are called light oil and heavy oil, respectively. Valdez et al. investigated the solubility of bio-crude by different polar and non-polar solvents [70]. They reported that the highest fraction of bio-crude dissolved was 39% by hexane or decane solvents, whereas it was only 30% for DCM. However, DCM solved bio-crude was found to have the highest carbon and hydrogen among seven solvents. Overall, decane was found to be the best solvent to separate bio-crude. A similar study was done by Yan et al. (2016) to investigate the efficiency of ten different solvents on bio-crude extraction [71]. According to their findings, isopropanol was slightly more successful than DCM in extraction. Xu and Savage (2014) studied the effect of centrifugation and solvent extraction order on bio-crude recovery [73]. They found that 43.7% bio-crude can be extracted if DCM is directly added to HTL effluent. However, a higher yield of 44.2% was reached by centrifuging the effluent first and then using solvent on both liquid and solid fractions separately. Mujahid et al. (2020) also reported that the centrifugation-first approach increased the bio-crude extraction yield from 47% to 64% [74]. 2.2.2 Hydrochar Hydrochar is the solid co-product in HTL, and it is the main product of hydrothermal carbonization (HTC) process, which typically utilizes lower temperatures (150-280°C). Hydrochar 20 in HTL process, also called solid residue, is formed by the recombination reactions of compounds in the aqueous phase and bio-crude. Along with recombination products, hydrochar also contains insoluble inorganics. Similar to bio-crude, hydrochar characteristics vary depending on the substrate and HTL conditions, and can consist of up to 80% inorganics [75]. Since bio-crude and hydrochar are formed from the same carbon sources, their yields are inversely proportional. According to the studies in the literature, hydrochar can be directly combusted for energy production, applied to soil for amendment and carbon sequestration, or activated for adsorption purposes [76]. Hydrochar also contains a decent amount of trace elements and phosphorus, which can be extracted for commercial use [77,78]. The valorization and/or safe disposal of hydrochar is one of the requirements to facilitate the sustainable development of HTL. 2.2.3 Hydrothermal liquefaction gaseous phase Gases in the HTL process are mainly produced during decarboxylation, decarbonylation, and deamination reactions. At reaction temperatures over 350°C and 340°C, respectively, water-gas shift [79,80] and methane forming reactions [44,81] also occur and produce H2 and CH4. The gaseous phase contains mainly CO2, which is around 80-95% [82,83]. Typical HTL gas composition is given in Figure 2.9. The CO2 content of the gaseous phase can be as high as 97%. Zhang et al. (2012) suggested that HTL gaseous phase can be used in algae growth due to the high CO2 content. Figure 2.9. Typical composition of HTL gaseous phase at 350°C [9]. HTL gaseous phase can also include small quantities of light hydrocarbons, such as ethane, ethylene, propene [85], isobutane, n-butane, 1-butane, isopentane, and 1-pentene [86]. Although these gases are generally in low concentrations, Wagner et al. [85] reported that in HTL of a polyhydroxy butyrate producing cyanobacteria, Synechocystis cf. salina, it is possible to generate propylene gas with a 2.6% yield along with bio-crude. Other14.35%C2-C5 : 4.18%N2 : 0.06%CO : 3%CO2 :85.65% CH4 : 1.49%H2 : 5.62% 21 2.2.4 HTL aqueous HTL aqueous stream contains both water soluble organic and inorganic residue compounds of the HTL process. Major organic compound groups in aqueous phase are carboxylic acids, alcohols, ketones, glycerol, aldehydes, phenolics, esters, ethers, amides, pyrazines, pyridines, and N&O-heterocyclic compounds [87â89]. Percentages of organic compound groups in HTL aqueous phases obtained from different types of substrates are presented in Figure 2.10. Figure 2.10. Distribution of organic compounds in HTL aqueous phase of various biomass [9]. According to Figure 2.10, the most abundant organic compound group in the HTL aqueous phase is carboxylic acids, except for HTL aqueous of cyanobacteria. The amount of carboxylic acids in HTL aqueous is directly proportional to the carbohydrate percentage of the biomass type due to HTL reaction pathways presented in Figure 2.1. Therefore, lignocellulosic biomass HTL aqueous was found to have the highest carboxylic acids due to its carbohydrate-rich cellulose and hemicellulose structure. For the same reason, HTL aqueous of macroalgae and some microalgae species also have high amounts of carboxylic acids. Cyanobacteria-based HTL aqueous, on the other hand, was found to be rich in amides as a result of the HTL reactions of lipids and proteins in cyanobacteriaâs structure. HTL aqueous of municipal sludge has the most balanced distribution with carboxylic acids, ketones, amines, and amides as major compound groups. Substrates with high protein content cause an abundance of amides in the aqueous phase. A high amide amount in the aqueous phase indicates that the substrate can be co-liquefied with a carbohydrate-rich substrate to promote the Maillard reactions. Likewise, a substrate with a high carboxylic acid concentration in the aqueous phase can be co-liquefied with a protein-rich substrate to obtain a higher bio-crude yield. CarboxylicAcidsAlcoholsKetonesPhenolsAminesAmidesN-HeterocyclicOtherCompounds0%20%40%60%80%100%Organic Compound Groups Lignocellulosic Municipal Sludge Microalgae Macroalgae Cyanobacteria 22 Due to deamination reactions of proteins, ammonia concentration is also high in the aqueous phase. Like hydrochar, HTL aqueous is also a suitable medium for nutrient recovery. Ovsyannikova et al. (2020) reported that they reached 99% P, 79.4% NH4-N, and 66.5% P, 19.4% NH4-N recovery by precipitation from HTL aqueous of municipal sludge and Spirulina (cyanobacteria), respectively [90]. On the other hand, Garcia Alba et al. (2013) investigated Desmodesmus sp. growth in HTL aqueous phase and reported growth reduction in HTL aqueous [91]. Chen et al. (2020) also found that even in 100 times diluted HTL aqueous, Chlorella vulgaris growth was reduced by 47% due to toxicity [92]. In the study by Das et al. (2020), they found that it is possible to grow Tetraselmis sp. in HTL aqueous if it is diluted enough to provide 50% of nutrients [93]. Anaerobic digestion is another valorization method for the HTL aqueous stream. However, according to the study by Zhou et al. (2015), HTL aqueous (104 g chemical oxygen demand (COD)/L) inhibits the batch anaerobic digestion process in mesophilic conditions (37°C) when its volume exceeds 13.3% of the reactor liquid volume [94]. Chen et al. (2020) reported that in up-flow anaerobic sludge blanket (UASB) systems that were being fed with HTL aqueous (2.5 g COD/L d), 61.6% and 45% of the COD contributed by HTL aqueous (10 g COD/L) were digested under mesophilic and thermophilic conditions, respectively [95]. They also emphasized that it was not possible to digest phenolic compounds, furans, and pyrazines in UASB system. On the other hand, it is known that digesting low concentrations of those compounds in batch anaerobic reactors is possible [96,97]. Elliott et al. (2013, 2014, 2017) suggested catalytic hydrothermal gasification (CHG) treatment for converting organic compounds in HTL aqueous into combustible gasses [98â100]. Xu et al. (2019) also followed the same approach and found that some N-containing compounds are difficult to be gasified [101]. According to the study by Cherad et al. (2016), it is possible to produce sufficient H2 gas from CHG of HTL aqueous for hydrodeoxygenation fuel upgrading of bio-crude, as one of the downstream processes [102]. Recycling HTL aqueous back to HTL process or headworks of a WWTP is another option. Ramos-Tercero et al. (2015), Biller et al. (2016b), and Shah et al. (2020) found that recycling aqueous phase up to six times in HTL increases energy recovery of the process [89,103,104]. However, Shah et al. (2020) also stated that the nitrogen content of bio-crude was increased by 55% after the sixth recycle, which is not desirable by end users [104]. Treatment of HTL aqueous with a microbial fuel cell is another promising topic of research interest. Studies investigating the treatment or valorization of HTL aqueous in the literature are summarized in Table 2.2. 23 Table 2.2. Hydrothermal liquefaction aqueous phase valorization and treatment studies in the literature. HTL Substrate Temperature (°C) Retention Time (min.) Treatment Method of HTL Aqueous Results Reference Municipal sludge 350 60 Biochar adsorption Biochar was able to adsorb 41.8% of phenolic compounds in the HTL aqueous. [105] Primary sludge 325 28 Wet oxidation Wet oxidation at 350°C and 180 minutes successfully treated organic pollution. [106] Municipal sludge 320 30 Anaerobic fermentation (mesophilic) Volatile fatty acid (VFA) production of 0.34 g CODVFA/g COD was achieved from HTL aqueous. [107] Corn stalk 300 60 Anaerobic digestion (mesophilic and thermophilic) 194 mL CH4/g COD and 137 mL CH4/g COD methane productions were observed in mesophilic and thermophilic conditions, respectively. [108] Distillerâs grains 340 16 Aqueous phase recycling Recycling the aqueous phase in the HTL increased the energy recovery to 95%. However, oxygen levels in bio-crude also increased each time aqueous phase was recycled. [89] Model compounds 320 30 Valorize HTL aqueous as a bactericide HTL aqueous was found to have equal germicidal effect with 100 µg/mL streptomycin. [109] Municipal sludge 300 60 Anaerobic fermentation with granular activated carbon (mesophilic) Granular activated carbon increased the acetoclastic methanogenesis and hydrogenotrophic methanogenesis by 60% and 50%, respectively. [110] Chlorella 300 30 Zeolite adsorption and anaerobic fermentation (mesophilic) Raw HTL aqueous inhibited anaerobic digestion when it was over 20% of the feed. Zeolite-treated HTL aqueous was anaerobically digested without inhibition up to 60% HTL aqueous to feed ratio. [111] 24 Table 2.2. Hydrothermal liquefaction aqueous phase valorization and treatment studies in the literature (continued). HTL Substrate Temperature (°C) Retention Time (min.) Treatment Method of HTL Aqueous Results Reference Desmodesmus sp. 300 5 Grow algae in HTL aqueous Nutrients in HTL aqueous were successfully used in cultivation of Desmodesmus sp. by 20 times dilution and adding more nutrients. [91] Swine manure 270 60 Anaerobic fermentation with biochar and granular activated carbon (mesophilic) Both additives increased the methane production in batch tests. Increase was due to the adsorption of inhibitors. No proof was found for enhanced direct interspecies electron transfer (DIET) mechanism. [112] Algae and cyanobacteria 300-350 60 Grow algae and cyanobacteria in HTL aqueous Each microalga and cyanobacteria species were successfully grown in HTL aqueous in varying dilution ratios. [113] Swine manure - - Anaerobic digestion (mesophilic) HTL aqueous was digested without a problem under 6.7% concentration. Over 13.3%, lag phase started to increase, and significant inhibition was observed in methanogenesis. [94] Chlorella vulgaris - - Ion exchange resin treatment for algae growth Algae growing in raw HTL aqueous was found infeasible. Even the most tolerant algae species was inhibited by HTL aqueous. Resin treatment of HTL aqueous significantly improved the algae growth yield. [114] Saccharina spp. 350 70 Catalytic hydrothermal gasification COD of HTL aqueous was treated with 99.2% efficiency in catalytic hydrothermal gasification. Produced gas had 69% methane percentage. [99] Municipal sludge 350 30 Catalytic wet air oxidation Over a period of 200 hours of oxidation with the use of a CeZr catalyst, a reduction in COD by 50% and an increase in acetic acid concentration by 40% were achieved. [115] Wheat stem 350 15 Adsorption with activated hydrochar Total organic carbon removal of 91.8% and total phenolic compound removal of 99.9% was achieved. [116] 25 In terms of widely available and sustainable feedstock and bio-crude yield, one of the best options for the HTL process is to use municipal sludge as substrate by coupling the HTL process with a WWTP. In this configuration, HTL aqueous can be treated by returning it to the headworks of the WWTP. However, the exact effects of inhibitory aromatic compounds and high concentration of ammonia in the HTL aqueous on wastewater treatment processes are still unclear. Although carboxylic acid concentration can be tolerated due to the dilution of HTL aqueous with incoming wastewater, high nitrogen concentration of HTL aqueous can affect the biological treatment processes applied in the mainstream by bringing an extra load to the nutrient removal process. If the nutrient removal capacity of the biological treatment process is not sufficient for treating that extra load, nitrogen accumulation may occur in the plant. Zhuang et al. (2017) reported that in the HTL aqueous of municipal sludge, organic nitrogen is converted into ammonium (NH4+) by increasing the temperature, whereas no significant change was observed on nitrite (NO2-) and nitrate (NO3-) concentrations [117]. In their study, the ammonium concentration in the aqueous phase was found to be 2500 mg/L at HTL condition of 300°C and 30 minutes. Xu et al. (2018) also reported a 2210 mg/L total ammonia concentration in HTL aqueous phase of municipal sludge processed at 340°C [18]. Ammonia is mostly formed in the deamination reactions of amino acids. Therefore, compared to the HTL aqueous of primary sludge, HTL aqueous of secondary (biological) sludge has a higher total ammonia concentration due to its higher protein content. In a continuous-flow HTL system operating at 350°C, Maddi et al. (2017) found 3680 mg/L and 5700 mg/L total ammonia concentrations in HTL aqueous of primary sludge and secondary sludge, respectively [118]. Besides the aerobic or anaerobic treatability of HTL aqueous, its possible effects on settling and UV disinfection units should also be considered for the option of returning it to headworks of WWTP. In the literature, there are no studies investigating the effect of blending HTL aqueous to wastewater on settling and UV disinfection processes. As mentioned above, HTL aqueous contains a wide variety of organic and inorganic molecules. Many of the organic compounds in HTL aqueous were reported to be inhibitors for anaerobic consortium [94], microalgae [119,120], or even mammal cells [121]. These inhibitors might also change the microbiological structure of wastewater and cause a decrease in biomass settling efficiency. Furthermore, an increase in the concentration of some compounds, such as phenol [122] or melanoidins, which has a high adsorption capacity at 254 nm, can lower the UV disinfection process efficiency. 26 2.3 Impact of hydrothermal liquefaction conditions and feedstock on products Product yield and quality vary in HTL processes, mainly depending on the reaction temperature, retention time, type of biomass, solid content, catalyst type, and catalyst concentration. This product variation is caused by the effect of reaction contents and conditions on the complex kinetics of HTL reactions presented in Section 2.1. In order to evaluate the parameter effect on HTL, studies in the literature using different reaction temperatures, retention times, substrate types, and solid contents are summarized in Table 2.3. Table 2.3. Overview of reaction conditions and product yields in batch hydrothermal liquefaction research. HTL contents HTL conditions Bio-crude yield (%wt) Reference Substrate Substrate to water ratio (w:w) Reaction temperature Retention time L. digitata L. hyperborea L. saccharina A. esculenta 8:30 350°C 15 min 17.6% 9.8% 13% 17.8% [123] Dried sewage sludge 1:2 250-400°C 60 min 52% at 350°C [124] Food waste 3:20 240-295°C 0-30 min 27.5% at 240°C, 30 min [125] C. vulgaris S. dimorphous S. platensis C. fritschii 1:10 300, 350°C 60 min 46.6% at 300°C 35.8% at 350°C 35.5% at 300°C 38.6% at 300°C [113] Nannochloropsis sp. 1:20 200-500°C 60 min 43% at 350°C [126] Eight types of green landscaping waste 1:10 300°C 30 min Leaves: 41% for D. racemosum Branches: 51% for C. camphora [127] Microwave pretreated municipal sludge 3:20 340°C 30 min 35.4% [128] Mixed-culture algal biomass 1:4 260-320°C 0-90 min 49.9% at 300°C, 60 min [75] N. salina G. sulphuraria 1:20-1:10 310-350°C 5-60 min 59.1% at 310°C, 60 min, 1:20 ratio 31% at 350°C, 30 min, 1:20 ratio [129] G. sulphuraria 5587.1 G. sulphuraria SOOS 1:20-1:10 310-350°C 5-60 min 18.2% at 330°C, 60 min, 1:20 ratio 23% at 350°C, 5 min, 1:20 ratio [130] Arthrospira platensis rsemsu 1/02-P 3:10 270-330°C 0 min 45.7% at 330°C [131] Mixed culture of C. sorokiniana, C. minutissima, S. bijuga 7:32 350°C 60 min 26.7% [132] Tetraselmis sp. 3:17 325, 350°C 30-60 min 51.2% at 350°C, 30 min [93] Natural hay, oak wood, walnut shell, cellulose 1:5 240, 320°C 30 min 25% for walnut shell [133] Spirulina sp. Tetraselmis sp. 4:25 300-350°C 5 min 42% at 350°C 58% at 350°C [134] Desmodesmus sp. 3:40 300°C 5 min 40% [91] Desmodesmus sp. 2:25 175-450°C 5-60 min 49.4% at 375°C, 5 min [135] Mixture of C. sorokiniana DBWC2 Chlorella sp. DBWC7 K. pneumoniae ORWB1 A. calcoaceticus ORWB3 3:20 310°C 55 min 15% [136] Chlorella sp. FC2 1:10 350°C 60 min 14.72% [137] Arthrospira platensis 3:10 270°C 80 min 34.6% [138] Cyanophyta 1:20-3:10 320-370°C 10-70 min 39.54% at 370°C, 50 min, 1:10 ratio [139] Oil Mill Wastewater 3:25 240-300°C 15-45 min 58% at 280°C, 30 min [140] 27 Table 2.3. Overview of reaction conditions and product yields in batch hydrothermal liquefaction research (continued). HTL contents HTL conditions Bio-crude yield (%wt) Reference Substrate Substrate to water ratio (w:w) Reaction temperature Retention time Nannochloropsis sp. Sargassum sp. 3:20 260-340°C 30 min 54.11% at 320°C 9.49%. at 340°C [141] Raw empty fruit bunch Palm mesocarp fiber Palm kernel shell 1:10 330-390°C 30-240 min 38.24% at 390°C, 120 min 34.32% at 390°C, 60 min 38.53% at 390°C, 60 min [142] Nannochloropsis sp. 3:20 100-400°C 0.16-60 min Model prediction: 46% at 230°C [17] Chlamydomonas reinhardtii 7:50 220-310°C 60 min 58.7% at 310°C [143] Cyanobacteria sp., Bacillariophyta sp. 1:10 250-350°C 15-90 min 21.10% at 325°C, 45 min 18.21% at 325°C, 60 min [144] Spirulina platensis 1:5 350°C 60 min 40.7% [120] Animal by-products 1:1-1:5 150â290°C 5-15 min 74% at 250°C [145] Nannochloropsis sp., Chlorella sp. 3:20-1:4 220â300°C 30-90 min 55% at 260°C, 60 min, 1:4 ratio 82.9% at 220°C, 90 min, 1:4 ratio [146] Manure from six different animal species 1:5 310â340°C 30 min 30.85% at 340°C for swine manure [147] N.gaditana S.almeriensis 1:10 350°C 15 min 60% 57.7% [148] Spirulina platensis Miscanthus Primary municipal sludge 1:20-1:4 250-350°C 5-31 min Model predictions: 40% at 300â320°C, 16â18 min, 1:5 46.9% at 330°C, 10 min, 3:20 42.3% at 320°C, 18 min, 3:20 [149] α-cellulose Xylan Soy protein Sunflower oil Starch Lignin Food waste Garden waste Paper waste 1:1 260-300°C 20 min 3% at 300°C 97% at 300°C 22% at 300°C 95% at 260°C 3% at 300°C 34% at 300°C 79% at 280°C 33% at 300°C 74% at 280°C [150] Chlorella sorokiniana 1:12-1:3 Direct HTL: 220-300°C Sequential HTL: 220-320°C 5-60 min 30% at 240°C, 20 min, 1:9 ratio [151] Domestic sewage sludge (DSS) Monoraphidium sp. (KMC4) Mixture of KMC4 to DSS (3:1 ratio) 1:10 275-350°C 15-60 min 22% at 350°C, 30 min 33% at 325°C, 30 min 39.6% at 325°C, 45 min [152] Dewatered sewage sludge 13:100 300-400°C 30-120 min 64% at 325°C, 30 min [74] Four species of marine and two species of freshwater microalgae 1:15 330-341°C 5 min Freshwater species: 26.2% for Oedogonium Marine algae species: 19.7% for Derbesia [153] Lipid Carbohydrate Lignin Protein 3:10 250-350°C 0-60 min 100% at 250°C. 0 min 42.83% at 350°C, 0 min 37.27% at 300°C, 0 min 4.4% at 250°C, 0 min [154] Mixture of lipid, carbohydrate, lignin, and protein 3:10 250-350°C 5-60 min ~30% at 350°C, 30 min [155] Gracilaria gracilis Cladophora glomerata 1:10 350°C 15 min 15.7% 16.9% [156] Chlorella vulgaris 11:100 220-265°C 30 min 42.2% at 240°C for the sixth time recycling of HTL aqueous [103] Nannochloropsis gaditana Chlorella sp. 1:10 180-330°C 30 min 47.5% at 300°C 32.5% at 275°C [157] 1:1 Water hyacinth-Para grass mixture 1:6 260-300°C 15 min 13.34% at 280°C [158] 28 Table 2.3. Overview of reaction conditions and product yields in batch hydrothermal liquefaction research (continued). HTL contents HTL conditions Bio-crude yield (%wt) Reference Substrate Substrate to water ratio (w:w) Reaction temperature Retention time Lipid extraction residue of of Dunalliela sp. 4:25 350°C 60 min 11.81% [159] Rice straw 1:6 280-320°C 15 min 17% at 280°C [160] Spirulina Chlorella Enteromorpha prolifera Nannochloropsis sp. Dunaliella salina Euglena Cyanophyta 1:4 280-350°C 30 min 32.5% at 350°C 37.5% at 350°C 27% at 280°C 37% at 280°C 17% at 350°C 29% at 350°C 30.5% at 280°C [161] Nannochloropsis Spirulina 1:10 220-300°C 60 min 41% at 280°C 33.3% at 260°C [162] Mixed algae 1:10-1:5 260-340°C 30-90 min 8.4% at 300°C, 60 min, 1:5 ratio [163] Corn straw Peanut straw Soybean straw Rice straw 3:8 320°C 60 min 7.9% 14.6% 15.8% 15.1% [164] Nannochloropsis salina Spirulina platensis 1:3 220-375°C 10-30 min 46% at 310°C, 30 min 38% at 350°C, 30 min [165] Nannochloropsis sp. 1:20-7:20 250-400°C 10-90 min 50% at 300°C, 10-20 min [166] Spirulina Swine manure Digested anaerobic sludge 1:5 300°C 30 min 32.6% 30.2% 9.4% [167] Raw Scenedesmus, Defatted Scenedesmus Spirulina 1:5 300°C 30 min 45.4% 36.5% 31.2% [168] Arthrospira platensis 3:10 270-300°C 80 min 32.6% at 330°C [169] Amphora sp. Spirulina sp. Tetraselmis sp. 4:25 350°C 5 min 55% 50% 66% [170] Spirulina, Synechococcus/Anabaena Synechocystis Soybean Rapeseed oil Cornflour Polyhydroxy butyrate 4:25 300-360°C 0 min 36.9% at 340°C 31.2% at 340°C 16.6% at 340°C 28% at 360°C 100% at 320-360°C 3% at 340°C 8% at 300°C [85] Secondary sewage sludge 1:10 260-350°C 10 min 22.9% at 340°C [18] Nannochloropsis sp. 7:50 350°C 20 min 43.9% [72] Chlorella Sewage sludge Mixture (1:3,1:1,3:1) 1:10 340°C 30 min 22% 24% 26.8% for 1:1 mixing ratio [171] Duckweed (Lemna minor) 1:3 350°C 30 min 26% by isopropanol extraction 24% by dichloromethane extraction [71] Animal carcass 1:20-1:5 230-350°C 10-80 min 55.6% at 320°C, 60 min, 1:10 ratio [172] Spend coffee grounds 1:20-1:5 200-300°C 5-30 min 35.29% at 275°C, 10 min, 1:5 ratio [173] Chlorella pyrenoidosa 1:5 100-300°C 0-120 min 39.5% at 280°C, 120 min [174] Spirulina platensis Lemon peel Mixture (0-4:0-4) 1:10 280-350°C 10-60 min 19.74% at 315°C, 35 min 7.24% at 315°C, 35 min 26% at 336°C, 35 min, 4:1 blending [175] Enteromorpha prolifera 3:17 300°C 30 min 17.8% [176] Swine manure Sewage sludge 11 algae species 1:5 280°C 30 min 33% 32% 35% for Chlamydomonas [84] Dunaliella tertiolecta 1:40-7:40 300-380°C 10-90 min 36.9% at 360°C, 30 min, 1:10 ratio [177] Sargassum tenerrimum 1:6 260-300°C 15 min 16.33% at 280°C [178] 29 Table 2.3. Overview of reaction conditions and product yields in batch hydrothermal liquefaction research (continued). HTL contents HTL conditions Bio-crude yield (%wt) Reference Substrate Substrate to water ratio (w:w) Reaction temperature Retention time Macroalgal mix 1:5 305-350°C 0 min 16% at 345-350°C [179] Sargassum tenerrimum 1:6 260-300°C 15 min 16.33% at 280°C [180] Ulva prolifera 1:6 200-310°C 10-30 min 12% at 290°C, 10 min [181] Human feces 1:20-1:4 260-340°C 10-50 min 34.44% at 340°C, 10 min, 1:4 ratio [182] Corn stover 1:6 250-375°C 0-60 min 42.61% at 300°C, 0 min [183] According to the result summary of HTL studies performed without catalyst in Table 2.3, in varying HTL conditions, the average bio-crude yields for microalgae, macroalgae, cyanobacteria, lignocellulosic and municipal sludge were 35.3%, 17.5%, 28.7%, 26.9%, and 39.4%, respectively. The two lowest average bio-crude yields, which belong to lignocellulosic and macroalgae substrates, indicate that due to the HTL pathways of carbohydrate-rich biomasses, the application of an alkaline catalyst is crucial to reach higher bio-crude yields. In HTL of carbohydrate-rich biomasses, volatile fatty acids can accumulate and decrease pH. As indicated by Elliot et al. [184], acidic conditions in HTL favors hydrochar generation instead of bio-crude. In Table 2.3, it can be clearly seen that the optimum HTL conditions and bio-crude yield results are highly dependent on the substrate type. Among the most investigated five substrate types, municipal sludge studies reported the highest bio-crude yields. This trend is consistent with biochemical compositions and HTL reaction pathways, as discussed in Section 2.1. As one of the most significant advantages of the HTL process, there is no limitations or pretreatment requirement for the types of substrate that are processed. Regardless of the moisture content and structure of organic substrate, it is possible to produce bio-crude with diverse quality and quantity. In the literature, microalgae, macroalgae, cyanobacteria, lignocellulosic biomass, and municipal sludge are the most studied HTL substrates. There are also studies that investigated the HTL of manure, food waste, slaughterhouse waste, and organic fraction of municipal solid waste [125,145,147,167,172,185â188]. To compare the bio-crude quality obtained from the five most promising HTL substrates, a van Krevelen diagram was created from the reported bio-crude data and presented in Figure 2.11. When both bio-crude quality and quantity are evaluated, it can be seen that municipal sludge and microalgae are the most promising substrates. Figure 2.11 also reveals that, regardless of the substrate type, application of a bio-crude upgrading process can be necessary to lower the oxygen percentage of bio-crude and reach the fuel quality of crude oil. 30 Figure 2.11. van Krevelen diagram of bio-crudes produced from different substrates in the literature [9]. In the literature, comparative studies on HTL of cyanobacteria, lignocellulosic biomass, and municipal sludge were performed to determine the effect of substrate type on bio-crude yield. For HTL at 350°C and 20 min conditions, Huang et al. [189] found 39.46%, 34.51%, and 21.14% bio-crude yields for municipal sludge, Spirulina, and rice straw, respectively. In a similar study, Anastasakis et al. [14] reported yields of 32.9% for Spirulina, 26.2% for Miscanthus, and 24.5% for municipal sludge in a continuous-flow HTL system. Furthermore, by using the same substrates, Madsen et al. [149] found 40%, 46.9%, and 42.3% bio-crude yields for Spirulina, Miscanthus, and municipal sludge, respectively, by optimized process conditions using a statistical method. 2.4 Use of catalyst in the hydrothermal liquefaction process The effect of catalyst varies in the HTL process, depending on the type of catalyst and substrate. Catalytic effect is achieved by altering the complex HTL reaction pathways or kinetics. However, due to the complexity of the process, catalytic effect can also be negative in terms of bio-crude yield. Therefore, catalyst type and concentration, which are specific for the substrate type, need to be selected carefully to enhance bio-crude yield and quality. To determine the substrate-specific catalyst performances, the overall effects of catalysts used by the studies in the literature are presented in Figure 2.12. 0.000.040.080.120.160.200.0 0.1 0.2 0.3 0.4 0.5 0.6 0.7Data Points Microalgae Macroalgae Cyanobacteria Lignocellulosic Municipal Sludge Crude OilHydrogen/CarbonOxygen/CarbonDehydrationDistribution Curves Microalgae Macroalgae Cyanobacteria Lignocellulosic Municipal SludgeDecarboxylation 31 Figure 2.12. Overall catalyst performance on different hydrothermal liquefaction substrates [9]. According to the results presented in Figure 2.12, it can be seen that lignocellulosic biomass is the most suitable biomass type for a catalytic HTL process. A high bio-crude yield enhancement for lignocellulosic biomass is reported for all five catalyst types, especially for alkaline and mineral catalysts. Due to the carbohydrate-rich structure of lignocellulosic biomass, the majority of the HTL reactions that can be catalyzed are based on carbohydrates. In uncatalyzed HTL of lignocellulosic biomass, carbohydrates are converted into organic acids. The accumulation of organic acids causes HTL reactions to favor hydrochar formation due to low pH. However, when the process is catalyzed, this pH change is avoided either by neutralizing pH or enhance degradation of organic acids. Although low pH during the HTL process is not favored, adding an acid catalyst in low concentrations initially can enhance the hydrolysis step and increase bio-crude yield for lignocellulosic biomass. For the substrates that are balanced in terms of carbohydrate and protein, catalysts are less effective compared to lignocellulosic biomass. Except for metal-based catalysts, mostly negative results were reported for microalgae, macroalgae, cyanobacteria, and municipal sludge. Particularly acid catalysts caused a significant loss in bio-crude yield for these substrates. Some promising results were reported for HTL of microalgae and cyanobacteria with a metal-based catalyst. The development of a reusable heterogeneous metal-based HTL catalyst can positively AlkalineCatalystsAlkali saltCatalystsAcidCatalystsMetal-basedCatalystsMineralCatalysts-60-50-40-30-20-100102030405060708090100110Change in Bio-crude Yield (%) Microalgae Macroalgae Cyanobacteria Lignocellulosic Municipal Sludge 32 affect the process's energy balance by decreasing the required reaction temperature while increasing the bio-crude yield. Therefore, it can be said that catalyst development will be an important step in the transition to full-scale applications. 2.5 Techno-economic analysis of the hydrothermal liquefaction process The major operating expenses in a full-scale HTL process are caused by substrate production or supply, dewatering, pumping, heating, phase separation, chemicals (solvents, catalysts), and waste management for product streams (aqueous and hydrochar). Among the major expenses, if algae or agricultural sources biomass is used, substrate supply would constitute the majority of the total cost [190]. Hence, the most economical option is to couple HTL plants with waste-producing facilities, such as wastewater treatment plants. As an important advantage among biomass valorization methods, HTL does not require a biomass dying process, which is an energy sink [27]. The bottleneck of the HTL process is the energy consumption for heating the substrate to the desired temperature. However, by incorporating a heat exchanger and efficient heat insulation, this energy consumption can be lowered. Furthermore, the application of the right catalyst to the HTL process can decrease the process temperature and input energy costs significantly. Energy consumption points and amounts of a continuous-flow HTL system are given in Figure 2.13. Calculations were done by assuming a 15% solid content of substrates and 50% dry ash-free (daf) bio-crude yield in the HTL process. Figure 2.13. Summary of economic analysis in the hydrothermal liquefaction process (daf: dry ash-free). In the literature, production costs of microalgae/cyanobacteria, macroalgae and lignocellulose were reported as US$ 1253 /ton (daf) [191], US$ 757 /ton (daf) [192] and US$ 96 33 /ton (daf) [193], respectively. Since it is a waste, and it is assumed that the HTL plant will be coupled with a wastewater treatment plant, the cost of municipal sludge production was assumed as US$ 0 /ton (daf). Since the substrate cost is the main cost in the total operational cost, municipal sludge-based bio-crude was found to be cheaper than the other bio-crudes. For the substrate of municipal sludge, microalgae or cyanobacteria, macroalgae, and lignocellulose, the cost of producing a liter of upgraded bio-crude by HTL was found to be US$ 0.74, US$ 3.24, US$ 2.25, and US$ 0.93, respectively. Results indicate that elimination or reduction in substrate costs plays a major role in the final product's price. Funkenbusch et al. [194] investigated coupling an HTL plant with a Kraft pulping process and found that Kraft lignin-based bio-crude can be produced for US$ 0.41-0.43 /L if the hydrodeoxygenation process for fuel upgrading is performed on-site. Zhu et al. [195] reported a US$ 1.29 /L bio-crude cost produced from lignocellulosic sources, whereas Ramirez et al.[196] found the minimum selling price of wood-based bio-crude as US$ 1.17 /L. Also, Perkins et al. [197] underlined that with current production costs, the minimum bio-crude selling price should be US$ 0.92-1.10 /L. These cost calculations do not include valorization or treatment of HTL aqueous, hydrochar, and gases. Among those co-products, HTL aqueous has the potential to affect bio-crude costs considerably. One of the most efficient ways of valorizing HTL aqueous is to use it in anaerobic hydrogen or methane production due to presence of soluble organics. Although producing hydrogen seems to be the right approach to decrease the cost of hydrodeoxygenation fuel upgrading process, HTL aqueous may not be an ideal substrate for biochemical hydrogen or methane production due to its high concentrations of inhibitory and recalcitrant compounds [110,198]. Si et al. [199] found that in an up-flow anaerobic sludge blanket reactor, by the addition of 8 g COD/L equivalent HTL aqueous, biohydrogen fermentation and subsequent methane fermentation were inhibited 100% and 50%, respectively. Another alternative is to dilute HTL aqueous and utilize it in an anaerobic digestion process. Produced methane from this process can be converted to electricity and heat on-site and used in HTL or hydrodeoxygenation processes. Coupling an HTL plant with an anaerobic digestion process was also found beneficial in terms of energy recovery [200,201]. However, adding an anaerobic process for hydrogen or methane production increases capital costs. Davidson et al. [202] calculated that if co-products are processed on-site, waste disposal would only increase the bio-crude cost by US$ 4 /L. If the HTL aqueous is used in a treatment process instead of valorization, capital investments required for condensed phase ketonization, dual-bed steam reforming, and catalytic hydrothermal gasification processes were reported as US$ 34.1M, US$ 34.1M [202] and US$ 16.8M [203], respectively. 34 The handling cost of the HTL aqueous phase from an HTL utilizing lignocellulosic biomass with low moisture would be lower since it can be recirculated in the HTL process. Knorr et al. [204] also stated that direct aqueous phase recirculation increases heat recovery significantly. They found that an HTL aqueous recirculating plant has the lowest operating cost at US$ 22.5M/year. However, it was also reported in the literature that over six times of recirculation of the HTL aqueous phase could lower bio-crude yield and quality [89,103,205]. Furthermore, recirculation causes the accumulation of inhibitors and recalcitrant substances, which increases the treatment cost of the aqueous phase. If the substrate is municipal sludge in the HTL process, the aqueous phase might be returned to the treatment plant to not to add an extra treatment cost to the process. However, the effect of the inhibitory compounds in the aqueous phase, mainly nitrogen-containing organics and aromatic compounds [206], on wastewater treatment processes is unknown. HTL gas and hydrochar are the other co-products that need handling. Hydrochar can be landfilled, combusted, or valorized by the application of further processes. Generally, hydrochar amount produced in the HTL process is so low (<20% on a dry basis) that its effect on total process cost is negligible. The only concerns about hydrochar are its heavy metal and persistent organic pollutant concentrations. However, Li et al. [147] reported that HTL reduced the high environmental risk due to the immobilization of heavy metals in the raw substrate (animal manures) to medium for hydrochar. Therefore, it is even possible to use hydrochar in soil amendment safely. Some studies also showed that HTL can destroy most persistent organic pollutants present in the substrate [207]. On the other hand, since the major fraction of HTL gas is CO2 and it does not contain any dangerous compounds, it can be released into the atmosphere. However, to make the HTL process more environmentally friendly, greenhouse gases, including CO2, can be captured by adsorption [208]. The addition of the gas adsorption process would cause an extra operating cost of US$ 60-190/ton of CO2 captured [209]. 2.6 Summary of findings from the literature review Key findings from the literature review section are given below: â¢ The literature on HTL of municipal sludge is limited, yet it is a highly promising feedstock because it is a waste stream, unlike other substrates that are specifically cultivated by costly processes for biofuel production. â¢ Municipal sludge is an ideal feedstock for the HTL process due to its optimal carbohydrate to protein ratio, which can be adjusted by the proportion of PS to SS ratio in the mix. 35 â¢ Compared to other feedstocks, except for lignocellulosic ones, municipal sludge generates less amount of inhibitory compounds in HTL aqueous phase, such as N-heterocycles. â¢ While there have been attempts to identify a treatment method for HTL aqueous, the literature does not present any cost effective, applicable techniques suitable for use in a WWTP for treating municipal sludge derived HTL aqueous. â¢ Among various feedstocks, municipal sludge generates one of the highest quality bio-crudes, which is crucial for minimizing the amount of organic material left in the aqueous phase. â¢ Integrating a HTL unit into a WWTP offers the advantage of on-site processing, potentially enabling bio-crude production at a low cost of approximately US$ 0.74/L, while allowing the on-site treatment of HTL aqueous produced in the process. â¢ No catalysts have been identified to enhance the HTL process of municipal sludge. â¢ Overall, HTL of municipal sludge appears to be a promising method for transforming WWTPs into biorefineries, provided the challenge of HTL aqueous treatment can be addressed. 36 Chapter 3: Characterization of Municipal Sludge-Derived Hydrothermal Liquefaction Aqueous Phase Under Various Reaction Conditions 3.1 Introduction In the HTL process, water acts as the reaction media. At temperatures over 300ºC, water has stronger solvent properties on both polar and nonpolar compounds than commonly known solvents, such as acetone, ethanol, or methanol [9]. This enhanced solvent property in the reaction media allows hydrolysis of biomass and thermochemical decomposition to take place in short amounts of time (5-30 min). Decomposed molecules then recombine and form bio-crude along with the large molecules that donât decompose under HTL conditions. However, many of the small molecules stay dissolved in water after the HTL reaction and cause a highly polluted liquid effluent named HTL aqueous. Municipal sludge-derived HTL aqueous phase contains high concentrations of VFAs (10-17 g/L), total ammonia (3.9-7.1 g/L), dissolved proteins (10-20 g/L), total phenols (1.5-2.3 g/L), and N-heterocyclic compounds. During HTL reaction, one kilogram of municipal sludge with 20 wt.% solids produce 0.81 L of HTL aqueous that has approximately 100 g COD/L. Due to its high volume and COD, the treatment of HTL aqueous is an issue that must be solved before a full-scale application of HTL in WWTPs. There have been investigations on HTL aqueous treatment and valorization by using adsorption [112,210], wet oxidation [106], mesophilic anaerobic digestion [94,107,112,210], thermophilic anaerobic digestion [95], recycling aqueous in the HTL reactor [89,103,205], algae growing [91,113], ion exchange resin treatment [92], and catalytic hydrothermal gasification [98,99]. However, due to the complex and inhibitory nature of HTL aqueous, an economical and practical processing method that can be applied at full-scale was not possible to determine. It is clear that identifying the ultimate HTL aqueous treatment or valorization method will require an extensive number of studies. In this chapter, the goal was to reveal the characteristics of HTL aqueous that would be produced in a municipal WWTP by processing mixed sludge under various HTL conditions. The characterization results were mathematically related to the HTL reaction conditions to identify sweet spots or HTL conditions to avoid undesired aqueous characteristics for specific downstream treatments. The ultimate objective of the study is to assist in determining a suitable HTL aqueous treatment method that will eliminate the biggest bottleneck of the HTL process and allow the incorporation of the process into wastewater treatment with its numerous environmental benefits. 37 3.2 Materials and methods 3.2.1 Sample preparation All experiments were conducted at the Bioreactor Technology Group (BTG) laboratory at UBCâs Okanagan Campus. Thickened screened primary sludge (TSPS) and thickened waste secondary sludge (TWSS) were obtained separately from Annacis Island WWTP in British Columbia, Canada. The daily volume ratio of primary to secondary sludge typically changes from 40:60 to 60:40% at WWTPs, depending on the scale and unit processes (sludge age, bioreactor configuration) implemented. To make the results representative of a typical WWTP, TSPS and TWSS were mixed at a 50:50 wt.% ratio. The obtained mixed sludge was centrifuged at 4000Ãg for 5 minutes and then at 20000Ãg for 10 minutes to achieve a concentration of 20 wt.% total solids (TS) before the HTL runs and was stored at 4°C. HTL runs were performed simultaneously with dewatering to decrease the storing time of mixed sludge. 3.2.2 Experimental design Two independent variables were used in the experimental design, HTL reaction temperature and retention time. The ranges for these independent variables, 290-360°C and 0-30 min, and the center point, 325°C - 15 min, were determined according to extensive literature review [9,30]. The zero-minute retention time represents starting the cooling right after the design HTL temperature was reached. A circumscribed central composite design (CCC) was used to determine the eight experimental points around the central point. The CCC design, which includes points at all axial positions that are at equal radial distances from the center, allows for an efficient investigation of the interaction and quadratic effects of the independent variables, namely HTL reaction temperature and retention time. A circumscribed design was preferred over a face-centered design because the CCC design places additional emphasis on the region around the center point, enabling more precise predictions within this critical area of the experimental space. The experimental design is illustrated in Figure 3.1. HTL runs for the eight experimental points were performed in duplicate and the center point was performed in quadruplicate. 38 Figure 3.1. Experimental design for HTL operating conditions for dewatered municipal mixed sludge (Severity is defined by Eq. 3.1 in Section 3.2.3). 3.2.3 Bench-scale HTL operation The HTL runs were performed using a Parr 4570 high temperature â high pressure (HT/HP) batch reactor (Parr Instrument Company, USA) with a maximum operating temperature of 500°C. The HT/HP reactor was made of corrosion-resistant alloy C-276 and the total volume was 1 L. In order to simulate a fast-heating full-scale system, the reactor was equipped with two heaters, a main reactor heating chamber, and a split ring heater, controlled by Parr 4848 and Parr 4838 reactor controllers, respectively. The head piece of the reactor included a magnetic drive mixing system and a cooling coil controlled by Parr 4848. In each run, the reactor was filled with 500 g of dewatered mixed sludge (20 wt.% TS) and sealed tightly. The headspace of the reactor was purged with nitrogen gas to avoid oxygen interference during the hydrothermal reaction and to simulate a continuous flow HTL reactor. After purging, the reactor was heated to the desired temperature while being mixed at 500 rpm. The mixing allowed equal distribution of heat during the reaction and cooling down. Heating the reactor to the desired temperature took approximately 1 h in each run. Once the target temperature was reached, the reactor was kept at that temperature according to the designed retention time. At the end of the retention time, the cooling system was activated and the reactor was cooled down to room temperature in 1-2 h. The reactor, cooling coil, and mixing system were washed with DCM after each run to avoid interference between runs. A photo of Parr 4570 HT/HP reactor and its internal components are given in Figure 3.2. 39 Figure 3.2. (a) Parr high temperature/high pressure reactor during operation and (b) internal components of the reactor system, which include a magnetic drive mixing system, a cooling coil, and a sampling tube. The HTL operation and the separation of HTL aqueous is summarized in Figure 3.3. After the reactor is cooled down, contents were transferred into a 1 L beaker. HTL aqueous was obtained by filtering the reactor contents with a 20 µm hydrophilic nylon membrane filter. The yield of the aqueous phase was 80-81 wt.% of the reactor effluent for all runs. The separation and characterization of bio-crude, hydrochar, and gaseous phase are outside the scope of this dissertation, which focuses solely on HTL aqueous, but can be found in Liu et al. (2022) [10] and Liu et al. (2023) [13], respectively. The obtained HTL aqueous samples were stored at 4°C and filtered using 0.45 µm membrane filters before characterization analyses of compounds in soluble (dissolved, <0.45 µm) phase. HTL reaction severity (dimensionless) was calculated using the following equation (Eq. 3.1): Severity = log [â« e(ð(ð¡)âðbÏ )ð¡0dt ] (3.1) where t is the time (minutes), T(t) is temperature at given time (t) (°C), Tb is the base temperature (100°C), and Ï is fitted parameter assigned the value of 14.75 [211]. 40 Using double heater, ramping times to 290°C, 300°C, 325°C, 350°C and 360°C were 50, 51, 54, 60 and 64 minutes. Cooling was performed using the cooling coil, resulting in no significant differences in cooling times (<1 min) across various target temperatures. Both heating and cooling effects were considered calculating the severity. Therefore, the discussion on HTL conditions affecting HTL aqueous characteristics includes considerations of both heating and cooling effects. Full temperature profiles of the runs can be found in Liu et al. (2022) [10]. Figure 3.3. Hydrothermal liquefaction of mixed sludge and separation of HTL aqueous 3.2.4 Characterization of hydrothermal liquefaction aqueous stream 3.2.4.1 Chemical oxygen demand (COD) The COD of the HTL aqueous samples was measured according to Standard Methods 5220 [212]. The test was performed in 15 mL borosilicate test tubes. Samples were digested using concentrated (98 wt.%) H2SO4 in the presence of HgSO4 and Ag2SO4 catalysts. The amount of oxygen spent in the test was donated by K2Cr2O7 which changes the absorbance of the liquid sample. After samples, K2Cr2O7, H2SO4, and catalysts were added, then tubes were capped and mixed using a vortex mixer. Digestion was performed at 150°C for 3 hours in a Thermotron 3200 temperature chamber (Thermotron Industries, USA). After the tubes cooled down to room temperature (21-23°C), absorbances at 600 nm were measured in GENESYSâ¢ 10S UV-VIS spectrophotometer (Thermo Fisher Scientific, USA). The calibration of the COD test was determined using potassium hydrogen phthalate (C8H5KO4) as standard. Photos of spectrophotometer used in colorimetric tests and colorimetric test tubes are given in Figure 3.4. Calibration curves of COD along with the rest of the colorimetric analyses can be found in Appendix A (Figure A1). 41 Figure 3.4. (a) GENESYSâ¢ 10S UV-VIS spectrophotometer used in the colorimetric tests and (b) test tubes used for colorimetric characterization analyses. 3.2.4.2 Total ammonia Colorimetric ammonia tests were performed following the Standard Methods 4500-NH3 [212]. This method converts ionized ammonia (NH4+) into unionized ammonia (NH3) by increasing pH over 12, therefore, it measures total ammonia (NH3 + NH4+). The chemistry of the method was based on reacting unionized ammonia with hypochlorite and phenol catalyzed by sodium nitroprusside to form a strong blue compound, indophenol, which can be measured by absorbance. The test was performed in 10 mL test tubes cleaned with 0.01 M HCl solution prior to the test to make them ammonia free. In order, complexing agent (KNaC4H4O6·4H2O and Na3O6H5O7·2H2O), alkaline phenol (C6H6O and NaOH), sodium hypochlorite (NaClO), and sodium nitroferricyanide (C5FeN6Na2O) solutions were added, and tubes were mixed with a vortex mixer after the addition of each solution. The tubes were capped and stored in the dark for 1 hour for the reaction to occur. Finally, the tubes were vortexed again and their absorbances were measured at 660 nm with GENESYSâ¢ 10S UV-VIS spectrophotometer (Thermo Fisher Scientific, USA). Calibration was performed with each batch of ammonia test using different dilutions of a 10 g/L NH4+-N equivalent stock NH4Cl solution. 3.2.4.3 Dissolved protein Concentrations of dissolved (<0.45 µm) proteins were measured using a modified Lowry method [213]. According to the Lowry method, peptides reacted with an alkaline copper solution (CuSO4, Na2CO3, NaOH, and KNaC4H4O6·4H2O), and then copper interacts with phosphomolybdic/phosphotungstic acid, also known as Folin-Ciocalteu reagent. The tests were performed in 10 mL borosilicate test tubes. After the samples and solutions are added to the test 42 tubes, they were mixed and left in the dark for 2 hours. At the end of the reaction time, the tubes were mixed again, and an absorbance measurement was performed at 750 nm with GENESYSâ¢ 10S UV-VIS spectrophotometer (Thermo Fisher Scientific, USA). A bovine serum albumin solution was used for the calibration of the test (Appendix A, Figure A1). 3.2.4.4 pH and total alkalinity The pH of HTL aqueous samples was measured with an electrochemical method using a dual-channel pH meter XL25 (Thermo Fisher Scientific, USA). Total alkalinity was measured by a titrimetric method according to Standard Methods 2320B [212]. Samples were titrated with 0.1 N H2SO4 solution till the pH decreased to 4.5. Alkalinity in CaCO3 equivalent was calculated depending on the amount of H2SO4 spent in the titration. 3.2.4.5 Total carbohydrate Concentrations of total carbohydrates in the HTL aqueous samples were measured with the phenol-H2SO4 (DuBois) method [214]. Following the method, carbohydrates were dehydrated into 2-furaldehyde (furfural, C5H4O2) and 5-hydroxymethyl furfural (HMF, C6H6O3) with H2SO4 in 15 mL borosilicate test tubes. These furanic compounds then reacted with phenol (C6H6O) to produce a gold-orange color. The absorbance change caused by the carbohydrates in the samples was detected by spectrophotometric measurement at 490 nm with GENESYSâ¢ 10S UV-VIS (Thermo Fisher Scientific, USA). A glucose solution prepared with anhydrous dextrose was used to determine the calibration (Appendix A, Figure A1). 3.2.4.6 Orthophosphate-P Phosphorus concentrations of HTL aqueous samples, in form of orthophosphate (PO4-3), were measured according to Standard methods 4500-P E [212]. Colorimetric orthophosphate tests were performed in 10 mL glass test tubes cleaned with 0.01 M HCl. Ammonium molybdate ((NH4)6Mo7O24·4H2O) and antimony potassium tartrate (K2Sb2C8H4O12·3H2O) were added to react with orthophosphate and form a heteropoly acid, phosphomolybdic acid. This acid was then reduced to molybdenum blue by adding ascorbic acid (C6H8O6). The absorbance increases caused by orthophosphate in samples were detected by GENESYSâ¢ 10S UV-VIS spectrophotometer (Thermo Fisher Scientific, USA) at 880 nm wavelength. A potassium dihydrogen phosphate (KH2PO4) solution was used as the standard for the test calibration (Appendix A, Figure A1). 43 3.2.4.7 Total phenolic compounds Total phenolic compoundsâ concentration was measured following the Folin-Ciocalteu procedure [215]. In 3 mL centrifuge tubes, samples were mixed with sodium carbonate (Na2CO3) and Folin-Ciocalteu reagent to reduce phenolic compounds into a blue-colored complex. The mixtures were transferred to a microplate, then kept at 40°C for 5 minutes to complete the reaction. The microplate was measured in Synergyâ¢ HT Microplate Multi-Detection Reader (BioTek Instruments, USA) at 735 nm wavelength. Gallic acid was used as the standard for calibration and results were given in mg gallic acid equivalent (GAE)/L (Appendix A, Figure A1). 3.2.4.8 Volatile fatty acids (VFA) Concentrations of acetic, propionic, butyric, valeric, and iso-valeric acids were measured with a 7890A gas chromatography (GC) system (Agilent Technologies, USA) equipped with a flame ionization detector (FID). A 25 m 19091 column (Agilent Technologies, USA) was used in the measurements utilizing He as the carrier gas at the flow rate of 40 mL/min. The GC-FID column, oven, and detector temperatures were 220°C, 70°C, and 300°C, respectively. 3.2.4.9 Furanic compounds, levulinic acid, ethanol, and glycerol An Agilent Infinity 1220 high-performance liquid chromatography (HPLC) system was used to determine the concentrations of 2-furaldehyde (furfural), HMF, levulinic acid, ethanol, and glycerol (glycerine). The HPLC was equipped with a 1260 Infinity II refractive index detector (RID) (Agilent Technologies, USA), and Aminex® HPX-87H column (Bio-Rad Laboratories, USA). The column and RID temperatures were 60°C and 50°C, respectively. A 0.005 M H2SO4 solution was used as the mobile phase at a 0.6 mL/min flow rate. Photos of the GC-FID system mentioned in Section 3.2.4.8 and HPLC-RID system are given in Figure 3.5. 3.2.4.10 Gas chromatography-mass spectrometry (GC-MS) analysis The composition of the HTL aqueous samples was analyzed using an Agilent 6890N GC system (Agilent Technologies, USA) coupled with a 5975B mass selective detector (MSD). Separation was achieved on an Agilent DB-Petro column (Agilent Technologies, USA). The temperatures of the injector and the column were maintained at 300°C and 325°C, respectively. For each analysis, 1 µL of the sample was injected using a 1:50 split ratio. The resulting chromatographic peaks were identified by comparing them to the standard compounds listed in Agilentâs GC-MSD chemical library. More details regarding the GC-MS analysis can be found elsewhere [216]. 44 (a) (b) Figure 3.5. (a) Gas chromatograph equipped with a flame ionization detector (GC-FID) and (b) high-performance liquid chromatograph with a refractive index detector (HPLC-RID) systems located at the Bioreactor Technology Group laboratory. 3.2.4.11 Ultimate analysis and inorganic elements The ultimate analysis was performed to detect the total C and N contents of HTL aqueous samples by a Costech ECS 4010 automated elemental analyzer (Costech Analytical Technologies, USA). Prior to inductively coupled plasma mass spectrometry (ICP-MS) analysis, samples were digested following the US Environmental Protection Agency (EPA) Method 3051A [217] with ETHOS-EZ microwave digestion system (Milestone Inc., USA). Extracts were diluted to 1% (v/v) nitric acid concentration. Measurement of inorganics elements was performed using an Agilent 8900 triple quadrupole ICP-MS (Agilent Technologies, USA). 3.2.4.12 Data analysis Throughout the chapter, error bars given indicate Â± standard deviation. Analysis of variance test (ANOVA) was conducted to reveal the statistical significance of direct, second degree, and interaction relation of independent variables and dependent variables. Also, the HTL aqueous characterization results were mathematically related to the independent variables of reaction temperature and retention time by performing regression analysis. The simulation revealed alterations in the characterization of HTL aqueous within the experimental design range, focusing on understanding trends rather than optimization. For each characterization analysis, six data points were analyzed, obtained from two HTL aqueous samples from different runs, with three subsamples taken from each run. Prior to ANOVA, normality of data was checked using Kolmogorov-Smirnov test on OriginPro 2023b software (OriginLab Corporation, Massachusetts, USA) (Appendix B, Table B7). ANOVA and the regression analysis were performed using Design 45 Expert 13 software (Stat-Ease Inc., USA) without any data transformation. While selecting the terms in the analysis, the highest coefficient of determination (R2), adjusted-R2, and predicted R2 were considered. The obtained regression equations were in the form of equation (Eq. 3.2) given below. Y = β0 + Î²1ð´ + Î²2ðµ + Î²3ð´ðµ + Î²4ð´2 + Î²5ðµ2 + Î²6ð´2ðµ + Î²7ð´ðµ2 + Î²8ð´2ðµ2 (3.2) where Y is dependent variable (mg/L or unitless), β0 is intercept, β0-8 are regression coefficients, A is reaction temperature (°C) and B is retention time (minutes). 3.3 Results and discussion 3.3.1 Chemical oxygen demand (COD) The COD concentrations of HTL aqueous samples represent the amount of organic molecules remaining in a dissolved state in the process water after the reaction is complete. COD was measured by chemically oxidizing all organics in the HTL aqueous samples and measuring the amount of oxygen spent during oxidation in mg O2/L unit. In HTL, the COD change in different HTL aqueous samples can be seen as a combined effect of hydrolysis, decomposition, recombination, char-forming (condensation, cyclization, and polymerization), and gasification reaction rates, therefore, the relationship between reaction conditions and COD level of HTL aqueous is quite complex that cannot be explained with a linear pattern. The COD values of HTL aqueous samples within the designed HTL conditions (290-360°C and 0-30 min) were found to be in the range of 8,711-102,845 mg O2/L, given in Figure 3.6(a). The high COD levels of HTL aqueous were caused by the hydrolyzed basic organic molecules that did not form bio-crude or hydrochar, and molecules that were not gasified into CO2, CO, or CH4, along with a portion of bio-crude due to the enhanced solvent properties of water under HTL conditions [9]. These dissolved organic molecules were mainly N-heterocyclics, organic acids, amides (including amino acids and peptides), amines, ketones, phenolics, esters, and alcohols (Appendix B, Tables B1-6). There was not a significant variation in COD results of HTL aqueous samples obtained at milder conditions than the center experimental point (325°C, 15 min). However, under more severe conditions than the central point, COD was found to decrease due to the gasification and recombination (bio-crude forming and char-forming) reactions. This pattern in the COD change can also be seen as a loss of carbon to the gas and solid phases [10]. Furthermore, HTL conditions around the central point were identified as temperature and retention time that accumulate the most amount of carbon in the bio-crude phase (optimum point being 46 332°C-16.9 min, severity of 8.2) [10]. This may contribute to the observed 10,000 mg/L drop in COD under conditions of 325°C and 30 minutes, which have a severity level similar to the optimum point. (a) (b) (c) (d) (e) (f) Figure 3.6. Dewatered mixed sludge derived HTL aqueous characterization results of (a) COD (n=6), (b) dissolved protein and total ammonia (n=6), (c) VFAs (n=6), (d) total alkalinity and pH (n=4), (e) phosphorus (PO4-3-P) (n=6), (f) total carbohydrates and total phenolic compounds (n=6) at various reaction temperatures and retention times (GAE: gallic acid equivalent). 47 The COD concentration of HTL aqueous is highly dependent on feedstock characteristics (mainly carbohydrate and protein concentrations), HTL reaction conditions, and initial solid concentration of feedstock. Therefore, various COD concentrations were reported in the literature for municipal sludge based HTL aqueous. Chen et al. reported 80-105 g COD/L concentrations for HTL aqueous obtained from dewatered municipal sludge (15.5 wt.% TS) processed at 200-320°C and 30 min [218]. By using sludge from the same WWTP in this study (Annacis Island, Delta, BC, Canada), Maddi et al. obtained HTL aqueous with 40.8, 73, 48.2 g COD/L from primary sludge (11 wt.% TS), secondary sludge (8.1 wt.% TS) and digested sludge (11.5 wt.% TS), respectively [118]. When the COD results in this study and the literature are evaluated, it can be seen that processing secondary sludge in HTL causes higher COD levels in HTL aqueous. This could be due to the higher protein content of secondary sludge linked to extra- and intra-cellular biopolymers that remain as dissolved amides in HTL aqueous. 3.3.2 Dissolved proteins and total ammonia Under subcritical HTL reaction conditions, proteins hydrolyze but do not totally decompose into smaller compounds. Therefore, it is possible to observe high protein concentrations in HTL aqueous depending on the protein content of the feedstock and reaction conditions. In this study, dissolved protein concentrations of 11,591 ± 270 and 21,493 ± 589 mg/L were measured under the second most and least severe conditions of 350°C-25.6 min (severity of 8.81) and 300°C-4.4 min (severity of 6.79), respectively. After N-heterocyclic compounds, proteins were found to be the most abundant compound group under milder HTL conditions than the middle point of 325°C-15 min (severity < 8.8). The results indicate that more proteins were decomposed as the reaction severity increased as the decomposition of proteins is the main source of ammonia in HTL aqueous. Under hydrothermal conditions up to 300°C, ammonia only comes from the decomposition of proteins, whereas over 300°C, amides and N-heterocyclic compounds, such as pyridine, decompose and contribute to ammonia formation [219]. In the literature, a protein concentration of 14,455 mg/L was reported by Chen et al. for an HTL aqueous obtained from dewatered municipal sludge with 15.5 wt.% TS [220]. Furthermore, the authors reported that under 320°C-30 min condition, asparagine, threonine, serine, cysteine, lysine, histidine, and arginine were completely decomposed, whereas glutamic acid, valine, leucine, γ-aminobutyric acid, and ornithine were the main amino acids remaining in HTL aqueous [220]. The decomposition of cysteine and methionine could be a possible source of sulfur in HTL that causes H2S and CH4S formation in the gas phase. 48 Since the decomposition reaction of protein deamination (Reaction 3.1) causes ammonia accumulation in HTL aqueous, protein and ammonia concentrations were inversely related. Both protein and ammonia concentrations were highly linked to HTL reaction conditions. Measured dissolved protein and total ammonia concentrations in HTL aqueous samples are given in Figure 3.6(b). Total ammonia concentrations of 3,929 ± 47 and 5,799 ± 50 mg/L were obtained under conditions of 300°C-4.4 min (severity of 6.79) and 350°C-25.6 min (severity of 8.81), respectively. As there was still over 10,000 mg/L protein left in HTL aqueous derived at 350°C-25.6 min, longer retention times could result in higher ammonia concentrations. It was found that reaching supercritical conditions significantly increases the degradation of remaining proteins [221]. Under catalytic hydrothermal gasification (CHG) conditions, where 99% of the organics were gasified, ammonia concentrations up to 19,367 mg/L were observed [222]. Higher ammonia concentrations provide an opportunity for N recovery from HTL aqueous by applying processes such as ammonia stripping and struvite precipitation. However, high total ammonia levels can also be a drawback due to its inhibitory effect on downstream biological treatment units. R â CH(NH2) â COOH deamination â R â C(O) â COOH + NH3 Reaction (3.1) 3.3.3 Volatile fatty acids (VFAs) Volatile fatty acids (carboxylic acids) are one of the major compound groups in HTL aqueous due to the number of decomposition reactions that form them. VFAs are formed in the decomposition of carbohydrates, proteins, and lipids [9]. Hydrothermal carbohydrate decomposition pathway mostly ends up as VFAs if there are no amino acids for reducing sugars to go under Maillard reaction [56]. Also, some of the amino acids form VFAs after they lose the amine group in the deamination reaction. Lipids are the least likely group to form VFAs in HTL reactions, however, severe hydrothermal conditions and catalytic effects are known to crack long chains into short chain fatty acids [223]. In HTL aqueous from municipal sludge, the shortest fatty acid detected was acetic acid due to formic acid being unstable under hydrothermal conditions and gasified into CO, CO2, and H2 [224]. In this study, VFAs were the second most abundant group at the central point and more severe conditions (severity â¥ 7.9). Fatty acid compounds of acetic, propionic, butyric, valeric, iso-valeric, hexanoic, 4-methyl valeric, heptanoic, and decanoic acids were detected by GC-FID in all HTL aqueous samples, but only acetic, propionic, butyric, valeric, and iso-valeric acids were quantifiable (Figure 3.6(c)). As a result of hydrothermal decomposition, concentrations of VFAs 49 increased as the molecular chain became shorter except for the 325°C-30 min condition where valeric acid concentration was higher than butyric acid. This is possibly caused by the decomposition of a longer chain fatty acid at this point. As decomposition products, all quantifiable VFAs yielded the lowest concentrations under the least severe HTL condition of 300°C-4.4 min. The highest acetic, propionic, and butyric acid concentrations were observed in HTL aqueous obtained under 325°C-0 min condition as 8,653 ± 11 mg/L, 3,229 ± 108 mg/L, and 2,385 ± 31 mg/L, respectively. Valeric and isovaleric acids, on the other hand, were the highest under 325°C-30 min and 360°C-15 min conditions. Five-carbon fatty acids having higher concentrations under more severe HTL conditions indicate that they were decomposed from longer chain fatty acids, possibly from lipid decomposition intermediates. The acetic acid concentration was found to decrease under more severe HTL conditions than 325°C-0 min. The first reason for this decrement could be the partial gasification of acetic acid that converted it into CO2 and CH4 (Reaction 3.2). The second reason could be acetic acid reacting with ammonia and forming acetamide (Reactions 3.3 and 3.4) [225]. GC-MS results, given in Section 3.8, revealed significant increases in acetamide concentrations under 325°C-15 min and 360°C-15 min conditions, supporting the occurrence of a reaction between acetic acid and ammonia. CH3COOH gasification â CH4 + CO2 Reaction (3.2) CH3COOH + NH3 â CH3COO(NH4) Reaction (3.3) CH3COO(NH4) dehydration â CH3CO(NH2) + H2O Reaction (3.4) Ammonia is also known to react with longer-chain VFAs than acetic acid and form ammonium salts, such as ammonium propionate and ammonium butyrate. Similar to ammonium acetate, these compounds then dehydrate under HTL conditions and form fatty acid amides of propanamide, butyramide [226], and their methylated versions of N-methyl, N-ethyl, N,N-diethyl. Also, long-chain fatty acid amides were reported to contribute to bio-crude formation [18]. The reason for decreasing propionic and butyric acid concentrations could be their reactions with ammonia. Owing to these reactions, fatty acid amides become one of the major organic compound groups in the HTL aqueous. In a WWTP scenario, all VFAs and fatty acid amides would be important substrates for downstream anaerobic digestion of HTL aqueous for further energy recovery. If the treatment plant does not have anaerobic digesters, these compounds can also be removed during aerobic treatment due to their high biodegradability but would cause much higher biomass yields and cause an increase in secondary sludge generation. 50 VFA concentrations of HTL aqueous reported in the literature vary due to the differences in the feedstock characteristics and HTL reaction conditions [49]. Maddi et al. reported 3,243 mg/L, 3,303 mg/L, 3,078 mg/L acetic acid and 856 mg/L, 965 mg/L, 427 mg/L propionic acid concentrations in aqueous samples obtained from the HTL of primary sludge (11 wt.% TS), secondary sludge (8.1 wt.% TS) and digested sludge (11.5 wt.% TS), respectively [118]. The difference in the VFA concentrations reported in [118] compared to this study was caused by the solid content and the type/characteristics of feedstock. Chen et al. measured 8,569 mg/L acetic acid, 3,936 mg/L propionic acid, 5,631 mg/L butyric acid, 2,700 mg/L valeric acid, and 1,853 mg/L valeric acid in municipal sludge-derived HTL aqueous obtained under 320°C-30 min condition, which was similar to the VFA results of this study [107]. Since VFAs are mostly formed as a result of carbohydrate decomposition reactions, lignocellulosic feedstocks tend to have the highest VFAs in HTL aqueous phase. Panisko et al. [227] reported acetic acid concentrations as high as 10,800 mg/L for corn stover derived HTL aqueous phase. Municipal sludge, on the other hand, is a more balanced substrate in terms of carbohydrate to protein ratio, which makes it a suitable substrate for HTL [9]. However, results of this study revealed that if feedstock mixed sludge is dewatered to 20 wt.% total solids, total VFA concentration in the HTL aqueous phase can be as high as 17,014 mg/L. This VFA amount represents 9.7 wt.% of the organic fraction of dewatered sludge. Recovering carbon from this portion by utilizing the VFAs in the aqueous phase in a downstream process would increase the overall energy recovery of the HTL process significantly. 3.3.4 Total alkalinity and pH In HTL, water is used as the solvent and medium where all the reactions take place. Water, which is already a good solvent for polar compounds due to being polar, becomes a strong solvent for non-polar molecules too owing to its relative permittivity becoming lower than 20 in the HTL conditions. All the dissolved polar and non-polar compounds cause HTL aqueous to have a strong pH buffering capacity. The total alkalinity range of municipal sludge-derived HTL aqueous in this study was 6,253-14,756 mg CaCO3/L. The lowest and the highest alkalinity levels were measured for 300°C-4.4 min (severity of 6.79) and 350°C-25.6 min (severity of 8.81) conditions, respectively. Alkalinity was found to be directly proportional to the reaction severity. These levels indicate that changing the pH of HTL aqueous for a downstream process would not be economical. However, high alkalinity can be useful in biological treatment processes, especially under high loading rates, where stable pH levels are desired. Total alkalinity and pH results are given in Figure 3.6(d). Although adjusting pH would be difficult due to the high alkalinity levels creating strong buffering (resistance) capacity, the pH of HTL aqueous was found suitable for most of the 51 downstream processes for being close to neutral. The pH was ranging between 6.4-7.9, which is acceptable for both aerobic and anaerobic treatment in a WWTP. The pH levels were found to change depending on VFA and ammonia concentrations. The lowest pH of 6.4 was measured for 325°C-0 min condition which also had the highest VFA concentration of 17 g/L. Since the ammonia concentrations were low under mild HTL conditions of 290°C-15 min, 300°C-4.4 min, and 325°C-0 min, slightly acidic pH levels of 6.4-6.7 were observed. Under more severe HTL conditions, pH started to increase due to the formation of ammonia and the decrease in VFAs described in Sections 3.2 and 3.3. The highest pH of 7.9 was measured for the second most severe HTL condition of 350°C-25.6 min which also had the highest ammonia concentration. Because municipal sludge is a balanced substrate in terms of carbohydrate to protein ratio (1:2), pH levels of municipal sludge derived HTL aqueous stay close to the neutral level. HTL aqueous pH levels can be as low as 3.68 for carbohydrate-rich substrates, such as lignocellulosic biomass [228], and as high as 8.64 for protein-rich complex substrates, such as cyanobacteria [229]. Furthermore, it was reported that when only glucose was used as feedstock in HTL, it caused a pH of as low as 1.5 in the aqueous phase, whereas when only protein was used, the aqueous phase had a pH of 9.5 [230]. Acidic conditions in HTL are known to promote recombination reactions and the formation of hydrochar, which causes less energy recovery in the form of bio-crude, therefore, decreasing the overall energy recovery in the process. Mixed municipal sludge, however, is a balanced feedstock in terms of carbohydrates and protein. Owing to this balance, carbohydrates can be recovered as bio-crude in the Maillard reaction of reducing sugars and amino acids taking place in HTL aqueous [10]. As pH results revealed, the amount of VFA and ammonia forming in the HTL of municipal sludge allow pH to stay close to the neutral level, which prevents the reaction to shift to hydrochar formation. 3.3.5 Phosphorus Along with nitrogen, phosphorus is another important nutrient that needs to be recovered in order to make HTL a process that supports the circular economy concept. Under HTL conditions, all organic P is converted into inorganic forms of apatite phosphate (AP) and non-apatite inorganic phosphate (NAIP) that accumulate in hydrochar, while a portion of P remains in the aqueous phase as orthophosphate ion (PO43-) [78,231]. Within the experimental conditions of this study, 85-92% of P from municipal sludge remained in the hydrochar phase after HTL [13]. In the aqueous phase, on the other hand, 332-587mg P/L equivalent orthophosphate concentrations were measured (Figure 3.6(e)). The highest P in the aqueous phase was detected under the 52 second mildest HTL condition of 300°C-4.4 min. As the severity increased, P concentration in the aqueous phase decreased due to hydrochar formation. The lowest P concentration was measured under the most severe HTL condition of 360°C-15 min (severity of 8.98), indicating that reaction temperature was the primary effective parameter on hydrochar formation, as the models also revealed by Liu et al. (2022) [10]. These P concentrations would be beneficial as a nutrient for the downstream biological treatment processes in a WWTP. However, total ammonia concentrations in the range of 3,231-5,799 mg/L show that PO43- concentration in the HTL aqueous was not high enough to use struvite precipitation to recover both N and P nutrients efficiently. The amount of P remaining in the HTL aqueous phase may vary depending on the WWTP process configuration. If the WWTP is utilizing a biological nutrient removal (BNR) process, municipal sludge would have much higher P concentrations which would lead to higher P in hydrochar and HTL aqueous phases. Therefore, P concentrations may vary highly depending on the feedstock. For the HTL aqueous of primary, secondary, and digested sludge obtained from the same WWTP (Annacis Island, Delta, BC, Canada) as in this study, 24 mg/L, 765 mg/L, and 62 mg/L phosphate anion concentrations (PO43-) were reported, respectively [118]. Phospholipids are an important source of phosphorus in the HTL. Studies in the literature showed that orthophosphate concentrations in the HTL aqueous phase can exceed 3,000 mg/L if a phospholipid-rich substrate is hydrothermally liquefied, such as microalgae [102,113]. 3.3.6 Total carbohydrates and total phenolic compounds In HTL of complex organic substrates, such as municipal sludge, carbohydrates are the first fraction that degrades under hydrothermal conditions. Over 200°C, carbohydrates dehydrate into furanic compounds and after various reactions, they form VFAs [232]. Also, carbohydrates have the least contribution to bio-crude formation [233]. Within the experimental range of this study, total carbohydrate concentration in HTL aqueous ranged from 229 to 961 mg/L (Figure 3.6(f)). The concentrations showed a clear pattern as they were inversely proportional to the severity. HPLC analysis was used to quantify concentrations of carbohydrate degradation products of furanic compounds and levulinic acid. In addition to furans, carbohydrate degradation also forms aldehydes, ketones, and phenolic compounds [234]. In another study, for municipal sludge processed at 290°C-30 min and 320°C-30 min, HTL aqueous carbohydrate concentrations of 162 and 158 mg/L were reported, respectively [218]. Similar carbohydrate levels in the HTL aqueous were also found for algal feedstock [235,236]. 53 Also, it was revealed that using alkaline catalysts significantly increased the amount of carbohydrates in HTL aqueous [235], indicating the decomposition of carbohydrates could be catalyzed by acidic conditions. For this reason, the use of alkaline catalysts has been widely studied for the HTL of carbohydrate-rich substrates, such as lignocellulosic biomass [9]. Phenolic compounds can contribute to the bio-crude formation or stay in the aqueous phase in such forms as phenol, cresol, and 4-ethylphenol. Although in the HTL field phenolic compounds are mostly associated with lignin, they also form as a degradation product of carbohydrates, proteins, and lipids [237]. However, the degradation of lipids and proteins is limited under HTL conditions. Therefore, it is possible to say that when municipal sludge is used as the feedstock in HTL, phenolic compounds in the aqueous phase mainly come from carbohydrate degradation and slightly from the degradation of lipids and proteins. Glucose alone was reported to degrade into significant amounts of phenolic compounds under HTL conditions [230]. In this study, after HTL of mixed municipal sludge, the highest total phenolics concentration was found as 2,299 mg GAE/L for the least severe HTL condition of 300°C-4.4 min. As the reaction severity increased, total phenolic levels gradually decreased, and the minimum concentration was measured as 1,450 mg GAE/L for the 350°C-25.6 min condition. The reason for this decrease could be the recombination reactions of phenolic compounds into bio-crude and hydrochar as the reaction severity increased. Phenols are known to form light polycyclic aromatic hydrocarbons that stay in the bio-crude, and heavy polycyclic aromatic hydrocarbons that form the hydrochar [10,238]. 3.3.7 Furanic compounds, levulinic acid, ethanol, and glycerol By using HPLC, ethanol, carbohydrate degradation products of furfural, HMF, levulinic acid, and lipid degradation product of glycerol concentrations were measured in all HTL aqueous samples. Ethanol, furfural, HMF, and levulinic acid were detected, but it was not possible to quantify them with HPLC due to their low concentrations (<50 mg/L). Because of the low concentration of carbohydrate degradation products, it is possible to say that the reaction of carbohydrates to VFA occurs at a high rate under hydrothermal conditions. Glycerol concentration, on the other hand, was found to be in the range of 421-688 mg/L, as shown in Figure 3.7. The highest glycerol concentration was found for the least severe HTL condition (severity of 6.79) of 300°C-4.4 min. Levels of glycerol decreased as the HTL reaction conditions became more severe. The lowest concentration was found for the most severe condition of 360°C-15 min (severity of 8.98). Although increasing severity causes more lipids to degrade and form glycerol, the aqueous concentrations became lower with the increasing severity due to 54 glycerol not being stable under hydrothermal conditions, and its degradation rate being higher than the rate of lipid degradation. Under subcritical conditions, glycerol degrades into acrolein and acetaldehyde [239]. Glycerol was found to be beneficial as a cosolvent and reactant in the HTL that increases the process efficiency [240,241]. Glycerol was also reported to stay in the bio-crude phase in some studies [14,242]. Figure 3.7. Glycerol concentrations of HTL aqueous samples (n=6) derived from dewatered municipal mixed sludge at various reaction temperatures and retention times. 3.3.8 Gas chromatography-mass spectrometry results A qualitative GC-MS analysis was performed to determine the relative abundancies of the compound groups in HTL aqueous samples obtained from 290°C-15 min, 325°C-15 min, and 360°C-15 min, corresponding to HTL severities of 6.8, 7.8, 8.8, respectively. The relative abundances and the major compounds are given in Figure 3.8. According to the findings, the most abundant compound group in HTL aqueous for all three conditions was N-heterocyclic compounds. Under hydrothermal conditions, N-heterocyclics mostly form during the Maillard reaction which involve reducing sugars and amino acids [9]. The N-heterocyclic compounds are known for their high cytotoxicity [243] and low biodegradability [244] which means their high levels in HTL aqueous can cause a major issue for its downstream biological treatment. The most abundant N-heterocyclic compounds in HTL aqueous were 2-methylpyrazine, 2-piperidonone, 2-pyrrolidinone, 1,3-diazine, 2-ethylpyrazine, 2,5-dimethylpyrazine, 3-pyridinol, and 3-pyridinol. The abundance was found to decrease by the increasing reaction temperature from 325°C to 360°C, supporting the studies in the literature stating that N-heterocyclic compounds degrade under severe HTL conditions [219,245]. The same N-heterocyclic molecules were detected in the bio-crude phase and a decrease in their abundance was observed by the increasing reaction severity [10]. 55 56 Figure 3.8 Major compound groups detected by GC-MS in HTL aqueous samples derived from dewatered municipal mixed sludge under conditions of (a) 290°C-15 min, (b) 325°C-15 min, and (c) 360°C-15 min. Organic acids were identified as the second most abundant compound group. The findings of GC-MS confirm the organic acid concentrations quantified by GC-FID. As in GC-FID findings, the highest abundance of organic acids in the aqueous phase was at the HTL temperature of 325°C. Also, the concentration of valeric acid increased at 360°C. In addition to GC-FID, GC-MS identified 2-methylbutyric acid and 3-phenylpropionic acid as major organic acid compounds in HTL aqueous. At 290°C, large ketone and amino acid-based esters were detected, indicating that the decomposition of the complex organic structure was not complete. At HTL of 325 and 360°C, these compounds were decomposed into smaller molecules. At these temperatures, 2-aminophenol was identified as one of the major amines which have inhibition potential in biological treatment systems [243]. As described in Section 3.3, HTL aqueous contains amides as a result of VFA and ammonia reactions during the process. These amides were mainly acetamide, propionamide, N-methyl acetamide, N-ethyl acetamide, N-methyl propionamide, N,N-dimethyl propionamide, and butanamide. Longer chain amides, such as octanamide and decanamide, forming as a reaction of long chain fatty acids and ammonia, were found to stay in the bio-crude phase [10]. The hydrothermal pathway from municipal sludge to compounds in HTL aqueous detected by GC-MS is presented in Figure 3.9. The full lists of GC-MS results for all three HTL conditions are given in Appendix B (Tables B1-6).57 Figure 3.9. Hydrothermal formation pathways of compounds and groups detected in GC-MS analysis. 58 3.3.9 Ultimate analysis and inorganic elements The ultimate analysis was performed to determine the amount of C, N, and S remaining in the HTL aqueous phases and the results are presented in Figure 3.10 (a). The S content was found under the detection limit while the C content was found in the range of 2.55-3.00%, showing a similar pattern to COD levels as C and H are the most effective two elements that determine the COD level. As the HTL reaction severity increased, C amount in the aqueous phase gradually decreased because of gasification and char-forming reactions. If reaction severity was further increased to supercritical conditions, it would be possible to gasify all organic carbon [222]. The N content of HTL aqueous samples was in the 0.71-0.86% range, and there was no significant pattern with HTL severity. The N content measured in HTL aqueous comes from the N in ammonia, amines, amides, and N-heterocyclic compounds. Although N-heterocyclic compounds and proteins decompose more by the increasing reaction severity, the ultimate analysis revealed that these decomposition reactions do not cause N to shift to gas, hydrochar, or bio-crude phase. Along with the ultimate analysis, to detect the concentrations of trace elements, an ICP-MS analysis was performed, and the results are presented in Figure 3.10 (b) and (c). According to the results, P concentration was consistent with the colorimetric method, indicating that approximately all P in HTL aqueous was in orthophosphate (POâ³â») form. S levels of 169-331 ppm were detected in HTL aqueous samples that decreased by increased reaction severity, possibly due to shifting to hydrochar and gaseous phases. A major source of S in the HTL aqueous phase was the decomposition of S containing amino acids of methionine, cysteine, and taurine. The S in HTL aqueous was reported to be in cyclic compounds like thiazoles [9] or bonded to other complex molecules and formed thione as revealed by the GC-MS analysis given in Section 3.8. As most of the inorganic elements stay in the hydrochar phase in HTL [13], there was no unusually high concentration detected in the ICP-MS analysis. The levels of Al and Fe were caused by the chemicals used in wastewater treatment and sewage collection systems, such as Al2(SO4)3 and FeCl3. Except for Fe and S, there was no clear pattern of concentration change at different reaction severity. There was a significant decrease in Fe levels at higher reaction severities, indicating a shift to hydrochar. The ICP-MS results revealed that regardless of reaction condition, HTL aqueous has sufficient macro- and trace elements required for microbial growth for a potential downstream biological treatment of aqueous. 59 (a) (b) (c) Figure 3.10. (a) Ultimate analysis (n=2), and (b), (c) trace element analysis (n=4) results of HTL aqueous samples derived from dewatered municipal mixed sludge (MS) at various reaction temperature and retention times. 3.3.10 ANOVA test and regression analysis The statistical relation of independent variables, HTL reaction temperature and retention time, and dependent variables of aqueous characterization parameters (pollutants/compounds quantified) were investigated using the ANOVA test. The analysis result for each characterization parameter is given in Table 3.1. The results were evaluated according to a 95% confidence interval. The ANOVA tests of all parameters were significant due to p values being <0.05. According to the ANOVA test results, except for acetic acid and propionic acid, the effect of HTL reaction temperature on all parameters was found statistically significant. For these parameters, 60 retention time was found to be more effective within the experimental design conditions of this study (290-360°C, 0-30 min). For retention time, on the other hand, only the effect on total phenolics concentration was found insignificant (p> 0.05). For all measured parameters, at least one of the independent factors was statistically significant. Although retention time is not as effective as reaction temperature in the severity formulation, the ANOVA test results indicate that it is also equally important for changing HTL aqueous characteristics. This effect was more obvious on individual VFA concentrations. The most significant term on all aqueous characterization parameters was retention time (B), whereas the least significant was AB2 (reaction temperature Ã retention time2). The least significant term, AB2, was kept in the test to be able to use the A2B2 term without disturbing the model hierarchy. Table 3.1. ANOVA test results of HTL aqueous characterization data. Parameter Terms A B AB A2 B2 A2B AB2 A2B2 ANOVA test p-value COD <0.0001 0.0003 0.2755* 0.0003 <0.0001 0.0002 0.5449* 0.0023 <0.0001 Dissolved proteins <0.0001 <0.0001 <0.0001 <0.0001 <0.0001 - 0.1844* - <0.0001 Total ammonia <0.0001 <0.0001 0.0792* <0.0001 <0.0001 - <0.0001 - <0.0001 Total alkalinity <0.0001 <0.0001 0.2229* 0.0166 0.0001 0.2529* 0.0089 <0.0001 <0.0001 pH <0.0001 <0.0001 0.0333 0.0213 0.0003 0.0123 0.4112* <0.0001 <0.0001 Phosphorus (POâ³â») <0.0001 0.0329 0.0015 0.4265* 0.0164 0.1767* 0.5878* 0.0008 <0.0001 Total carbohydrates <0.0001 <0.0001 <0.0001 <0.0001 <0.0001 0.0378 <0.0001 0.0002 <0.0001 Total phenolics <0.0001 0.0648* 0.0008 0.0005 - <0.0001 - - <0.0001 Glycerol <0.0001 0.0049 0.0884* 0.2114* 0.4863* 0.3680* 0.7378* <0.0001 <0.0001 Total VFAs 0.0005 0.0068 <0.0001 0.1548* 0.0018 <0.0001 0.0563* <0.0001 <0.0001 Acetic acid 0.1182* 0.0009 0.0001 0.0867* 0.0015 <0.0001 0.8657* <0.0001 <0.0001 Propionic acid 0.2923* 0.0002 0.0007 <0.0001 0.5316* <0.0001 0.7425* <0.0001 <0.0001 Butyric acid 0.0131 <0.0001 0.0743* <0.0001 0.0006 <0.0001 <0.0001 <0.0001 <0.0001 Valeric acid <0.0001 <0.0001 0.0028 0.0028 0.0368 0.0005 0.8789* <0.0001 <0.0001 Isovaleric acid <0.0001 0.0001 0.8073* 0.0634* 0.1418* 0.0147 0.0010 <0.0001 <0.0001 A: Temperature (°C), B: Retention time (min), *: Statistically insignificant terms (for 95% confidence interval) The characterization data of HTL aqueous was used in the regression analysis to mathematically relate the independent variables of reaction temperature and reaction time to dependent variables. The regression equations for all parameters are given in Table 3.2. It was possible to model the pattern of dependent parameters with high coefficient of determination (R2) values of 0.9260-0.9975. Adjusted-R2 values, which are a better indicator due to considering the number of independent variables used for prediction, were within the range of 0.8722-0.9956. 61 Predicted-R2 values, which show the fit of regression equations on actual experimental data, were in the 0.8400-0.9910 range. The highest prediction power was found for the regression equation of carbohydrates as the pattern given in Section 3.3.6 was very clear. The lowest R2, adjusted-R2, and predicted-R2 values were for the phosphorus equation. Using the regression equations, contour plots of predicted HTL aqueous characterization parameters and VFAs were presented in Figure 3.11 and Figure 3.12, respectively. These regression equations could be utilized by wastewater utilities to predict the HTL aqueous characterization depending on the HTL process temperature and retention time fluctuations and make necessary decisions for downstream treatment processes. Table 3.2 Regression equations of HTL aqueous characterization parameters. Parameter Regression coefficients Fit statistics Intercept A B AB A2 B2 A2B AB2 A2B2 R2 Adjusted R2 Predicted R2 COD -240690 2166 -48452 307 -3.5 3346 -0.48 -21 0.03231 0.9551 0.9224 0.8592 Dissolved proteins 246531 -1220 -2688 6.9 1.6 44 - -0.11 - 0.9881 0.9826 0.9729 Total ammonia -63121 356 1783 -5.1 -0.5 -62 - 0.18 - 0.9840 0.9767 0.9660 Total alkalinity 252596 -1653 -48126 305 2.8 1730 -0.48 -11 0.01707 0.9898 0.9823 0.9775 pH 163.4 -0.99 -22 0.13 0 0.63 0 0 0.00001 0.9717 0.9511 0.9074 Phosphorus (POâ³â») 25218 -149 -3029 18.6 0.22 89 -0.03 -0.55 0.00086 0.9260 0.8722 0.8400 Total carbohydrates -5619 55.67 1988 -13.4 -0.11 -60 0.02 0.39 -0.00063 0.9975 0.9956 0.9910 Total phenolics 86818 -510 -4375 26.5 0.76 - -0.04 - - 0.9459 0.9266 0.9203 Glycerol 25466 -151 -3283 20.3 0.23 105 -0.03 -0.65 0.001 0.9588 0.9288 0.8810 Total VFAs -1087780 6721 124560 -762 -10 -3543 1.16 22 -0.03325 0.9801 0.9657 0.9470 Acetic acid -455608 2832 48359 -297 -4.3 -1255 0.45 7.7 -0.01189 0.9580 0.9275 0.8891 Propionic acid -232449 1441 22725 -139 -2.2 -603 0.21 3.7 -0.00569 0.9643 0.9384 0.9016 Butyric acid -158855 979 19259 -117 -1.5 -418 0.18 2.5 -0.00378 0.9774 0.9610 0.9242 Valeric acid -156108 955 21047 -129 -1.5 -791 0.20 4.9 -0.0075 0.9809 0.9669 0.9314 Isovaleric acid -84757 515 13170 -80 -0.77 -477 0.12 2.9 -0.0044 0.9801 0.9657 0.9327 A: Temperature (°C), B: Retention time (min) Contour plots given in Figure 3.11, drawn using the regression equations show the changes in HTL aqueous characterization under various HTL conditions. The patterns related to hydrothermal conditions for all parameters were clear except for COD. The pattern for aqueous COD level was caused by the number of complex reactions that affect it, such as hydrolysis, decomposition, recombination, char-forming, and gasification. Other parameters were not affected by so many reactions; therefore, it was easy to relate them to HTL reaction conditions. According to the contour plots, total ammonia, total alkalinity, and pH were directly related to 62 reaction severity, whereas dissolved proteins, phosphorus, total carbohydrates, total phenolics, and glycerol were inversely proportional to reaction severity. In Figure 3.12, acetic, propionic, and butyric acids were favoring 325-330°C and 0 min conditions. Since they have the highest concentration among VFAs, they also caused total VFAs to be the highest under these conditions. Valeric and isovaleric acids were the highest under 325°C-30 min and 360°C-15 min conditions. However, their concentrations decrease at 350-360°C reaction temperatures and retention times beyond 20 min, possibly due to decomposition into smaller VFAs. (a) (b) (c) (d) (e) (f) (g) (h) (i) Figure 3.11 Predicted values of (a) COD, (b) dissolved protein, (c) total ammonia, (d) total alkalinity, (e) pH, (f) phosphorus (PO4-3-P), (g) total carbohydrates, (h) total phenolics, and (i) glycerol by the regression analysis for HTL aqueous derived from dewatered municipal mixed sludge at various reaction temperatures and retention times (GAE: gallic acid equivalent). 63 The regression models in this study were built using the characterization of 20 HTL aqueous samples obtained under 9 different HTL conditions. The obtained regression equations only represent the circular area staying within the experimental conditions as can be seen in Figure 3.11. Graphs of one-factor models illustrating variations in HTL aqueous characteristics, based on individual independent parameters, are provided in Appendix A (Figures A1-3). For a wider application, models can be enhanced by using larger data sets with more experimental points and wider ranges. The use of an increased number of data points would result in a model with higher resolution and more precise predictions. (a) (b) (c) (d) (e) (f) Figure 3.12. Predicted values of (a) total VFAs, (b) acetic acid, (c) propionic acid, (d) butyric acid, (e) valeric acid, and (f) isovaleric acid obtained by the regression analysis for HTL aqueous derived from dewatered municipal mixed sludge at various reaction temperatures and retention times. 3.3.11 Evaluation of on-site biological treatment options for HTL aqueous in a WWTP On-site treatment of HTL aqueous would be the most economical option due to its high volume and resource (biogas, nutrients) recovery potential from elevated dissolved organics, such as VFA, and ammonia levels. Three HTL aqueous biological treatment options in a WWTP would be aerobic treatment (returning HTL aqueous to secondary treatment), mesophilic anaerobic digestion, or thermophilic anaerobic digestion. In this section, potential positive and negative 64 effects of HTL aqueous characterization parameters on these three treatment options are evaluated based on the various compound ranges obtained. A summary of the evaluation is given in Table 3.3. Table 3.3. Potential effect of HTL aqueous characteristics on on-site biological treatment options. Parameter Aerobic treatment Mesophilic anaerobic treatment Thermophilic anaerobic treatment COD ο ο ο Dissolved proteins ο ο ο Total ammonia - -- -- Total alkalinity + + + pH + + + Phosphorus (PO4-3-P) + + + Total carbohydrates ο ο ο Total phenolics -- -- -- Glycerol - ο ο VFAs ο + + Trace elements + + + N-heterocyclic compounds - -- -- ο: No negative effect on treatment at the measured range +: Beneficial for biological treatment at the measured range -: Can have negative effects at the measured range --: Can have significant negative effects at the measured range Compared to wastewater and even primary and secondary sludge, HTL aqueous has higher COD levels of 87-103 g/L. Preferably, high COD wastes should be treated in anaerobic digestion systems, however, if the WWTP does not have digesters, it could be returned to aerobic treatment units and treated along with wastewater. As long as the organic loading rate ranges of 0.1-0.3 g COD/LÂ·d for aerobic treatment, 1-2 g COD/LÂ·d for mesophilic anaerobic treatment [246], and 2-3 COD/LÂ·d for thermophilic anaerobic treatment [247] are not exceeded, COD level would not be a source of inhibition in biological systems. Dissolved proteins, carbohydrates, VFAs, and glycerol can be efficiently removed in biological treatment systems. However, if HTL aqueous is returned to aerobic treatment, glycerol could cause a bubbling issue in the aeration tanks by lowering the surface tension of bubbles. Therefore, glycerol concentration should be closely monitored before introducing HTL aqueous to aerobic treatment. Although ammonia can be an important N source in biological treatment, its concentrations in HTL aqueous samples are sufficient to inhibit microorganisms. In aerobic treatment, HTL 65 aqueous would be blended into wastewater which would bring ammonia concentration to insignificant levels. However, mesophilic and thermophilic anaerobic digestion systems cannot handle ammonia concentrations higher than 4,000 mg/L and 5,500 mg/L, respectively [248]. Hence, ammonia could cause total inhibition of anaerobic digestion if HTL aqueous is fed into digesters without dilution. To avoid ammonia inhibition, co-digestion of HTL aqueous with another substrate can be a good strategy. Like N, P is another essential nutrient for biological treatment. The PO4-3-P levels measured in HTL aqueous, 332-587 mg/L, would be beneficial for biological treatment as a nutrient and these P levels do not have inhibition potential. Along with these essential nutrients, it was found that HTL aqueous also contains sufficient levels of trace elements that are required by microorganisms in biological treatment. HTL aqueous pH levels were in the range of 6.4-7.9. The operation pH range for aerobic treatment, mesophilic and thermophilic digestion are 6-8, 6.5-8 [246], and 6-8.5 [249]. Even though the pH of HTL aqueous fits the operation ranges of all three treatment options, HTL aqueous has the potential of causing pH fluctuations in biological treatment processes if it is used as the sole substrate due to the different consumption rates of VFAs and ammonia. Under a situation like this, high alkalinity levels in HTL aqueous would provide buffering capacity for a stable pH to a certain extent which is desirable for all three biological treatment processes. Parameters that cause the biggest challenge for the biological treatment of HTL aqueous are phenols and N-heterocyclic compounds. At low concentrations, some of the members of these organic molecule groups are biodegradable [250,251]. However, phenol levels of 1,000 mg/L [252] and 1,250-1500 mg/L [253] were reported to cause total inhibition of aerobic and anaerobic systems, respectively, whereas N-heterocyclicsâ inhibitory effects are well-known even at concentrations as low as 79 mg/L (pyridine) [243]. The co-existence of both groups of inhibitors in HTL aqueous makes it more challenging to treat it biologically. It was reported that phenol concentrations over 400 mg/L inhibit the degradation of pyridines causing a combined inhibition effect [254]. A pretreatment, such as advanced oxidation, could be applied to decrease the phenolic and N-heterocyclic concentrations prior to biological treatment. Removal of N-heterocyclic compounds would also decrease the recalcitrant COD significantly, which would allow lower COD concentrations in WWTPâs final effluent. According to the characterization results, HTL aqueous obtained under 325°C-0 min condition was the most suitable for biological treatment as most of its COD was caused by VFAs, proteins, carbohydrates, and glycerol. Also in this condition, the concentration of possible inhibitory compounds, such as ammonia and phenol, was lower compared to other HTL aqueous 66 samples. Although it is possible to determine the HTL operation condition that produces HTL aqueous with the highest biodegradability, the HTL process should not be designed for this condition. As HTL is a biomass to biofuel conversion method that consumes significant amount of energy, it should always be designed for the highest bio-crude production [10]. Under the optimum bio-crude production condition (332°C-16.9 min) [10], most carbon and larger molecules shift to the bio-crude phase, leaving less COD and possible inhibitors in the HTL aqueous. Hence, it is possible to produce the maximum amount of bio-crude without compromising much from HTL aqueous biodegradability. 3.4 Summary An extensive characterization of dewatered (20% TS) municipal mixed sludge derived HTL aqueous phases, obtained at 290-360°C and 0-30 min conditions, was performed. Regression analyses were performed to create mathematical relations between HTL temperature/retention time conditions to its aqueous characteristics. Furthermore, HTL characterization data and the trends in pollutant/compound levels were used to evaluate potential downstream biological treatment routes for HTL aqueous as on-site treatment. The key findings were as follows: â¢ HTL aqueous contains dissolved protein concentrations up to 22 g/L that decompose into ammonia and fatty acids as the reaction severity increases. The degradation of proteins causes total ammonia concentrations as high as 7.1 g/L at the HTL condition of 350°C- 25.6 min. â¢ The N-heterocyclic compounds were identified as the most abundant organic molecule group in HTL aqueous. Due to their inhibitory nature, N-heterocyclics would be the biggest obstacle to biologically treating HTL aqueous in a WWTP adopting HTL as sludge treatment. â¢ HTL aqueous was found to have a suitable pH of 6.4-7.9 and elevated total alkalinity levels of 6.3-14.8 g CaCO3/L which would help maintain stable pH during downstream biological treatment. â¢ HTL aqueous was found to contain sufficient macro- (N, P, K, S) and micro-trace elements for biological treatment. â¢ Volatile fatty acids were identified as the second most abundant group in HTL aqueous. The total VFA levels were measured as high as 17 g/L for 325°C-0 min. In this condition, acetic, propionic, and butyric acids were at the highest concentrations, whereas longer chain VFAs favored more severe HTL conditions. Solely based on the characterization results, the 67 aqueous derived at 325°C-0 min is found to be the most suitable stream for anaerobic digestion owing to its elevated VFA levels. The next chapter (Chapter 4) will verify this by bioreactor testing. â¢ Along with N-heterocyclic compounds, phenols, and ammonia were detected as possible inhibition sources for biological treatment processes. Total phenolic levels of 1.5-2.3 g GAE/L were measured, which could be inhibitory for both aerobic and anaerobic systems. Furthermore, the ammonia concentration of HTL aqueous was 3.9-7.1 g/L, which would be an inhibitor for anaerobic treatment units. Chapter 4 will assess biodegradability of the identical aqueous streams under aerobic and anerobic conditions. The ANOVA tests conducted in this chapter revealed that, with the exception of reaction temperature's impact on acetic acid and propionic acid concentrations and retention time's effect on phenolic compounds, HTL conditions significantly influenced (p<0.0001) all other HTL aqueous characteristics. According to these findings, Hypothesis 1 has been validated. 68 Chapter 4: Analyzing four biological treatment methods for treating hydrothermal liquefaction aqueous phase derived from sludge in municipal wastewater treatment plants 4.1 Introduction Using the existing infrastructure in a WWTP, both aerobic and anaerobic treatment could be potential pathways of HTL aqueous processing. The AD of HTL aqueous would increase the overall process energy recovery by converting organics in HTL aqueous into CH4. Aerobic treatment of HTL aqueous would consume energy, but this process could be an alternative for WWTPs that do not have anaerobic digester infrastructure in place. Compared to AD, aerobic treatment would also be less affected by the inhibitory compounds [255]. The third on-site treatment option in a WWTP is dark fermentation which could provide highly needed H2 for fuel upgrading (hydrodeoxygenation) of the bio-crude produced in the HTL process. To treat municipal sludge-derived HTL aqueous phase, mesophilic AD [107,110], adsorption [105,210], wet oxidation [106], membrane distillation [256], aqueous phase recycling [257], and ammonia removal [258] have been investigated. Despite these efforts, an economic HTL aqueous treatment method that could be applied in full-scale plants has not been identified. To the best of my knowledge, aerobic treatment of HTL aqueous has never previously been investigated. In this chapter, four on-site biological treatment scenarios were investigated with HTL aqueous samples obtained at 290-360°C reaction temperatures and 0-30 minutes retention times using dewatered mixed sludge as feedstock. The primary objective of this chapter was to identify a suitable biological method for treating HTL aqueous derived from municipal sludge, which bottlenecks the full-scale application of HTL process in WWTPs. In alignment with this objective, treatment yields from four biological treatment methods were investigated. The secondary objective of this chapter was to reveal the biodegradability levels of HTL aqueous samples obtained from various HTL conditions in Chapter 3. This objective was met through investigation of the relative biodegradability and inhibitory nature of the HTL aqueous by product produced through differing reaction temperatures and retention times. 4.2 Materials and methods 4.2.1 Experimental design and sample preparation The HTL operation was performed using a 1 L HT/HP Parr 4570 reactor (Parr Instrument Company, USA). Nine HTL runs were designed using a circumscribed composite design around the center point of 325°C â 15 min, within 290-360°C and 0-30 min ranges. Information regarding HTL operation and experimental design were given in Chapter 3, Section 3.2.3 and 3.2.4.12, 69 respectively. Four scenarios were tested as potential on-site biological treatments of HTL aqueous in a WWTP. In scenario I, the aerobic biodegradability of HTL aqueous was studied to test if HTL aqueous could be treated aerobically along with wastewater. In scenarios II and III, mesophilic and thermophilic methane potentials and anaerobic biodegradability of HTL aqueous samples were investigated, respectively. In the last scenario (IV), the hydrogen potential of HTL aqueous was investigated in the dark fermentation process. Tested scenarios are summarized in Figure 4.1. Figure 4.1. On-site HTL aqueous treatment pathways of (a) scenario I: aerobic treatment, (b) scenario II: mesophilic anaerobic digestion, scenario III: thermophilic anaerobic digestion, and (c) scenario IV: dark fermentation. 70 4.2.2 Sample and inoculum characterization Detailed characterizations of feedstock mixed sludge and inocula used in aerobic/anaerobic tests were performed (Appendix C, Table C1). Total/volatile solids (TS/VS), total ammonia, COD, and total phenolic compound concentrations were measured following the methods given in Section 3.2.4. Total suspended solids (TSS) and volatile suspended solids (VSS) were determined following Standard Methods 2540 [212]. The VFA concentrations of HTL aqueous samples were measured using a 7890A gas chromatography â flame ionization detector (GC-FID) system (Agilent Technologies, USA) as described in Section 3.2.4.8. Ultimate analysis of mixed sludge was performed by the method described in Section 3.2.4.11. Detailed HTL aqueous characterizations are given in Chapter 3. Ash, carbohydrate, protein, and lipid content analyses were performed on all sludge types. Ash content was measured following Standard Methods 2540 [212]. Sludge samples were dried at 105°C for 24 hours before carbohydrate, protein, and lipid analyses. Carbohydrate contents were measured using the phenol-H2SO4 (DuBois) method [214]. Protein contents were calculated by multiplying the nitrogen (N) contents of the samples by 6.25. Finally, lipids were determined by hexane extraction according to Standard Methods 3540 [212]. 4.2.3 Aerobic biodegradability test (Scenario I) Batch aerobic biodegradability tests were conducted according to Standard Methods 5210D [212], using an online PF-8000 respirometer (Respirometer Systems and Applications, USA). In these tests, 600 mL borosilicate reactor bottles were used with 500 mL liquid volume. A sample (HTL aqueous) volume of 5 mL was employed in each experiment. The aerobic inoculum was acquired from a continuously fed lab-scale complete mix aerobic treatment unit being fed with synthetic wastewater and HTL aqueous for a period of 6 months at room temperature (22-23°C). In each test, sufficient inoculum was provided to maintain a ratio of 10 g CODsample/g VSSinoculum. Aerobic inoculum characterization is given in Appendix C (Table C1). Additionally, reactors were supplied with buffer (pH 7), nutrient, mineral, and nitrification inhibitor solutions, the details of which are provided in Table 4.1. Finally, CO2 absorption tubes filled with 5 mL 30% KOH were placed in the headspaces of reactors. The reactors were then sealed and placed in water baths working at 20°C. 71 Table 4.1. Buffer, nutrient, and inhibitor solutions added to biochemical oxygen demand tests. Mineral Base I Compound Contribution Compound Contribution CoCl2â 6H2O 0.062 mg Co/mL Na2MoO4â 2H2O 0.0020 mg Mo/mL FeCl2â 4H2O 1.126 mg Fe/mL NiCl2â 6H2O 0.0062 mg Ni/mL MnCl2â 4H2O 0.0139 mg Mn/mL Na2SeO4 0.0104 mg Se/mL H3BO3 0.0044 mg B/mL CuCl2â 2H2O 0.0026 mg Cu/mL ZnCl2 0.0119 mg Zn/mL Mineral Base II Compound Contribution CaCl2 5.4 mg Ca/mL MgCl2â 6H2O 2.36 mg Mg/mL Nutrient Base Compound Contribution NH4Cl 13.9 mg N/mL Na2SO4 3.4 mg S/mL KH2PO4 2.7 mg P/mL Aerobic Test Buffer Base Compound Contribution KH2PO4 Stable pH (7.0) Nitrification Inhibitor Compound Contribution C4H8N2S (N-Allylthiourea) Inhibit nitrification Throughout the test, the reactors were mixed at 600 rpm to allow oxygen in the headspace to diffuse into the liquid. The respirometer fed each reactor bottle with pure oxygen at a pressure of 500 mm H2O and recorded the gas flow at 10-minute intervals. Oxygen consumption from blank reactors (inoculum only) was subtracted from others to calculate substrate-specific oxygen consumptions. Tests continued until oxygen consumption ceased, allowing the measurement of ultimate biochemical oxygen demand (UBOD). Photos of a respirometer reactor and RSA PF-8000 respirometer system are given in Figure 4.2. The aerobic biodegradability of HTL aqueous samples was calculated using the formulation given in Eq. 4.1. Aerobic biodegradability (%) =Ultimate biochemical oxygen demand (mg/L)Chemical oxygen demand (mg/L) (4.1) 72 (a) (b) Figure 4.2. (a) Image of a 500 mL borosilicate respirometer reactor with 30% KOH filled CO2 absorption tube in the headspace and (b) RSA PF-8000 respirometer during measurement. 4.2.4 Anaerobic biodegradability test (Scenario II and III) Batch biochemical methane potential (BMP) and biodegradability tests were performed following the method suggested by Angelidaki et al. (2009). The anaerobic tests were conducted in 160 mL glass Wheaton bottles with 70 mL working volume. Mesophilic (35°C) and thermophilic (55°C) inoculum were obtained from lab-scale 3 L digesters, which were semi-continuously fed with mixed sludge obtained from Annacis Island WWTP and operated at a hydraulic retention time of 20 days for a period of 3 months. Each BMP assay was supplemented with 40 mL of inoculum and adequate amount of HTL aqueous to maintain a food/microorganism (F/M) ratio of 0.25 g CODHTL aqueous/g volatile solids (VS)inoculum. In a working volume of 70 mL, the amount of HTL aqueous for mesophilic and thermophilic AD tests ranged from 1.4-1.6 mL and 1.55-1.79 mL, respectively. As HTL aqueous contains sufficient levels of alkalinity and macro/micro nutrients, no additional supply was added [260]. A total of eight BMP assays were conducted to characterize HTL aqueous samples derived from each sludge type (duplicate HTL runs per sludge type and 4 replicate BMP assays per HTL aqueous sample). Among eight BMPs, two were used for monitoring pH and volatile fatty acid (VFA) levels (methods described in Section 3.2.4). BMP experiments were performed using Innova®44 rotary shaking incubators (New Brunswick Scientific Co. Inc., USA) set to 35°C for mesophilic and 55°C for thermophilic experiments, with 90 rpm orbital mixing. On certain days, BMP assays were taken out of the shaking incubators and cooled down to room temperature. The amount of generated biogas and the composition of biogas were 73 measured using the manometric method and Agilent 7820A gas chromatography equipped with a thermal conductivity detector (GC-TCD) (Agilent Technologies, USA), respectively. After measuring the pressure, a 500 µL gas sample was collected using a glass syringe and then injected into a GC-TCD equipped with a packed column (Agilent G3591 â 80002/80003). The carrier gas, helium, was flowing at a rate of 28 mL/min during the injection. The oven, inlet, and detector temperatures in the GC were 70°C, 120°C, and 150°C, respectively. Images of GC-TCD, shaking incubators, and 160 mL BMP bottles are presented in Figure 4.3. (a) (b) (c) Figure 4.3. Images of (a) Agilent 7820A gas chromatography - thermal conductivity detector, (b) New Brunswick Innova®44 shaking incubators, and (c) batch of BMP assay bottles. Obtained N2, CO2, and CH4 percentages in the gas samples were used for calculating methane production. The calculated gas productions were normalized for standard temperature and pressure (STP: 0°C, 1 atm). The measured gas pressure was converted to volume using Eq. 4.2, and the resulting volume was normalized for STP using Eq. 4.3. 74 âN(t) = [y(t)P(t)VhRT] â [y(t â 1)P(t â 1)VhRT] (4.2) âV(t) = âN(t)RT0P0 (4.3) Where ÎN(t) is the change in normality of CH4 between times t and t-1 in the headspace, y(t) is percent CH4 in the headspace (%), P(t) is measured gas pressure at time t (Pa), Vh is headspace volume (m3), R is universal gas constant (8.314 m3·Pa/KÂ·mol), T is temperature of the reactor at the time of measurement (K), ÎV(t) is the change in gas volume of CH4 produced between times t and t-1, T0 is standard temperature (273.15 K), and P0 is standard pressure (101325 Pa). Anaerobic biodegradability was calculated by dividing specific methane yields obtained from the HTL samples by the theoretical methane yield of 350 mL CH4/g CODdegraded at STP (Eq. 4.4). Anaerobic biodegradability (%) =mL CH4 produced/g CODadded350 mL CH4/g CODdegradedà 100 (4.4) 4.2.5 Dark fermentation test (Scenario IV) The dark fermentation test was performed following the biochemical hydrogen potential (BHP) method described by Eskicioglu et al. (2017) [261]. The test was conducted using 160 mL Wheaton bottles with 40 mL working volume. The inoculum for the BHP test was obtained from mesophilic semi-continuous flow lab-scale fermenters fed with mixed sludge for a minimum of 60 days, under 20 days hydraulic retention time. Prior to the test, the inoculum was pretreated at 90°C for 30 minutes to inactive mainly methanogenic microorganisms, sparing only bacterial spores. To each BHP test, 20 mL of pretreated inoculum, macroelements (0.8 g/L NH4Cl, 0.5 g/L KH2PO4), trace elements (1.5 g/L FeCl2, 60 mg/L H3BO3, 25 mg/L CoCl2, 117 mg/L MnSO4, 25 mg/L NiCl2, 70 mg/L ZnCl2, 15 mg/L CuCl2, 25 mg/L Na2MoO4), vitamin solution (5 mg/L vitamin H, 5 mg/L vitamin B12, 0.5 mg/L vitamin B1), and pH buffer (0.1 M 2-(N-morpholino)ethanesulfonic acid) were added. The HTL aqueous dosage in each bottle was adjusted according to 10 mg CODHTL aqueous/mg VSinoculum ratio. To maintain this F/M ratio, the volume of HTL aqueous added to BHP tests ranged from 1.94 to 2.29 mL. The control assays with glucose were also conducted to test the activity of the pretreated inoculum and reliability of the other assay conditions including F/M ratio and buffer/nutrient media. After adding all contents, the initial pH of each reactor was set to 6 by adding H2SO4. Reactors were kept in Innova®44 rotary shaking incubators (New Brunswick Scientific Co. Inc., USA) working at 35°C and 90 rpm mixing. The gas pressure was 75 monitored using a digital manometer and the gas composition was measured by injecting a 500 µL sample into Agilent 7820A GC-TCD (Agilent Technologies, USA) until the first signs of hydrogen consumption (reduction of H2%) at the headspace. GC configuration, column, and gas flow rates were identical to the AD test, shared in Section 4.2.4., with the exception of the gas carrier being switched to Argon. To calculate the amount of H2 produced and to normalize gas production for STP, Eq. 4.2 and Eq. 4.3, given in Section 4.2.4. were utilized. The inoculum activity test with glucose (positive control) resulted in 1.2±0.2 mole H2/mole glucose (146±29 mL H2/g glucose) (Appendix C, Figure C1), which agrees with literature [262,263]. An image of the BHP assay bottles is given in Figure 4.4. Figure 4.4. Biochemical hydrogen potential assays. 4.2.6 Data analysis Error bars given throughout the chapter indicate Â± standard deviation. The ANOVA test was performed on BOD5, BMP, and calculated aerobic/anaerobic biodegradability index values of HTL aqueous samples to reveal the statistical significance of the relationship between dependent variables and the HTL conditions. Data normality was assessed with the Kolmogorov-Smirnov test using OriginPro 2023b software (OriginLab Corporation, Massachusetts, USA) (Appendix C, Table C2). Experimental design and ANOVA test were explained in Section 3.2.4.12. The oxygen uptake curves from the respirometric test and the cumulative methane generation curves from mesophilic and thermophilic BMP tests were fitted to the modified Gompertz equation to estimate the reaction kinetic parameters Kinetic modeling was conducted on OriginPro 2023b software (OriginLab Corporation, Massachusetts, USA). The modified Gompertz equation used for kinetic modeling is given in Eq. 4.5. 76 BMPð¡ or BODð¡ = ð Ã exp {âexp [ð ð Ã eðà (ð â ð¡) + 1]} (4.5) Where t is specific incubation time (h), P is maximum specific methane production (mL CH4/g CODadded for BMP) or oxygen consumption (mg O2/L for BOD), Rm is maximum reaction rate (mL CH4/g CODadded·h-1 for BMP and mg O2/LÂ·h-1 for BOD), and λ is lag phase duration (h). 4.3 Results and discussion 4.3.1 Characterization of HTL aqueous samples Prior to the biodegradability experiments, a detailed characterization was performed on the HTL feedstock (Table 4.2) and HTL aqueous samples (Chapter 3). Feedstock characterization results were found in the range of previous findings in the literature [189,264]. According to the characterization results, given in Table 4.3, the HTL aqueous samples have COD values ranging from 87 g/L to 103 g/L. It is observed that COD levels dropped as the severity of the HTL process increased, potentially correlated to the carbon shifting to bio-crude and gas phases due to recombination and gasification reactions, respectively. Aqueous characterization also revealed that a significant contributor to the measured CODs was VFAs in the range of 10 to 14.3 g/L. VFA concentrations were found to be favorable for AD, however possible inhibitory compounds, ammonia, and phenolics were also found in high concentrations. Furthermore, HTL aqueous contains N-heterocyclic compounds, which are known to inhibit AD cultures [49]. More detailed characterization of HTL aqueous samples used in this chapter can be found in Chapter 3. Table 4.2. Characterization data of mixed municipal sludge used as feedstock in hydrothermal liquefaction process (n=3). Parameters Mixed municipal sludge pH 5.97±0.04 Total solids (%, w/w) 4.03±0.02 Volatile solids (%, w/w) 3.54±0.02 Total COD (mg/L) 64106±3565 Soluble COD (mg/L) 9846±1316 Carbohydrates (%, db) 27.8±2.4 Proteins (%, db) 33.8±0.4 Lipids (%, db) 12.6±0.25 Ash (%, db) 12.3±0.56 C (%, db) 47.6±0.14 H (%, db) 6.82±0.07 N (%, db) 5.41±0.07 S (%, db) 0.91±0 O (%, db) 28.8±0.14 COD: Chemical oxygen demand, db: Dry basis 77 Table 4.3. Characterization results of HTL aqueous samples obtained from mixed sludge at various conditions. Reaction temperature (°C) Retention time (min) Chemical oxygen demand (mg/L) Total ammonia (mg/L) Total phenolics (mg GAEa/L) Total VFAb (mg/L) 290 15 100286±3729c 4506±255 2144±73 12442±279 300 4.4 101759±4497 3929±47 2299±47 9963±103 300 25.6 102845±3839 4501±50 1649±58 12398±102 325 0 96486±548 4861±106 1564±8 14267±150 325 15 101216±2634 5979±37 1665±23 12728±296 325 30 90514±1535 5893±5 1711±37 12447±341 350 4.4 90126±2961 6077±29 1657±22 11526±135 350 25.6 93771±658 7052±50 1450±0 11918±156 360 15 87412±4168 6351±34 1538±16 12656±227 aGallic acid equivalent, bAcetic acid + propionic acid + butyric acid, c± values represent standard deviation of 6 replicates. 4.3.2 Scenario I â Aerobic treatment Respirometric BOD5 tests were performed with all HTL aqueous samples to measure aerobic biodegradability. The oxygen uptake and oxygen uptake rates for 5 days are presented in Figure 4.5. The measured BOD5 values were in the range of 41 g/L and 54 g/L for HTL aqueous samples obtained at 350°C-25.6 min and 325°C-0 min (no retention at 325°C), respectively. The highest oxygen-consuming HTL aqueous sample, 325°C-0 min, had the highest VFA concentration and the second highest carbohydrate and protein concentrations of all the HTL aqueous samples [260]. Aligned with the observed COD trend, BOD5 levels also decreased as the reaction severity increased. This indicates that aerobically biodegradable compounds also shifted to other phases of the HTL as the severity increased. When HTL reaction conditions became more severe (325°C-0 min to 360°-15 min), the decreases in the HTL aqueous COD and BOD5 were found to be approximately 9.4% and 20%, respectively. Due to the BOD5 level decreasing at a greater rate than the COD levels, it can be said that the biodegradability would decrease at HTL conditions more severe than 325°C-0 min. This could be a result of an increased gasification reaction rate after 325°C causing smaller molecules to decompose and form H2, CO2, CO, and CH4 [81]. On the other hand, larger molecules that are challenging to degrade biologically, such as N-heterocyclics and phenolics, were less affected by the increasing HTL severity [260]. Therefore, the decrease in BOD5 was higher than COD when the reaction temperature increased from 325°C to 360°C. 78 Figure 4.5. Respirometric oxygen uptake and uptake rate results of municipal sludge derived HTL aqueous samples obtained under various temperature and retention times (shaded regions indicate error bands with standard deviation). 79 No acute inhibition was observed during the aerobic tests at a concentration of 1% (v/v) HTL aqueous. However, there was a slight time difference observed between the peaking of oxygen uptake rate graphs given in Figure 4.5. For HTL aqueous samples obtained around 350-360°C, there was a clear increase in the lag phase indicating some inhibition. For example, at 360°C-15 min condition, the oxygen uptake rate peak was observed at the 20th hour, which is later than those obtained below 350-360°C. During the aerobic test for HTL samples of 350°C-25.6 min and 360°C-15 min conditions, a small secondary peak was observed on the oxygen uptake rate graphs, which indicates a shift in microbial structure or substrate type. All these findings support the observation that HTL aqueous samples obtained under 350-360°C conditions caused inhibition of the aerobic microbial community. The inhibition may be caused by the increased dissolved bio-crude in the aqueous phase caused by the enhanced solvent properties of water at higher temperatures. Most of the non-biodegradable compounds forming in hydrothermal reactions have strong bonds that can withstand hydrothermal conditions. As these molecules are also uncommon in the environment, microorganisms are not equipped with enzymes that can degrade most of them. Hence, a significant portion of COD in HTL aqueous samples was not utilized by microorganisms in the five-day respirometric test. To reveal the kinetic parameters, oxygen uptake curves were fitted using the Modified Gompertz equation. The obtained kinetic parameters are given in Table 4.4.Table 4.4. Modified Gompertz model kinetic parameters of five-day biochemical oxygen demand of HTL aqueous samples. HTL condition P (mg O2 L-1) Rm (mg O2 L-1 h-1) λ (h) R2 Adj-R2 290°C-15 min 51823 922 0.91 0.9816 0.9815 300°C-4.4 min 46870 831 1.27 0.9829 0.9828 300°C-25.6 min 49686 914 1.81 0.9804 0.9803 325°C-0 min 51681 833 0.98 0.9728 0.9727 325°C-15 min 47400 884 1.05 0.9802 0.9802 325°C-30 min 45748 697 1.03 0.9757 0.9756 350°C-4.4 min 43325 693 5.21 0.9859 0.9859 350°C-25.6 min 39139 634 3.34 0.9814 0.9813 360°C-15 min 40514 777 8.03 0.9822 0.9822 The trend observed in kinetic parameters supported the inhibition caused by the HTL samples obtained around 350-360°C conditions. The maximum reaction rate (Rm) and lag phase (λ) parameters indicate that adapting to the substrate and aerobic degradation became more challenging for microorganisms as the HTL reaction severity increased. This effect was most 80 pronounced for HTL aqueous samples obtained at 360°C-15 min with a lag phase of 8 hours. In general, for batch aerobic tests, there was no significant inhibition leading to process failure observed. However, there were signs of inhibition for HTL aqueous samples obtained at severe conditions (350-360°C) that could be increasingly problematic in a downstream biological treatment system. Aerobic degradation of HTL aqueous should be tested on a continuous-flow system with various HRT and SRT combinations to better understand reactor stability and the potential chronic inhibitory effect. 4.3.3 Scenario II â Mesophilic anaerobic treatment The cumulative mesophilic (35°C) methane production from the HTL aqueous samples plateaued in 15 days (360 hours), as observed in Figure 4.6. There was no significant methane production between days 15 and 25. As the reaction media of the HTL process, HTL aqueous undergoes extensive chemical changes under near-supercritical temperatures. As a result, prior to BMP tests, the feedstock (mixed sludge) had already undergone significant thermal hydrolysis and decomposition during the HTL runs. The hydrolysis process, which is mostly the rate-limiting step in AD, and the decomposition process involve the cleavage of chemical bonds in complex organic molecules, into simpler soluble forms, such as VFAs, protein fragments (amino acids, amino acid esters, N-heterocyclic compounds, amines, and amides), and monosaccharides [260]. If not inhibitory, these hydrolyzed and decomposed compounds are expected to be more readily bioavailable and can be metabolized by microorganisms, leading to the rapid consumption of substrate and the subsequent cease of methane production in a BMP test. Figure 4.6. Mesophilic biochemical methane potential results of various HTL aqueous samples (n=6). 81 The measured specific methane production results fell within the range of 151-179 mL/g COD, with the exception of the HTL aqueous samples obtained at the 360°C and 15 min condition. The similarity in methane potential among the HTL aqueous samples could be attributed to their comparable VFA and COD concentrations, as shown in Figure 3.6. VFAs are significant substrates in the AD process and represent one of the major organic molecule groups present in the HTL aqueous samples, so VFA concentrations play a crucial role in determining methane potential. This means that in AD, HTL aqueous samples with the same VFA concentrations should theoretically have similar methane potentials. Using the mesophilic BMP results, it was calculated that 36.3% (300°C, 4.4 min) to 51.9% (350°C, 25.6 min) of the produced methane directly originated from the utilization of VFAs (C2-C5 VFAs) in theory (calculations provided in Appendix C, Table C3). The VFA and pH levels in mesophilic BMP tests are given in Figure 4.7 and Figure 4.8, respectively. Due to the high alkalinity of HTL aqueous samples, the pH was in a range of 7.53-7.99, without buffer addition. The lowest pH of 7.53 was observed in BMP reactors set up with HTL aqueous of 360°C-15 min at day 25 as a result of VFA accumulation. Results showed low VFA concentrations at day 25 indicating successful completion of the batch digestion tests, with the exception of the BMP test treating the aqueous derived at 360°C-15 min. In the failing test treating HTL aqueous sample obtained at 360°C-15 min, the VFA results revealed that bacteria started to perform acidogenesis and acetogenesis after a lag phase between the 10th and 15th day. This caused the accumulation of VFAs up to 991 mg/L at day 25. As there was no methane production along with accumulating VFAs, it can be concluded that the acetoclastic methanogenic archaea were inhibited by HTL aqueous obtained at 360°C-15 min. Some N-heterocyclic and phenolic compounds that HTL aqueous contains could be the cause of this inhibition. The GC-MS results for the HTL aqueous samples indicated that there was a high abundance of these N-heterocyclic and phenolic compounds in the sample obtained at 360°C, including pyridines (pyridine, 3-aminopyridine, 3-picoline, 1-methyl-2(1H)-pyridinethione), pyrrolidinones (2-pyrrolidinone, 1-methyl-2-pyrrolidinone, 1-ethyl-2-pyrrolidinone), piperidinones (δ-valerolactam, 1-ethyl-6-methyl-3-piperidinone), pyridinols (3-pyridinol, 6-methyl-3-pyridinol, 2-6-dimethyl-3-pyridinol), and phenolic compounds (phenol, p-cresol, aminophenol) [260]. Comparing the HTL aqueous sample obtained under the 360°C-15 min condition with the other samples, the most significant difference was in N-heterocyclic compounds, specifically pyridines, and pyrrolidinones. As mentioned by Fernandez et al. (2018), these compounds have also been found in HTL aqueous derived from algae and are known to cause inhibition in semi-continuously fed anaerobic digesters [265]. 82 Figure 4.7. Volatile fatty acid concentrations of mesophilic biochemical methane potential reactors throughout the test (n=2). 83 Figure 4.8. pH of mesophilic biochemical methane potential reactors throughout the test. Pham et al. (2013) reported cytotoxicity for all the mentioned N-heterocyclic and phenolic compounds, with aminophenol identified as the most cytotoxic compound [243]. In their study, OâConnor and Young (1996) discovered that 2-aminophenol and 3-aminophenol exhibited significant inhibitory effects in AD, whereas the isomer 4-aminophenol did not show any inhibitory behavior [266]. Li et al. (2019) used zeolite to adsorb inhibitors from Chlorella-derived HTL aqueous and achieved a 117% methane production increase by removing N-heterocyclic compounds (F/M ratio of 1 g COD/g VS, COD: 2 g/L) [111]. Based on these findings, it can be concluded that aminophenol, specifically the isomer 2-aminophenol, was identified as one of the inhibitors present in the HTL aqueous samples. Kinetic modeling was performed on the BMP results by fitting the curves with the Modified Gompertz equation. The calculated ultimate methane productions (P), maximum biochemical reaction rates (Rm), and lag phases (λ) were provided in Table 4.5. Notably, no significant variation was observed in the ultimate methane productions and maximum reaction rates. The impact of skipping hydrolysis in AD was evident in the calculated lag phases. Since the substrate had already undergone thermochemical hydrolysis, there was no lag phase observed in the mesophilic BMP tests. Due to the early acute process failure (Figure 4.6), no kinetic parameters could be calculated for the AD of HTL aqueous sample obtained at 360°C-15 min.84 Table 4.5. Modified Gompertz model kinetic parameters of mesophilic biochemical methane potential results of HTL aqueous samples. HTL condition P (mL CH4/g COD) Rm (mL CH4/(g CODÂ·h)) λ (h) R2 Adj-R2 290°C-15 min 149 3.41 0 0.8800 0.8201 300°C-4.4 min 138 3.62 0 0.9087 0.8630 300°C-25.6 min 148 3.97 0 0.9003 0.8505 325°C-0 min 155 3.90 0 0.9041 0.8561 325°C-15 min 142 3.90 0 0.9187 0.8780 325°C-30 min 153 4.18 0 0.9098 0.8647 350°C-4.4 min 140 3.48 0 0.8780 0.8170 350°C-25.6 min 135 4.04 0 0.9281 0.8922 360°C-15 min 149 3.41 - - - In summary, mesophilic AD of HTL aqueous samples that were produced at lower temperatures than 360°C was possible. With the exception of the 360°C-15 min and 350°C-25.6 min samples, approximately 50% of theoretical methane production yield at STP (350 mL CH4/g CODdegraded) was reached for all samples (Figure 4.6). The yield decreased to 43% for the 350°C-25.6 min sample and 0% for 360°C-15 min, indicating that toxicity of HTL aqueous increases for more severe conditions than 350°C-25.6 min. It should be noted that the AD results in this study represent the ultimate methane potential of HTL aqueous samples by using a low F/M ratio of 0.25 g CODHTL aqueous/g VSinoculum (â2 g COD/L) to avoid inhibition. Although the batch tests were mostly successful, the successful performance under this low F/M ratio does not guarantee the stability of an HTL aqueous-fed continuous-flow anaerobic digester. Unlike batch systems, continuous-flow reactors are susceptible to chronic inhibition that can occur due to microorganisms being exposed to inhibitory compounds for a prolonged time or due to the accumulation of certain inhibitory compounds. The BMP results in this study only indicate the maximum methane yield that could be reached from municipal sludge derived HTL aqueous in a controlled anaerobic digester if no inhibition or organic overloading occurs. The results also provide insight into which HTL (temperature/retention time) conditions to avoid due to the increased toxicity of HTL aqueous to beneficial anaerobic cultures. Characterization of HTL aqueous is highly dependent on the substrate type and reaction conditions [49]. Carbohydrate-rich feedstocks, such as lignocellulosic biomass, produce an HTL aqueous phase that contains higher VFA concentrations, whereas protein-rich feedstocks produce more nitrogen containing molecules and higher toxicity in the HTL aqueous [9]. 85 Therefore, there could be differences in HTL aqueous biodegradability and toxicity findings by feedstock. Wang et al. (2021) reported failing BMP tests at pH 7, 7.5, and 8 when the test was set up at a ratio of 1 g CODHTL aqueous/g VSinoculum using municipal sludge derived HTL aqueous [258]. However, they produced 225 mL CH4/g COD in a mesophilic BMP test when the HTL aqueous was treated with biochar and struvite precipitation. Chen et al. (2020) explained how they treated cornstalk derived HTL aqueous with 62% COD removal at mesophilic conditions using a continuously fed reactor receiving HTL aqueous load of 2.5 g COD/d [95]. Using the same feeding rate in a sequencing batch reactor, Usman et al. (2019) reported that they treated 55% of COD of HTL aqueous obtained at 300°C-1 h when municipal sludge with 15.5% initial TS was used as feedstock [110]. Summarizing both the results of this study and previous results in the literature, it can be concluded that, without pretreatment, HTL aqueous can be anaerobically digested in mesophilic conditions if diluted to a level adequate to prevent inhibition on the anaerobic culture. Another alternative could be the co-digestion of HTL aqueous with more biodegradable co-substrates to provide dilution. In a WWTP, sludge centrate obtained from dewatering sludge for HTL process or even primary treatment effluent could be used to dilute HTL aqueous before sending it to anaerobic digesters. The methane yield of HTL aqueous could be increased by incorporating pretreatment. Two HTL aqueous pretreatment methods in literature have shown promising results for detoxification, advanced oxidation, and adsorption. The first viable pretreatment method, adsorption, has the advantage of targeting several known inhibitory compounds, like phenolics, while not adsorbing VFAs [105]. Adsorbents can also be added into AD systems which results in simultaneous adsorption and AD. Adsorbent-added HTL aqueous BMP tests were reported to produce up to 10.4% and 200% more methane than control (no adsorbent) by Wang et al. (2022) and Si et al. (2019), respectively [112,267]. The second viable pretreatment method, advanced oxidation, has the ability to convert complex molecules into VFAs and is more economically feasible for full scale application as compared to adsorption. Silva Thomsen et al. (2022) performed wet air oxidation to HTL aqueous at 350°C for 10 minutes and reported that total organic carbon was reduced by 80%, while VFAs increased by 100% [106]. Furthermore, the wet air oxidation process efficiency was reported to increase with the use of inexpensive catalysts such as CeZr, WO3, and ZrO2 [115]. Ozonation is also an alternative advanced oxidation process that can oxidize phenols, cyclic hydrocarbons, and partially N-heterocyclic molecules into VFAs [268]. 86 4.3.4 Scenario III â Thermophilic anaerobic treatment Similar to the mesophilic BMP tests, the thermophilic BMP tests (55°C) took only 15 days to reach most of the methane production. The thermophilic BMP results of HTL aqueous samples are given in Figure 4.9. The inhibitory effect of HTL aqueous was much more prominent for the thermophilic BMP test results. This was expected as thermophiles are more susceptible to inhibition compared to mesophiles due to reduced microbial diversity at elevated temperatures [269]. BMP tests performed with HTL aqueous obtained at 360°C-15 min were completely inhibited and did not show a discernable methane volume throughout the test. The rest of the samples had a two-step cumulative methane production curve, which is an indication of inhibition [270]. With the exception of the HTL aqueous obtained at 360°C-15 min, specific thermophilic methane productions ranged from 102 mL/g COD (350°C-25.6 min) to 122 mL/g COD (325°C-0 min). These values correspond to 29-35% anaerobic biodegradability (calculated by Eq. 4). It is important to note that thermophilic BMP results were, on average, 32% lower than the mesophilic BMP results. The kinetic parameters of the thermophilic BMP test are given in Table 4.6. Figure 4.9. Thermophilic biochemical methane potential test results of HTL aqueous samples (n=6). Thermophilic AD has some key advantages over mesophilic AD, such as higher TS removal, pathogen elimination, and higher reaction rate. However, when HTL aqueous derived at high temperature/pressure is used as the substrate in AD, no TS removal and pathogen elimination is required. Furthermore, due to the substrate being hydrolyzed thermochemically, thermophilic AD also loses its high-rate advantage. In order to determine the reaction kinetics, the thermophilic BMP curves were fitted using the Modified Gompertz equation. According to the reaction kinetics given in Table 4.6, similar to the mesophilic tests, there was no lag phase detected in thermophilic BMP tests. The maximum reaction rates of the thermophilic tests compared to the mesophilic 87 tests were ten-fold lower, providing another indication of inhibition. As thermophilic AD loses the common advantages over mesophilic AD when digesting HTL aqueous, along with considering more robust mesophilic operation due to wider microbial diversity, it can be concluded that mesophilic conditions are more favorable and reliable for the digestion of HTL aqueous. These batch test results indicate that in a continuously fed system implemented at full-scale, mesophilic AD would be a better option in terms of biochemical process stability as acute inhibition was observed in thermophilic AD. This inhibition causes significantly lower organic removal efficiencies and methane production yields. Chen et al. (2020) reported that the treatment yield of cornstalk-derived HTL aqueous-fed reactors decreased to 45% in thermophilic conditions versus 62% in mesophilic conditions [95]. In conclusion, based on ultimate methane production, treatment efficiency, and resistance against inhibition, mesophilic AD showed better overall performance compared to thermophilic AD for the treatment of HTL aqueous. Table 4.6. Modified Gompertz model kinetic parameters of thermophilic biochemical methane potential results of HTL aqueous samples. HTL condition P (mL CH4/g COD) Rm (mL CH4/(g CODÂ·h)) λ (h) R2 Adj-R2 290°C-15 min 124 0.27 0 0.9310 0.8850 300°C-4.4 min 127 0.27 0 0.9540 0.9234 300°C-25.6 min 113 0.25 0 0.9576 0.9293 325°C-0 min 137 0.30 0 0.9444 0.9073 325°C-15 min 126 0.33 0 0.9609 0.9348 325°C-30 min 123 0.31 0 0.9582 0.9304 350°C-4.4 min 116 0.27 0 0.9708 0.9513 350°C-25.6 min 110 0.29 0 0.9713 0.9521 360°C-15 min - - - - - The VFA concentrations and pH levels in thermophilic BMP tests are given in Figure 4.10 and Figure 4.11, respectively. The pH range in the thermophilic BMP test was 7.78-8.15. There was a slight increase in pH on the second day, possibly due to the hydrolysis of soluble proteins in HTL aqueous causing ammonia concentration to increase. The VFA results showed the presence of significant levels of propionic acid during the initial ten days in all bottles. Similar to the mesophilic BMP test, except for the 360°C-15 min sample, VFAs were consumed in 25 days. In the BMP of HTL aqueous obtained at 360°C-15 min, VFA concentrations continuously increased throughout the duration of the test. Nonetheless, this increase occurred at a notably slower pace compared to the mesophilic BMP test. This indicates inhibition of bacteria that caused slower acidogenesis and acetogenesis rates. Also, no biogas generation, along with VFA accumulation, showed that thermophilic archaea were completely inhibited by HTL aqueous. 88 Figure 4.10. Volatile fatty acid concentrations of thermophilic biochemical methane potential reactors throughout the test (n=2).89 Figure 4.11. pH of thermophilic biochemical methane potential reactors throughout the test. 4.3.5 Scenario IV â Dark fermentation The dark fermentation test was performed by BHP assays using a thermally pretreated inoculum. Due to the elimination of the majority of both bacteria and archaea by pretreatment, this test relies on spore-forming bacteria, such as Clostridium [262], to start consuming the substrate and producing hydrogen. In this study, the amount of HTL aqueous in the BHP reactors was approximately 40 times higher than in the BMP tests. This is required by the method to detect hydrogen production more accurately, provide sufficient substrate to bacteria, and suppress the growth of hydrogen-consuming archaea which grow at a much slow rate. However, the HTL aqueous caused inhibition, and the recovery of hydrogen-producing bacteria from the spores did not occur. Therefore, no hydrogen potential was detected for the samples. The second reason for the failure of the BHP tests could be the substrate type. Si et al. (2019) mentioned that glycerol in the HTL aqueous could be used as a BHP substrate. Furthermore, simple sugars are the ideal substrate for BHP production. However, HTL aqueous samples used in this study contain 421-688 mg/L glycerol and 229-961 mg/L carbohydrates, equivalent to only 1-2% of their COD concentrations [260]. Xu et al. (2022a) reported that HTL aqueous can have an inhibition effect equivalent to 100 μg/ml streptomycin on bacteria [109]. As biochemical hydrogen production is performed by 90 bacteria, it can be concluded that sludge-derived HTL aqueous generated at 290-360°C is not suitable for biochemical hydrogen production due to its low substrate concentration and high inhibition. 4.3.6 ANOVA test and regression analysis The statistical significance between the independent variables of HTL reaction conditions (reaction temperature and retention time) and biodegradability responses were investigated using ANOVA test, given in Table 4.7. According to the ANOVA results, the direct effect of reaction temperature, retention time, and their interaction (AB) were found significant on BOD5 and the ratio of BOD5/COD. The least effective term on the aerobic biodegradability responses was the interaction between reaction temperature and quadratic effect of retention time (AB2) that was insignificant (95% confidence interval). For mesophilic and thermophilic anaerobic biodegradability responses, the effect of HTL retention time was found insignificant. The ANOVA test results revealed that the anaerobic biodegradability of HTL aqueous samples primarily varies with the HTL temperature as highly significant terms associated with the HTL reaction temperature (A, A2, AB2, A2B2). Table 4.7 ANOVA test results for aerobic/anaerobic biodegradation assessment for HTL aqueous samples. Parameter Terms A B AB A2 B2 A2B AB2 A2B2 ANOVA test p-value BOD5a <0.0001 <0.0001 0.0007 0.1221 0.2403 0.0007 0.7476 <0.0001 < 0.0001 BOD5/CODb <0.0001 0.0014 0.0006 0.2231 <0.0001 0.1654 0.3623 <0.0001 < 0.0001 Mesophilic methane potential and anaerobic biodegradability <0.0001 0.8008 0.0418 <0.0001 0.0476 0.9294 <0.0001 <0.0001 <0.0001 Thermophilic methane potential and anaerobic biodegradability <0.0001 0.1742 0.3503 <0.0001 0.5999 0.7208 <0.0001 <0.0001 <0.0001 A: HTL reaction temperature (°C), B: HTL retention time (min), aFive-day biochemical oxygen demand, bChemical oxygen demand The ANOVA test enabled the establishment of mathematical relationships between independent variables of reaction conditions and dependent variables of biodegradability responses. The generated regression equations and the fit statistics are given in Table 4.8. These equations can be used to calculate BOD5, aerobic biodegradability index, methane potentials, and anaerobic biodegradability of HTL aqueous samples using reaction conditions falling within the 91 circumscribed central composite design in this study (290-360°C, 0-30 min). As there were multiple highly significant terms in the ANOVA test, mostly associated with reaction temperature, it was possible to generate equations with high adjusted coefficient of determination (Adjusted-R2) levels of 0.9362-0.9789. Table 4.8. Regression equations of aerobic/anaerobic biodegradation assessment for HTL aqueous samples. Parameter Regression coefficients Fit statistics Intercept A B AB A2 B2 A2B AB2 A2B2 R2 Adjusted R2 Predicted R2 BOD5a -1640852 10504 179581 -1103 -16.28 -4690 1.69 28.9 -0.045 0.9711 0.9501 0.8866 BOD5/CODb -15.3 0.098 2.12 -0.01 15 Ã10-5 -0.07 2 Ã10-5 0.00041 -64Ã10-8 0.9631 0.9362 0.8837 Mesophilic methane potential 2401 -16.4 -1072 7.25 0.029 36.2 -0.012 -0.244 41Ã10-5 0.9859 0.9747 0.9494 Mesophilic anaerobic biodegradability 686 -4.68 -306 2.07 0.0084 10.3 -0.0035 -0.069 12Ã10-5 0.9859 0.9747 0.9494 Thermophilic methane potential 1424 -9.34 -791 5.26 0.0164 25.8 -0.0087 -0.17 29Ã10-5 0.9883 0.9789 0.9615 Thermophilic anaerobic biodegradability 407 -2.67 -226 1.50 0.0047 7.37 -0.0025 -0.049 8Ã10-5 0.9883 0.9789 0.9615 A: HTL reaction temperature (°C), B: HTL retention time (min), aFive-day biochemical oxygen demand, bChemical oxygen demand Mathematical predictions of the regression models are given in Figure 4.12 as contour plots, illustrating how the responses vary with alterations in reaction conditions. One factor model predictions are given in Appendix C (Figure C2). According to the modelâs predictions, reaction conditions around 290°C-17 min and 325°C-0 min points were identified as the conditions producing HTL aqueous that is easily aerobically biodegradable. As the HTL reaction severity increased, the aerobic biodegradability index decreased to values under 0.5, changing the classification of HTL aqueous from easily biodegradable to biodegradable [272]. Within the experimental design range of this study, no HTL aqueous sample was found in the refractory classification (BOD5/COD â¤ 0.3). There were also increased lag phases and decreased reaction rates for HTL aqueous samples produced at temperatures higher than 340°C in the BOD5 tests. Therefore, it can be concluded that temperatures over 340°C should be avoided in order to obtain HTL aqueous that can be aerobically treated more effectively. Similar results were found for mesophilic and thermophilic anaerobic biodegradability models. Temperatures higher than 350°C, clearly produce HTL aqueous that inhibits AD. The highest anaerobic biodegradability was detected for HTL aqueous samples produced under the conditions of 330-340°C and 0-2 min retention time, which coincided with the highest VFA concentration. 92 (a) (b) (c) (d) (e) (f) Figure 4.12. Predicted values of (a) five-day biochemical oxygen demand, (b) aerobic biodegradability index (BOD5/COD), (c) mesophilic methane potential, (d) mesophilic anaerobic biodegradability, (e) thermophilic methane potential, and (f) thermophilic anaerobic biodegradability by the regression analysis for HTL aqueous derived from dewatered municipal mixed sludge at various reaction temperatures and retention times (red dots indicate experimental points). 93 4.3.7 Advantages and disadvantages of treatment options In microbial communities, diversity is the key to utilization of a wide range of molecules as well as increased resistance to inhibition. Aerobic microorganisms have an advantage in microbial diversity due to aerobic respiration, which more efficiently produces energy through its biochemical processes and metabolizes a wider range of both organics and inorganics [273]. The microbial diversity of the four tested scenarios from most diverse to least diverse and aligned with biodegradability and resistance to inhibition is: aerobic treatment, mesophilic AD, thermophilic AD, and lastly dark fermentation [274]. Hence, it can be asserted that aerobic treatment is the least prone to inhibition while achieving the highest biodegradability. However, aerobic treatment has significant operational and environmental drawbacks, namely high energy consumption and greenhouse gas emissions. Activated sludge systems are estimated to have US$10/1000 m3 wastewater equivalent aeration cost [275]. Treating HTL aqueous along with wastewater could lead to a significant increase in aeration costs, primarily due to elevated levels of BOD. Another significant consequence of aerobic HTL aqueous treatment is the notable increase in secondary sludge generation, given that approximately 50% of the treated COD transforms into microorganism biomass. AD, in contrast, offers enhanced energy recovery and has relatively limited operating expenses compared to aerobic treatment, with its primary costs involving pumping and mixing. AD of HTL aqueous in mesophilic conditions was reported to contribute a 10.1-11.5% increase in overall process energy recovery, resulting in over 80% recovery overall. The challenge of AD for the treatment of HTL aqueous arises from the synergistic effect of multiple inhibitory thermochemical reaction products. At low concentrations, this inhibition can be avoided. However, in a full-scale application, dilution might be difficult due to the increased AD reactor size that the diluted sample will require. Due to higher inhibition, thermophilic AD should not be considered as a viable treatment method for full scale application. AD systems have also been reported to exhibit potential coupling with microalgae cultivation systems, offering an additional feedstock stream for co-liquefaction with municipal sludge [276,277]. In summary, both aerobic and anaerobic treatment scenarios have their advantages and disadvantages for full scale applications and may both be considered viable options depending on the existing infrastructure for any particular WWTP. 4.4 Summary In this chapter, four different on-site HTL aqueous treatment scenarios were studied for a WWTP. The key findings of this chapter are given below: 94 â¢ According to the biodegradability index (BOD5/COD), HTL aqueous samples obtained at reaction temperature under 340°C were found to be easily aerobically biodegradable. â¢ Increased lag phase and decreased reaction rate were observed in the aerobic treatment of HTL aqueous produced at 350-360°C, indicating minor inhibition. â¢ In mesophilic conditions, except for HTL aqueous obtained under 360°C-15 min condition, methane potential range was in the 151 mL CH4/g COD (350°C-25.6 min) and 179 mL CH4/g COD (325°C-0 min) range, representing 43-51.1% anaerobic biodegradability. â¢ Thermophilic AD was found to be much weaker against inhibition compared to mesophilic AD. In thermophilic AD, measured specific methane yields were in 103 CH4/g COD (350°C-25.6 min) and 122 mL CH4/g COD (325°C-0 min) range. These values were 32% lower than the mesophilic AD results. â¢ HTL aqueous obtained at 360°C was unsuitable for AD. The possible inhibitory compounds in this sample were identified as pyridines, pyrrolidinones, piperidinones, pyridinols, and phenolic compounds. â¢ VFA concentrations during BMP tests showed that HTL aqueous obtained at 360°C caused complete inhibition on archaea under mesophilic conditions. Under thermophilic conditions, on the other hand, not only did archaea experience complete inhibition, but bacterial activity was also significantly reduced. â¢ Municipal sludge derived HTL aqueous was found unsuitable for biochemical hydrogen production due to its composition and inhibitory nature. â¢ HTL aqueous obtained under 325°C-0 min reaction condition was found as the most biodegradable in both aerobic and anaerobic tests. These results verify the predicted biodegradability performances, based on HTL aqueous characterization, in Chapter 3. â¢ Aerobic (scenario I) and mesophilic anaerobic (scenario II) were found as the most suitable options for the biological treatment of HTL aqueous. However, further investigation of both methods using continuous flow reactors is required. Based on the findings in this chapter, HTL temperature significantly affected biodegradability, while retention time influenced only aerobic biodegradability (p < 0.0001). Additionally, microbial activity decreased with increasing HTL severity, validating Hypothesis 2. The biodegradation performance results obtained in Chapter 4 are HTL feedstock (municipal sludge type/characteristics) specific. Since municipal sludge varies depending on the plant design and operation, verification of these results for different sludge streams are needed, which will be assessed in Chapter 5. 95 Chapter 5: Characterization and biodegradability assessment of hydrothermal liquefaction aqueous phase derived from various sludge streams of different treatment plants 5.1 Introduction There have been quite informative HTL aqueous characterization studies in the literature performed on feedstocks of algae [278,279], food waste, oleaginous yeast, corn stover lignin [118], corn stover, pine [227], rice straw [280], municipal sludge [260], and municipal PS and SS [281]. These studies revealed that HTL aqueous characteristics are highly dependent on feedstock composition, especially carbohydrate, protein, and lipids ratios. Moreover, toxicity and biodegradability of HTL aqueous also vary based on substrate type and HTL reaction conditions. In Chapters 3 and 4, the variations in the characterization and biodegradability of HTL aqueous derived from mixed sludge (mixture of PS and SS at a ratio of 50:50 wt.%) within HTL temperature ranges of 290-360°C and retention time of 0-30 minutes was explored. However, PS and SS ratios in WWTPs may differ over time, depending on the operational needs of the WWTPs, which will cause a shift in the HTL aqueous characterization. In this chapter, PS and SS from two large-scale regional municipal WWTPs were mixed in various ratios and used in HTL to reveal the effect of PS to SS ratios on the HTL aqueous characteristics. Furthermore, HTL aqueous samples derived from mesophilic and thermophilic digested sludge (DS) were compared. Finally, a comprehensive biodegradability assessment was performed on HTL aqueous samples to determine their aerobic, mesophilic anaerobic, and thermophilic anaerobic biodegradability levels. Based on unsuccessful performance of aqueous in dark fermentation in Chapter 4, it was eliminated from Chapter 5. Another objective of this chapter was to correlate HTL feedstock characteristics in terms of carbohydrate, protein, lipid, ash contents (analytical methods that could not be utilized in other chapters) with their aqueous streamsâ biodegradability. 5.2 Materials and methods 5.2.1 HTL feedstock sample preparation and characterization Thickened PS and SS samples were acquired from two regional WWTPs (Annacis Island and Lulu Island plants) in British Columbia, Canada, that is referred to as WWTP 1 and WWTP 2 throughout the chapter. The PS samples were collected from primary clarifiers working at 1 h and 2.2 h detention times. At the plant, the collected PS was then thickened and screened for dewatering. Both treatment plants use trickling filters followed by aerobic suspended growth 96 systems with short retention times as biological treatment units. In both WWTPs, excess sludge from the biological system was thickened using dissolved air flotation systems. At UBCâs BTG laboratory, for both WWTPs, thickened PS and SS were blended in three different (w/w) ratios based on total solids (TS) to represent high-PS (PS:SS ratio: WWTP1: 59:41%, WWTP2: 70:30%), average (PS:SS ratio: WWTP1: 45:55%, WWTP2: 57:43%), and high-SS (PS:SS ratio: WWTP1: 31:69%, WWTP2: 39:61%) conditions. Blending ratios were determined according to the annual data from each WWTPs, using 95th percentiles (top and bottom 5% of data points were excluded from the distribution). Blended sludge samples were first centrifuged at 4000Ãg for 5 min and then at 20000Ãg for 10 minutes to reach 20 wt.% TS. Thermophilic and mesophilic DS from WWTP 1 and WWTP 2 were obtained as dewatered cakes with 26 wt.% and 23.3 wt.% TS concentrations, respectively, and used as received in the HTL runs. Dewatering of digestates was conducted by centrifuging at 4000Ãg after addition of 9-12.7 kg/ton Zetag® 7563 (BASF SE, Germany) polymer on site. Total and volatile solids concentrations of various HTL feedstock (mixed sludge and DS) were tabulated in Appendix D (Table D2). Finally, carbohydrate, lipid, and protein content of the feedstock mixed sludge were measured using the methods described in Section 4.2.2. 5.2.2 Bench-scale HTL operation The HTL operation was performed using a Parr 4570 1 L bench scale batch HT/HP reactor (Parr instrument company, USA). Detailed HTL procedure was provided in Section 3.2.2. HTL runs were conducted at 332°C and 16.9 min retention time, as determined to be the optimal condition for bio-crude production [10]. Ramping up to 332°C required approximately 1 h in each HTL run, and cooling to room temperature took 2 h, starting from the end of retention time. Reactor contents were collected and filtered using a 0.45 µm membrane to separate HTL aqueous from bio-crude and hydrochar. After this separation, remaining particulate matter-free HTL aqueous samples were stored at 4°C until used in chemical characterization (dissolved) and biodegradability tests. HTL operation was conducted in duplicate for each sludge sample and obtained aqueous samples were used independently in the characterization and biodegradability tests to obtain more data points. 5.2.3 Chemical characterization of HTL aqueous Chemical characterization analyses of COD, total ammonia (NH3), orthophosphate phosphorus (PO4-3-P), total alkalinity, total sugar, total phenolic compounds, and pH were performed following the methods provided in Section 3.2.4. 97 Short-chain and medium-chain fatty acids (SCFA and MCFA) concentrations were measured by a 7890A GC-FID system (Agilent Technologies, USA) equipped with a 19091 HP-FFAP column (25 m Ã 0.32 mm Ã 0.5 µm) (Agilent Technologies, USA) as described in Section 3.2.4.8. HTL aqueous samples were extracted in dichloromethane prior to quantification of selected N-heterocyclic compounds and ketones in GC-MS analysis. Extracts were analyzed using Agilent 8860 GC equipped with DB-5 Ultra column (30 m Ã 0.25 mm Ã 0.25 µm), and 5977B MS detector (Agilent Technologies, USA). Helium was used as carrier gas in GC-MS analysis. The mass spectrometer was run in full scan mode (40 to 450 atomic mass unit). 5.2.4 Aerobic biodegradability tests The determination of aerobic biodegradability was performed as ultimate biochemical oxygen demand (UBOD) tests according to the Standard Methods 5210D [212] using an online (real-time) RSA PF-8000 respirometer (Respirometer Systems and Applications, USA). Complete information on the respirometric BOD measurements and aerobic biodegradability index calculations can be found in Section 4.2.3. 5.2.5 Anaerobic biodegradability tests The anaerobic biodegradability of HTL aqueous samples was measured via biochemical methane potential (BMP) assays for mesophilic and thermophilic conditions following the method described in Section 4.2.4. Both mesophilic (35°C) and thermophilic (55°C) inoculum were obtained from lab-scale 3 L digesters, which were semi-continuously fed with mixed sludge taken from WWTP1 and operated at a hydraulic retention time of 20 days for a period of 3 months. In a working volume of 70 mL, the amount of HTL aqueous for mesophilic and thermophilic AD tests ranged from 1.45-2.15 mL and 1.38-2.05 mL, respectively. HTL aqueous had sufficient levels of micro- and macro-nutrients to support biological activity, thus no additional nutrient was added. Likewise, no buffer was added due to the high alkalinity of HTL aqueous. Detailed information regarding the incubation, gas measurements, pH and VFA monitoring, and anaerobic biodegradability calculations are given in Section 4.2.4. 5.2.6 Data analysis Error bars given throughout the chapter indicate Â± standard deviation. To evaluate the reaction kinetic parameters of maximum specific methane production (P) (mL CH4/g CODadded for BMP) or oxygen consumption (P) (mg O2/L for BOD5), maximum reaction rate (Rm) (mL CH4/g 98 CODadded for BMP and mg O2/L for BOD5), and lag phase duration (λ) (h). More detail regarding the kinetic modeling was given in Section 4.2.6. A Pearson correlation analysis was conducted to identify the level of correlation between HTL feedstock characteristics, HTL aqueous composition, and HTL aqueous biodegradability. The Pearson correlation coefficients were calculated on OriginPro 2023b software (OriginLab Corporation, Massachusetts, USA). 5.3 Results and discussion 5.3.1 Sludge characterization The HTL feedstock characterization results, in terms of ash, carbohydrate, protein, and lipid content, for mixed sludge (high PS, average, and high SS cases) and anaerobically digested sludge are presented in Table 5.1 for both WWTPs. It was revealed that 82-88% of mixed sludge and 77% of digested sludge consisted of carbohydrates, proteins, lipids, and ash on a dry basis. Sum of carbohydrate, protein, lipid, and ash content in municipal sludge was around 80-90% of total solids, similar to the findings in the literature [189,264]. This is due to organic molecules that are not classified as carbohydrates, proteins, or lipids. These molecules are mostly short/medium chain fatty acids, vitamins, and fibers. Based on the annual plant operation data, the mixed sludge in WWTP 1 exhibited a SS ratio of up to 69%, while WWTP 2 showed a higher PS ratio, reaching up to 70%. In the sludge composition of WWTP 1, the carbohydrate content increased in correlation with the PS ratio, while the protein content increased corresponding to the SS ratio. This was expected as SS has higher protein content than the PS due to its excess biomass comprised of aerobic microorganisms. The fluctuation in PS:SS ratio in WWTP 1 was determined to have no significant impact (p-value = 0.2765) on the lipid content in the municipal sludge. For mixed sludge of WWTP 2, on the other hand, the lowest PS content was 39%, which is 9% higher than that of WWTP 1, causing much less fluctuation in carbohydrate content. Furthermore, there was a clear increase in protein and a decrease in lipid content as the SS ratio increased in mixed sludge. The municipal sludge characterization in this study aligns with the results reported in the literature [189,264]. Anaerobically digested sludge feedstocks were predominantly composed of ash and proteins due to approximately half of the organic material already utilized in AD. Carbohydrate and lipid levels were significantly lower (average carbohydrate: 9.2% and average lipid: 6.8%) compared to mixed sludge feedstocks (average carbohydrate: 26.5% and average lipid: 14.4%). Despite DS having the drawback of lower organic matter content, unlike mixed sludge, it 99 possesses the advantage of being easily dewatered to higher solid levels. With protein content close to mixed sludge at high SS cases (38.4%), it was determined that mesophilic DS had higher organic matter levels when compared to thermophilic DS. In comparison to mixed sludge, both DS samples have the potential to yield a significant amount of hydrochar in the HTL process, due to their high ash content. Table 5.1. Composition of sludge feedstocks used in hydrothermal liquefaction runs (n=3). Sludge origin Sludge type Carbohydrates (%, db.) Proteins (%, db.) Lipids (%, db.) Ash (%, db.) WWTP 1 Mixed sludge High PS case PS:SS=59:41%* 31.9 ± 3.5** 31.9 ± 0 13.6 ± 0.1 8.9 ± 0.2 Mixed sludge Average case PS:SS=45:55%* 25.9 ± 2.8 35.6 ± 0 13.7 ± 0.5 10.2 ± 0.2 Mixed sludge High SS case PS:SS=31:69%* 18.7 ± 1.6 39.1 ± 0.5 13.2 ± 0.2 11.2 ± 0 Digested sludge Thermophilic 10.2 ± 0.8 29.4 ± 0 7.9 ± 0 29.2 ± 0.2 WWTP 2 Mixed sludge High PS case PS:SS=70:30%* 29.3 ± 1.6 27.2 ± 0.5 17.5 ± 0.1 8.7 ± 1 Mixed sludge Average case PS:SS=57:43%* 28.7 ± 1.1 34.1 ± 0.5 15.4 ± 0.1 9.5 ± 0.2 Mixed sludge High SS case PS:SS=39:61%* 24.8 ± 2.8 39.4 ± 0 13.1 ± 0.2 10.8 ± 0.1 Digested sludge Mesophilic 8.1 ± 0.9 38.4 ± 0.5 5.6 ± 0.1 24.5 ± 0.2 WWTP: Wastewater treatment plant, db.: Dry basis, PS: Primary sludge, SS: Secondary sludge *: Total solids-based ratio, **: Data represent arithmetic average Â± standard deviation of 3 replicates. 5.3.2 Chemical characterization of HTL aqueous Characterization results of municipal sludge from two different WWTPs showed significant differences, suggesting that the HTL aqueous generated in these treatment plants would also show variations. In the HTL process, organics that do not form bio-crude or hydrochar, do not recombine into bio-crude, and are not gasified into HTL gas, remain in the reaction media, HTL aqueous. A detailed chemical characterization was performed to uncover changes in the HTL aqueous characteristics depending on the substrate composition (Figure 5.1). COD is the most effective analysis for quantification of the organic content in HTL aqueous samples. COD levels for HTL aqueous sample derived from mixed sludges (dewatered to 20% wt. TS) of WWTP 1 and WWTP 2 were in the 75700-86300 mg/L and 59070-70940 mg/L ranges, respectively. A noticeable upward trend in COD was observed as the SS ratio in the mixed sludge increased with the exception of WWTP 1 high SS case. Higher COD levels were caused mostly by protein degradation products of N-heterocyclic compounds and VFAs. 100 (a) (b) (c) (d) (e) (f) Figure 5.1. Chemical characteristics of (a) chemical oxygen demand, (b) protein, total ammonia, (c) fatty acids, (d) orthophosphate-P, total sugar, (e) total alkalinity, pH, and (f) total phenolic compounds of HTL aqueous derived from various sludge types (n=6). High PS, average, and high SS correspond to PS:SS mixing ratios (% TS wt.) of 59:41%, 45:55%, 31:69% and 70:30%, 57:43%, 39:61% for WWTP 1 and WWTP 2, respectively. 101 Previous findings have consistently highlighted these as the two most abundant organic compound groups in HTL aqueous derived from municipal sludge [260]. HTL aqueous generated from municipal sludge of WWTP 2 exhibited significantly lower COD compared to WWTP 1. The COD levels at average cases were 86300 mg/L for WWTP 1 and 62200 mg/L for WWTP 2. The most likely cause of this COD difference was the carbohydrate contents of the feedstocks. WWTP 1 mixed sludge showed a substantial difference in carbohydrate content (18.7-31.9%) depending on the PS:SS ratio, whereas carbohydrate content did not show as much variation in WWTP 2 municipal sludge (24.8-29.3%). The impact of the feedstock carbohydrate content on HTL aqueous COD becomes evident when comparing the HTL aqueous phases of lignocellulosic feedstocks with those of protein-rich feedstocks [49]. Lignocellulosic feedstocks (carbohydrate-rich) mostly degrade into VFAs and generate less N-heterocycles in HTL aqueous. The COD concentrations in HTL aqueous samples derived from mesophilic and thermophilic DS were measured as 69200 mg/L and 65200 mg/L, respectively. Despite DS feedstocks having TS concentrations of 26% and 23.3 wt.%, their organic contents in HTL (18.4 wt.% and 17.6 wt.% VS in Appendix D, Table D2) were similar to mixed sludge (16.9-18 wt.% VS in Appendix D, Table D2), resulting in HTL aqueous COD levels comparable to mixed sludge-derived HTL aqueous. The concentrations of dissolved proteins and total ammonia in HTL aqueous, as presented in Figure 5.1(b), indicate that a higher protein content in the feedstock leads to an increased deamination rate [282]. Consequently, this results in reduced protein levels in HTL aqueous but higher concentrations of total ammonia. For all sludge types from both treatment plants, the concentrations of dissolved proteins in HTL aqueous were inversely proportional to the protein content of the feedstock, while total ammonia concentrations showed a direct proportionality with the feedstock protein content. These results indicate that the HTL condition used in this study, 332°C and 16.9 min, was sufficiently severe to decompose proteins. For WWTP 1, dissolved protein concentration in HTL aqueous decreased as the SS ratio increased due to deamination. For WWTP 2, on the other hand, the dissolved protein concentration was similar for all PS:SS ratios. However, total ammonia concentration increased from 3790 to 7830 mg/L, related to the higher protein concentration in the feedstock. Mesophilic DS had the highest protein-containing feedstock (89.5 g protein/kg feedstock) that resulted in 12000 mg/L total ammonia concentration in its HTL aqueous phase. Fatty acids, generated from decomposition of carbohydrates and lipids and deamination of proteins [9], are given in Figure 5.1(c). The total fatty acid concentration was found to be directly proportional to the SS ratio in mixed sludge. This change was particularly evident in municipal 102 sludge from WWTP 1, where the total fatty acid concentration increased by 2000 mg/L between the high PS case (PS:SS:59:41% TS wt.) and the average case (PS:SS:45:55% TS wt.), and similarly between the average case and the high SS case (PS:SS:31:69% TS wt.). The highest fatty acid concentration was measured in the WWTP 1 high SS case as 16900 mg/L. The fatty acid concentrations between the two WWTPs differed significantly. This difference might be attributed to the protein content in the feedstock, which appears to have a more pronounced effect than carbohydrates. Alternatively, variations in protein and carbohydrate structures between sludges from two treatment plants could be contributing to the observed differences in the decomposition effect. Fatty acids are highly biodegradable and serve as a valuable substrate for methane production in AD. Fatty acid concentrations of thermophilic and mesophilic DS-derived HTL aqueous were 8550 mg/L (21 mg VFA/g VSsludge) and 6700 mg/L (20 mg VFA/g VSsludge), respectively. The difference in fatty acid generation was less pronounced in the production per gram of VS of sludge. The orthophosphate-P concentrations in HTL aqueous samples derived from municipal sludge increased with an increase in the SS ratio in the feedstock. This observation suggests that phosphorus primarily originates from microorganism biomass, given that P is a crucial element in biomass composition. Although approximately 90% of P ends up in the hydrochar phase [13], a noteworthy amount of orthophosphate-P was still present in the HTL aqueous phase. Increasing with SS ratio, WWTP 1 municipal sludge derived HTL aqueous orthophosphate-P concentration was in the range of 260-420 mg/L, while this concentration range was 200-465 mg/L for WWTP 2. These values might be much higher if these treatment plants were performing biological P removal. The mesophilic DS feedstock produced almost twice the orthophosphate-P concentration in HTL aqueous compared to the thermophilic, potentially attributable to the longer retention time in the digesters of WWTP 1 (thermophilic). Total sugar levels in HTL aqueous were lower than 210 mg/L, as carbohydrates are the first degrading group under hydrothermal conditions. For both treatment plants, total sugar concentrations were lower as PS ratio in mixed sludge decreased. The pH of HTL aqueous varies mainly depending on the concentration of total ammonia concentration. Therefore, the pH levels were observed to rise in correlation with the SS ratio in municipal sludge, a trend evident in the HTL of mixed sludge from both treatment plants. The pH range for WWTP 1 fell between 7.23 and 7.95, while for WWTP 2, it ranged from 7.04 to 7.9. Parallel to pH, total alkalinity showed the same trend and increased with the amount of SS in municipal sludge. Using feedstocks of high PS and high SS in HTL, the total alkalinity values of 103 HTL aqueous were doubled for WWTP 1 and tripled for WWTP 2. The total alkalinity levels in HTL aqueous for both treatment plants reached a range of 17600-17800 mg CaCO3/L. The highest protein loading in HTL was achieved using mesophilic digested sludge (DS) as a feedstock, resulting in a pH of 8.22 and total alkalinity of 31700 mg CaCO3/L. In comparison, thermophilic DS yielded a pH of 7.79 and a total alkalinity of 23400 mg CaCO3/L. All HTL aqueous samples were found suitable for biological treatment in terms of pH and alkalinity. Phenolic compounds form in the HTL process as a decomposition product of carbohydrates, proteins, and lipids [260]. Hence, looking at the total phenolic compound results given in Figure 5.1(f), it is evident that there was only a slight difference among all sludge types and treatment plants. Based on these findings, it can be concluded that the concentration of phenolic compounds in HTL aqueous does not vary significantly based on feedstock characteristics. Earlier findings suggest that, among organic groups, carbohydrates are the primary contributors to the formation of phenolic compounds in HTL [260]. The results of this study align with earlier findings, demonstrating that a higher concentration of carbohydrates in the feedstock (higher PS) led to a slightly higher phenolic compound concentration in HTL aqueous. These levels of phenolic compounds have the potential to inhibit both aerobic [252] and anaerobic microorganisms [253]. Therefore, HTL aqueous should not be utilized in biological treatment without dilution or pretreatment. Some N-heterocyclic molecules, such as imidazoles and pyridines, are well-known inhibitors that form during the degradation of proteins in the HTL process [283]. Along with N-heterocycles, certain ketones, such as pentanones and hexanones, are also formed during HTL. These compounds have the potential to inhibit microorganisms by interfering with enzymatic activity [284]. In Figure 5.2, concentrations of selected N-heterocycles and ketones detected by GC-MS are illustrated. Since N-heterocycles are protein decomposition products and ketones are mostly generated in carbohydrate degradation, increasing N-heterocyclic and decreasing ketone concentrations were observed as SS ratio in mixed sludge increased. In the high SS case of WWTP 2, the highest concentration of an N-heterocyclic compound was measured at 1290 mg/L for 2-methyl pyrazine. The highest total quantified N-heterocycles concentration was also obtained from the same feedstock as 2030 mg/L. Although pyrazine concentrations seem to be high compared to other N-heterocycles, they are known to be completely biodegradable [285]. However, the same cannot be said for pyridines. Pyridine and its derivatives have been proven to significantly inhibit microorganisms, and there are studies actively investigating the antimicrobial properties of these compounds for medical use [286]. The concentrations of cyclopentanone and 104 cyclohexanone in the aqueous samples from HTL were found to be at levels that are not inhibitory [287]. DS-derived HTL aqueous samples had significantly lower N-heterocycles concentrations compared to municipal sludge-derived HTL aqueous samples. This could be due to N-heterocycles forming earlier in the DS HTL process, linked to lower organic material in the reactor, and then degrading into smaller molecules. Figure 5.2. Quantitative analysis of selected N-heterocycles compounds and ketones in HTL aqueous samples using gas chromatography-mass spectrometry (GC-MS). High PS, average, and high SS correspond to PS:SS mixing ratios (% TS wt.) of 59:41%, 45:55%, 31:69% and 70:30%, 57:43%, 39:61% for WWTP 1 and WWTP 2, respectively. 5.3.3 Biochemical oxygen demand The BOD tests were conducted over a period of 17.5 days until all reactors stopped consuming oxygen. The UBOD results were 75-79% higher than BOD5 levels therefore it was concluded that the five-day test was not sufficient for the BOD testing with 1% HTL aqueous volume and at F/M ratio of 10 g CODsample/g VSSinoculum. The UBOD curves and oxygen uptake rates (OUR) for all HTL aqueous samples are illustrated in Figure 5.3. The results showed a direct correlation with COD levels, suggesting that the aerobic biodegradability did not vary significantly between different types of feedstock sludge. The highest UBOD of 68600±260 mg/L was measured for the HTL aqueous derived from WWTP 1 high SS case. For both treatment plants, the higher the SS ratio in the municipal sludge, the higher the UBOD levels became. The increase in UBOD levels can be attributed to the presence of biodegradable protein degradation products in the HTL aqueous, including fatty acids, amines, amides, amino acids, protein oligomers, and certain N-heterocyclic compounds. 105 Figure 5.3. Specific oxygen uptake and oxygen uptake rate of HTL aqueous samples measured in respirometric biochemical oxygen demand tests. High PS, average, and high SS correspond to PS:SS mixing ratios (% TS wt.) of 59:41%, 45:55%, 31:69% and 70:30%, 57:43%, 39:61% for WWTP 1 and WWTP 2, respectively. Lighter color indicates standard deviation. (n=2) 106 The UBOD increase related to SS sludge ratio in the mixed sludge was substantial for WWTP 1, whereas PS to SS ratio did not cause as much difference in UBOD levels for WWTP 2 mixed sludge derived HTL aqueous. The height of the initial peak in OUR suggested the presence of easily biodegradable substrates in the HTL aqueous samples. According to UBOD and OUR results, mixed sludge from WWTP 1 generated an HTL aqueous stream with much higher aerobically biodegradable matter, regardless of PS:SS ratio. The UBOD levels of these samples were 26-36% higher than those derived from WWTP 2 municipal sludge. This was expected due to HTL aqueous derived from WWTP 1 mixed sludge having significantly higher concentrations of COD and fatty acids. Similar UBOD levels were reported in the literature for PS- and SS-derived HTL aqueous with and without catalyst [281]. The UBOD levels of mesophilic and thermophilic DS-derived HTL aqueous samples were 54.2 g/L (226 mg O2/g VSDS) and 50 g/L (170 mg O2/g VSDS). In the UBOD test of DS sludge derived HTL aqueous samples, there were no sharp oxygen uptake rate peaks. This indicates that the readily available substrate levels were not as high as those derived from mixed sludge. There were no noteworthy levels of inhibition observed in the UBOD tests, as the HTL conditions in this study did not generate toxic HTL aqueous capable of inhibiting aerobic treatment at 1% concentration. For a potential full-scale HTL application, the concentration of HTL aqueous after dilution with wastewater was calculated as 0.1% [288]. 5.3.4 Biochemical methane potential The anaerobic biodegradability of HTL aqueous samples obtained from various sludge types was measured following the BMP protocol. Mesophilic and thermophilic test results, along with methane generation rates and VFA levels observed during AD, are presented in Figure 5.4 and Figure 5.5. Validations of BMP tests for both operation temperatures (35°C and 55°C) were conducted using glucose and given in Figure 5.4 and Figure 5.5. The BMP results showed a similar trend to UBOD results, indicating that the differences in aerobic and anaerobic biodegradability due to feedstock variations were changing in a similar manner. Following this trend, HTL aqueous derived from WWTP 1 mixed sludge yielded 214-243 mL CH4/g COD methane potential under mesophilic condition, which increased in correlation with the SS ratio. In contrast, the HTL aqueous derived from WWTP 2's mixed sludge yielded 209-218 mL CH4/g COD under mesophilic condition, showing no significant variation with changes in the PS to SS ratio. 107 Figure 5.4. Cumulative methane production curves, methane production rates, and volatile fatty acids concentrations of mesophilic biochemical methane potential tests conducted on HTL aqueous samples derived from various sludge types (n=6). High PS, average, and high SS correspond to PS:SS mixing ratios (% TS wt.) of 59:41%, 45:55%, 31:69% and 70:30%, 57:43%, 39:61% for WWTP 1 and WWTP 2, respectively. 108 Figure 5.5. Cumulative methane production curves, methane production rates, and volatile fatty acids concentrations of thermophilic biochemical methane potential tests conducted on HTL aqueous samples derived from various sludge types (n=6). High PS, average, and high SS correspond to PS:SS mixing ratios of (% TS wt.) 59:41%, 45:55%, 31:69% and 70:30%, 57:43%, 39:61% for WWTP 1 and WWTP 2, respectively. 109 The trend in BMP at thermophilic temperature was similar to that observed in the mesophilic tests, but 20-30% lower. Lower biodegradability in thermophilic conditions is expected as the microorganisms at that temperature are not as diverse as in mesophilic conditions, resulting in a community that is more susceptible to inhibition [95]. Furthermore, this reduced diversity limits the variety of substrates that can be utilized effectively. The highest methane production was observed for high SS case of WWTP 1 as 196 mL CHâ/g COD. This value reduced as the PS ratio in the feedstock mixed sludge increased. For the high PS case, the methane potential of HTL aqueous dropped to 162 mL CHâ/g COD. Methane potentials measured for WWTP 2 mixed sludge derived HTL aqueous were in the range of 147-152 mL CH4/g COD, showing no significant variation by PS to SS ratio. These methane potential trends for WWTP 1 and WWTP 2 were similar under both mesophilic and thermophilic conditions. The VFA concentrations showed that microorganisms in mesophilic condition had no issue in generating and consuming VFAs, which is an indication of healthy AD process. Similar healthy VFA consumption was also observed for thermophilic test with the exception of HTL aqueous obtained from WWTP 2 high SS case. This might be due to elevated N-heterocycles concentration in this sample, as mentioned in Section 5.3.2. Mesophilic and thermophilic DS-derived HTL aqueous samples generated 214 to 225 mL CH4/g COD under mesophilic condition and 157 to 164 mL CH4/g COD under thermophilic condition. In both mesophilic and thermophilic BMP assays, HTL aqueous derived from thermophilic DS produced more methane than mesophilic DS, despite having lower COD and UBOD levels. These results indicate that BMP is more closely correlated with readily biodegradable organics, such as VFA, than with COD levels. Similar methane yields were reported in the literature for HTL aqueous obtained at various temperatures by using different feedstocks such as swine manure [267], Spirulina [289], rice straw [228] and cornstalk [199]. When the results of this study are evaluated in conjunction with findings from the literature, it can be concluded that mesophilic conditions are more favorable for the AD of HTL aqueous, and the level of biodegradability significantly (p-value = <0.0001) vary depending on the feedstock 5.3.5 Kinetic modeling Kinetic modeling was performed on the results of UBOD, mesophilic AD, and thermophilic AD to calculate the kinetic parameters of specific maximum production (P), maximum reaction rate (Rm), and lag phase duration (λ) obtained for HTL aqueous samples. The calculated 110 parameters are given in Table 5.2. The kinetic parameters indicated the absence of a lag phase in all the tests, attributed to the HTL aqueous being a thermochemically hydrolyzed substrate. It also shows that there was no noteworthy level of inhibition at the outset of the tests. The value of Rm from BOD increased in correlation to the SS ratio in the mixed sludge as can be seen in Table 5.2. This suggests that the hydrothermal degradation of SS leads to a higher concentration of readily biodegradable organics in HTL aqueous compared to PS. The same trend was observed in mesophilic and thermophilic BMP results of HTL aqueous derived from WWTP 1 mixed sludge. However, in the high SS case of WWTP 2, the maximum reaction rate was lower than the average and high PS case. This indicates slight inhibition, possibly caused by inhibitory protein decomposition products in HTL aqueous. Table 5.2. Kinetic parameters of biochemical tests performed on HTL aqueous samples obtained from HTL of various sludge types. Biochemical oxygen demand test Kinetic parameters WWTP 1 WWTP 2 Digested sludge High PS Average High SS High PS Average High SS Mesophilic Thermophilic P 51711 60245 64847 38807 46610 50443 51318 47113 Rm 455 494 575 344 387 454 470 447 λ 0 0 0 0 0 0 0 0 R2 0.9532 0.9670 0.9673 0.9605 0.9663 0.9561 0.9659 0.9639 Adj-R2 0.9532 0.9670 0.9673 0.9605 0.9663 0.9560 0.9659 0.9639 Mesophilic biochemical methane potential test Kinetic parameters WWTP 1 WWTP 2 Digested sludge High PS Average High SS High PS Average High SS Mesophilic Thermophilic P 189 196 219 177 186 172 178 193 Rm 1.10 1.40 1.59 1.26 1.26 1.10 1.23 1.35 λ 0 0 0 0 0 0 0 0 R2 0.9423 0.9278 0.9510 0.9084 0.9095 0.9112 0.9051 0.9120 Adj-R2 0.9231 0.9037 0.9346 0.8778 0.8793 0.8816 0.8734 0.8827 Thermophilic biochemical methane potential test Kinetic parameters WWTP 1 WWTP 2 Digested sludge High PS Average High SS High PS Average High SS Mesophilic Thermophilic P 155 164 188 134 138 147 150 153 Rm 0.60 0.68 0.80 0.66 0.64 0.44 0.50 0.79 λ 0 0 0 0 0 0 0 0 R2 0.9527 0.9727 0.9859 0.9420 0.9370 0.9267 0.9280 0.9669 Adj-R2 0.9370 0.9635 0.9812 0.9226 0.9160 0.9022 0.9040 0.9558 P: Specific maximum production (mL CH4/g COD) (consumption for BOD, mg/L), λ: Lag phase (h) Rm: Maximum reaction rate (mL CH4/g CODÂ·h-1 for BMP and mg O2/LÂ·h-1 for BOD), Adj-R2: Adjusted R2 5.3.6 Biodegradability and principal component analysis The biodegradability levels of HTL aqueous samples in aerobic and anaerobic environments, obtained from Equations (4.1) and (4.2), are illustrated in Figure 5.6(a). The microbial diversity effect discussed in Section 5.3.4 was clear in biodegradability levels. The most 111 diverse community, aerobic microorganisms, achieved the highest biodegradability levels. Results indicate that 70-78% biodegradability can be achieved in aerobic treatment regardless of treatment plant or sludge type. Slightly higher biodegradability levels were reported by Kulikova et al. (2023) for HTL aqueous generated at 240-280°C from feedstock of PS an SS. In all tests, the HTL aqueous samples exhibiting the highest biodegradability were derived from the high SS case of WWTP 1 and the average case of WWTP 2. For WWTP 1, a clear trend was observed where biodegradability increased in correlation with the SS ratio in the mixed sludge. However, the high SS ratio also caused a significant increase in COD levels, and proteins in SS which may cause the formation of non-biodegradable degradation products. Mesophilic DS generated an HTL aqueous phase with higher COD and BOD levels compared to thermophilic DS derived HTL aqueous. This was also supported by kinetic parameters. Mesophilic DS derived HTL aqueous reached a slightly higher Rm value in the aerobic test. In contrast, thermophilic DS derived HTL aqueous had higher P and Rm values in both AD tests. This could be explained by the different protein decomposition levels during HTL. In HTL derived from mesophilic DS, a greater degree of protein decomposition occurred as supported by the elevated total ammonia levels, which likely contributed to higher concentrations of N-heterocycles. Higher degradation was observed, indicated by increased BOD; however, AD tests revealed the formation of inhibitory compounds such as pyridines, which reduced the reaction rate compared to that in thermophilic DS-derived HTL aqueous biodegradability tests. According to the Pearson correlation plot given in Figure 5.6(b), aerobic biodegradability was correlated to feedstock protein content. Anaerobic biodegradability, on the other hand, was strongly correlated with fatty acids and dissolved proteins in the HTL aqueous phase, as these parameters represent biodegradable compounds. Feedstock carbohydrate content was found to affect ketones, total phenolics, and total sugar concentrations of HTL aqueous. High protein content in the feedstock was found to lead to an increase in pH, as well as higher levels of total alkalinity, total ammonia, and orthophosphate-P concentrations. As carbohydrate and lipid degradation generate high concentrations of fatty acids, they were found inversely correlated to pH and total alkalinity. Finally, higher amounts of protein in the feedstock led to enhanced protein degradation rates and subsequent breakdown of protein degradation products, such as N-heterocycles. Therefore, the protein content of feedstock was found to be inversely correlated with the remaining N-heterocycles concentration in HTL aqueous. 112 (a) (b) Figure 5.6. (a) Biodegradability levels of HTL aqueous samples obtained from HTL of various sludge types and (b) Pearson correlation plot of feedstock characteristics, HTL aqueous composition, and HTL aqueous biodegradability. High PS, average, and high SS correspond to PS:SS mixing ratios (% TS wt.) of 59:41%, 45:55%, 31:69% and 70:30%, 57:43%, 39:61% for WWTP 1 and WWTP 2, respectively. Diameters represent the magnitude of the correlation coefficient. 113 5.4 Summary Key findings obtained from the results of Chapter 5 include: â¢ The PS to SS ratio in the mixed sludge exhibited variations throughout the year at the WWTPs, significantly impacting the carbohydrate and protein content of the mixed sludge (p-value<0.0001). â¢ Significant variations were observed in the mixed sludge characteristics between two WWTPs, leading to notable differences in the HTL aqueous phase obtained from these mixed sludges. â¢ Higher SS amount (higher protein content) in the mixed sludge caused HTL aqueous to have higher COD, total ammonia, N-heterocycles, fatty acids, orthophosphate-P, pH, and total alkalinity. â¢ Higher PS amount (higher carbohydrate) in the HTL feedstock mixed sludge led to an increase in dissolved protein, total sugar, total phenolic compounds, and ketone concentrations in the HTL aqueous. â¢ Mesophilic DS exhibited higher protein content compared to thermophilic DS. Consequently, the HTL aqueous phase derived from mesophilic DS contained elevated levels of total ammonia, orthophosphate-P, pH, and total alkalinity. However, the differences in COD, total phenolic compound, N-heterocycles, and ketone concentrations were negligible. â¢ In the UBOD tests, increased SS in the mixed sludge resulted in HTL aqueous with significantly elevated UBOD levels for WWTP 1. The trend was similar for the HTL aqueous derived from the mixed sludge of WWTP 2; however, the increase was not as pronounced as the one observed for WWTP 1. â¢ High SS in the HTL feedstock mixed sludge caused a higher methane potential for the HTL aqueous in WWTP 1. In contrast, the methane potential of HTL aqueous derived from WWTP 2 mixed sludge was not affected by the PS:SS ratio. â¢ Overall, the biodegradability levels of HTL aqueous samples were determined to be 69-78% under aerobic conditions, 58-70% under mesophilic anaerobic conditions, and 42-56% under thermophilic anaerobic conditions. â¢ The most biodegradable mixed sludge derived HTL aqueous was obtained from the high SS case for WWTP 1 and the average case WWTP 2. â¢ The biodegradability variation between mesophilic and thermophilic DS HTL feedstocks derived HTL aqueous was negligible. 114 â¢ Pearson's correlation analysis unveiled a positive correlation between aerobic biodegradability and feedstock protein content. Anaerobic biodegradability, on the other hand, was positively correlated to dissolved proteins and fatty acids in the HTL aqueous. Results in this chapter support that there is significant variation in HTL aqueous characteristics depending on the sludge composition and the source WWTP, thereby supporting Hypothesis 3. 115 Chapter 6: Treatment of municipal sludge derived hydrothermal liquefaction aqueous phase in continuous-flow activated sludge systems of wastewater treatment plants 6.1 Introduction In Chapters 4 and 5, aerobic treatment and mesophilic anaerobic digestion were identified as suitable biological processes to treat HTL aqueous derived from municipal sludge. While mesophilic AD would ideally suit the treatment of substrates with high COD levels, the presence of inhibitory compounds in HTL aqueous, such as pyridines and phenolic compounds, poses challenges for anaerobic treatment of HTL aqueous at itâs full strength (without dilution) [290]. Furthermore, only a minority of WWTPs are equipped with an anaerobic digester. Contrary to digesters, most of the treatment plants include secondary treatment units, which means that they are capable of performing aerobic biological treatment. Despite the added cost of aeration, treating HTL aqueous together with wastewater in aerobic treatment could be the most viable option for many WWTPs. This approach would enable the treatment of HTL aqueous by utilizing the existing infrastructure in WWTPs. Moreover, the dilution effect and the resilience of aerobic microorganisms to inhibitors in this approach would likely reduce the risk of inhibition in the biological treatment process. However, to the best of my knowledge, continuous-flow aerobic treatment of wastewater blended HTL aqueous has never been investigated. Only three studies in the literature have investigated the aerobic treatment of HTL aqueous, utilizing batch aerobic reactors to conduct BOD tests [281,288,291]. Prior to implementing full-scale aerobic treatment of HTL aqueous, a lab-scale continuous-flow bioreactor investigation is essential. The work in this Chapter investigated the impact of blending and treating HTL aqueous with influent wastewater in a continuous-flow activated sludge process on various aspects, including the characteristics of treated effluent, the stability of aerobic reactors, and the performance of downstream processes, such as secondary settling and UV disinfection. The main objective of this chapter was to identify potential challenges in the aerobic treatment of HTL aqueous before its implementation in WWTPs and to understand the effects of HTL aqueous blending on the characteristics of the effluent. The approach taken aims to uncover the unknown aspects of this treatment process and assess its effectiveness for potential larger scale HTL applications for municipal WWTPs. 116 6.2 Materials and methods 6.2.1 HTL feedstock preparation and operation Thickened primary and secondary sludge streams were received from Annacis Island WWTP located in Delta, British Columbia, Canada. Upon receipt, thickened primary and secondary sludge streams were combined in a 1:1 mass ratio, reflecting the average annual sludge generation that is representative of a typical WWTP. Sludge dewatering and HTL of the sludge cake were performed following the method described in Section 3.2.1 and Section 3.2.3, respectively. The HTL process was conducted at a temperature of 332°C and with a retention time of 16.9 minutes, in accordance with the identified optimum HTL conditions [10]. Detailed HTL operation procedure was described in Section 3.2.3. 6.2.2 HTL aqueous characterization Determination of COD, total alkalinity, total ammonium nitrogen (TAN), orthophosphate P, dissolved proteins, total sugar, and total phenolic compound concentrations were conducted using the methods described in Section 3.2.4. Five-day and ultimate BOD (BOD5/UBOD) levels of the HTL aqueous sample were measured using an online respirometer system as described in Section 4.2.3. Concentrations of fatty acids were determined using an Agilent 7890A GC-FID system as described in Section 3.2.4.8. Selected pyridines, pyrazines, indole, and cyclic ketones in the HTL aqueous sample were quantified using an Agilent 8860 GC-MS system mentioned in Section 5.2.3. All characterization analyses were performed in triplicate. 6.2.3 Continuous-flow activated sludge process design Two parallel sets of lab-scale aeration tanks, followed by settling tanks were designed and fabricated, to simulate a continuous-flow activated sludge process. The design of the units was performed using DesignSpark Mechanical 3-D CAD software (Ansys Inc., USA). The aeration tanks had a liquid volume of 1 L with 10 cm diffusers covering the entire bottom area to avoid any dead zones in the reactor. Settling tanks were designed as 500 mL (liquid volume) reverse conical units, enabling the settled sludge to compress. Volume ratios were selected to ensure sufficient retention times in the settling units. Designed units were fabricated using a modified Ender 3-V2 3D printer (Shenzhen Creality 3D Technology Co. Ltd., China). The units were printed using clear polyethylene terephthalate glycol (PETG) filament, while the gaskets between the units and lids were made from flexible thermoplastic polyurethane (TPU). The 3D printed reactors were epoxy coated to ensure water tightness. The 3D printer used in this chapter and the 3D modeled aerobic reactor design with settling tank are presented in Figure 6.1. 117 (a) (b) Figure 6.1. (a) Ender 3-V2 3D printer and (b) designed lab-scale treatment units in 3D CAD software. During operation, aeration tanks were placed on Fisherbrandâ¢ Isotempâ¢ 10Ã10 cm stirring plates (Thermo Fisher Scientific, USA), and a 2.5 cm stirrer bar was placed inside each tank. Along with pneumatic mixing provided by an aeration pump, 200 rpm magnetic stirring was also employed to ensure homogeneous operation. Feeds and effluents were kept at 6°C to decrease the risk of microbial decomposition during operation. Feeds were pumped into the aeration tank using a Masterflex® L/S peristaltic pump (Antylia Scientific, USA) and 3.2 mm Tygonâ¢ silver antimicrobial tubing (Saint-Gobain, France). A second peristaltic pump, using size 16 (3.1 mm inner diameter) Masterflexâ¢ Norpreneâ¢ pump tubing (Antylia Scientific, USA), was utilized to return the desired amount of biomass (sludge) from the bottom of the settling units to the aeration tanks. To prevent the spread of aerosols into the laboratory, the reactors were kept sealed and vented to a fume hood throughout the operation. Finally, the aerobic biological treatment was performed at room temperature (22-26°C). Photos of two parallel 3D printed lab-scale activated sludge systems are given in Figure 6.2. 118 Figure 6.2. Photos of (a) 3D printed, two parallel activated sludge treatment systems during the testing period, (b) top view of an empty aeration reactor with a 10 cm diameter diffuser at the bottom, and (c) HTL aqueous spiked reactor in operation, showing clear effluent and settled sludge. 6.2.4 Feed preparation for activated sludge process testing In this study, a complex synthetic wastewater formulation was utilized to ensure stable influent quality throughout the operation. For testing, complex synthetic wastewater was chosen instead of real wastewater to maintain consistent quality throughout the operation. Using actual sewage would have been impractical due to the substantial volume required to be transported from Vancouver weekly, in addition to the health and safety regulations associated with handling raw sewage. The formulation of the synthetic wastewater was a modified version of that used by Bracklow et al. [292]. The synthetic wastewater formulation used in this study is given in Table 6.1. Feeds for both activated sludge systems were prepared in 20 L batches every 3 days to ensure feed quality. The feeds were kept at 4°C before being transferred to 10 L feeding bags 119 that were kept at 6°C. The âcontrolâ activated sludge system was fed with synthetic wastewater, while the âtestâ activated sludge system had 1000 times diluted HTL aqueous in the synthetic wastewater. The dilution ratio was calculated using real WWTP data, under the assumption that the daily total volume of municipal sludge would be processed in an HTL unit. Throughout the operation, all prepared feeds and effluents were analyzed for the same parameters, explained in the following sections. Table 6.1. Synthetic domestic wastewater formulation used in this study. Ingredient Concentration (mg/L) Ingredient Concentration (mg/L) Peptone 17.4 H3BO4 0.3 Yeast extract 52.2 CuCl2 0.04 Milk powder 116.2 KI 0.06 Starch 122 MnSO4·H2O 0.32 Ammonium acetate 79.4 NaMoO4·2H2O 0.12 KH2PO4 23.4 ZnCl2 0.14 MgHPO4·3H2O 29 CoCl2·6H2O 0.3 Urea 91.7 FeSO4·7H2O 5.8 (NH4)Cl 12.8 N-Allylthiourea 10 6.2.5 Activated sludge process start-up, operation and monitoring The reactors were initiated with an aerobic inoculum, the details of which are provided in Appendix C (Table C1), sourced from the contact stabilization tanks of the Annacis Island WWTP (Delta, BC, Canada). Before initiating testing and data collection, the lab-scale activated sludge processes were supplied with synthetic wastewater for a period of 6 months to acclimatize the inoculum to the complex synthetic wastewater. The operation employed hydraulic retention times (HRT) of 3, 4.5, and 6 h, and a fixed sludge retention time (SRT) of 9 days, following the typical ranges for a complete mix activated sludge system treating municipal wastewater [246]. These HRT values corresponded to an F/M ratio of 0.6-0.8 kg BOD/kg mixed liquor volatile suspended solids (MLVSS)·d. However, the reactor operation failed at the 3-h HRT due to the high F/M ratio, resulting in slime bulking caused by high concentrations of extracellular polymeric substances. Therefore, the 3-h HRT was excluded from the results. Both 4.5- and 6-h HRTs were tested for 36 days (4ÃSRT) at steady state. The criteria for achieving steady state involved monitoring fluctuations in mixed liquor suspended solids levels (MLSS) and effluent COD concentrations, 120 with steady state confirmed when these fluctuations stayed below 5% across three sludge retention times (SRTs), totaling 27 days. For both 4.5- and 6-h HRTs, a return activated sludge ratio (R) of 0.3 was maintained. During the testing, pH levels in the aeration tanks were monitored by an Accumet AP110 portable pH meter (Thermo Fisher Scientific, USA). Aeration levels were manually adjusted with flowmeters, aiming to maintain a dissolved oxygen (DO) concentration of 2-3 mg/L. Dissolved oxygen concentrations and temperatures in the aerobic reactors were monitored using an Orion Starâ¢ A323 dissolved oxygen portable meter (Thermo Fisher Scientific, USA). Along with pH, temperature, and dissolved oxygen, MLSS and MLVSS levels of the aeration tanks were also measured daily as a parameter for reactor stability. Additionally, the level of TSS of the settled sludge was monitored every three days by extracting samples from the bottom of the settling tanks. MLSS and MLVSS measurements were conducted following Standard Methods 2540 [212]. The measured concentrations of MLSS were utilized to determine the volume of sludge required to be removed in order to maintain the 9-day SRT. Based on these calculations, the corresponding volumes of waste activated sludge were manually withdrawn from settling tanks on a daily basis. The sludge volume index (SVI) of the aerobic reactors was measured every three days according to Standard Methods 2710D [212]. The SVI measurements were performed in 100 mL graduated cylinders for 30 minutes. Results were calculated in mL/g TSSaeration tank. Also, the zeta potential of the mixed liquors was measured using a Malvern Zetasizer Nano ZS (Malvern Instruments, Netherlands). These were done to assess the impact of the HTL aqueous phase on the settling performance of the sludge. 6.2.6 Effluent characterization The effluents from the activated sludge processes were collected in 10 L bags that were stored in a refrigerator maintained at 6°C. Composite samples were collected daily and subsequently characterized. Characterization of the samples was conducted at varying intervals, daily, every three days, or every six days, depending on the specific parameter being measured. Daily parameters were COD (Section 3.2.4.1) and UV absorbance. The UV absorbance test was performed using a Synergyâ¢ HT microplate multi-detection reader (BioTek Instruments, USA) operated at a wavelength of 254 nm to determine the effect of untreated HTL aqueous compounds on the UV disinfection performance. The 254 nm wavelength was selected because it is the peak absorption wavelength for DNA. Concentrations of TAN, orthophosphate phosphorus (PO4-3-P), and total phenolic compounds were determined every three days according to the methods shared in Section 3.2.4. Finally, BOD5 levels of the treated effluents were determined every six days using an online respirometer following the method given in Section 4.2.3. 121 6.2.7 Data analysis Error bars given throughout the chapter indicate Â± standard deviation. Two-sample t-tests were conducted on the measured characteristics of the feeds, effluents, and activated sludge system operation parameters to assess the statistical significance of the differences. The objective of this analysis was to determine whether blending HTL aqueous into wastewater significantly impacts the operation and characteristics of the influent and effluent. Two-sample t-tests were performed using OriginPro 2023b software (OriginLab Corporation, Massachusetts, USA), and the test results are presented in Appendix E (Tables E1-4). In this chapter, the terms âsignificantâ and âinsignificantâ were used solely to refer to t-test results at 95% confidence level (α=0.05) throughout the results and discussion. 6.3 Results and discussion 6.3.1 HTL aqueous characteristics Prior to its use in the continuous-flow activated sludge system, the municipal sludge derived HTL aqueous phase, produced at 332°C and a retention time of 16.9 minutes, was thoroughly characterized. The characterization results, given in Table 6.2, were in line with the previous findings in the literature [49,260]. Based on the findings, it was observed that 60% of the COD of HTL aqueous was easily biodegradable, as indicated by the BOD5/COD ratio. Additionally, 77% of the COD was found to be utilizable by aerobic microorganisms, indicating the approximate level of COD removal that could be achieved through aerobic treatment. This level was similar to the ones reported in the literature for municipal sludge derived HTL aqueous [281,288]. The HTL aqueous phase typically contains high concentrations of ammonia, produced through the hydrothermal decomposition of proteins in deamination reactions. Although most of the phosphorus during the HTL process is mineralized and retained in the hydrochar phase, the HTL aqueous phase also contains a notable amount of phosphorus in the form of orthophosphate (PO4-3) [13,30]. The TAN and phosphorus levels measured in the HTL aqueous phase, when blended with influent wastewater at a 1:1000 (v/v) ratio, would result in an increase of 4.9 mg/L of TAN and 0.35 mg/L of PO4-3-P in the feed of the test activated sludge system. Since nitrogen and phosphorus are essential nutrients for biological treatment processes, the concentrations of these elements in the HTL aqueous phase are beneficial. They could provide the necessary nutrients required for the microbial activity involved in the biological removal of COD. Likewise, the pH and 122 alkalinity levels measured in the HTL aqueous phase were found to be suitable for biological treatment, providing an optimal environment for the microbial processes involved in the treatment. Table 6.2. Characterization results of hydrothermal liquefaction aqueous phase utilized for testing continuous aerobic treatment by spiking into synthetic wastewater in this study. Parameters HTL aqueous COD (mg/L) 85900±1000 BOD5 (mg/L) 51800±1100 UBOD (mg/L) 65900±3400 UBOD/COD (%) 77 Total ammonium nitrogen (mg/L) 5950±90 Orthophosphate-P (mg/L) 416±8 pH 7.85 Total alkalinity (mg CaCO3/L) 11500±20 Total phenolics (mg GAE/L) 1600±10 Dissolved proteins (mg/L) 12000±175 Total sugar (mg/L) 151±3 Acetic acid (mg/L) 6800±170 Propionic acid (mg/L) 2660±20 Butyric acid (mg/L) 1120±3 Valeric acid (mg/L) 256±8 Iso-valeric acid (mg/L) 700±53 Caproic acid (mg/L) 290±12 Iso-caproic acid (mg/L) 57±5 Heptanoic acid (mg/L) ND 2-Methyl-pyridine (mg/L) 39±3.2 2-Ethyl-pyridine (mg/L) 4.6±0 2,6-Dimethyl-pyridine (mg/L) 6.7±0.1 5-Ethyl-2-methyl-pyridine (mg/L) 4.7±0.1 2-Methyl-pyrazine (mg/L) 972±25 2-Ethyl-pyrazine (mg/L) 260±20 2,3-Dimethyl-pyrazine (mg/L) 163±13 2,3,5-Trimethy-pyrazine (mg/L) 145±18 2,3-Diethyl-5-methyl-pyrazine (mg/L) 3.4±0.2 Indole (mg/L) 2.3±0 3-Methyl-cyclopentanone (mg/L) 20±0.2 Cyclohexanone (mg/L) 5.3±0.2 4-Phenyl-2-butanone (mg/L) ND COD: Chemical oxygen demand, BOD5: Five-day biochemical oxygen demand, UBOD: Ultimate biochemical oxygen demand, GAE: Gallic acid equivalent, ND: Not detected Among major organic compound groups, carbohydrates are the first to be degraded under hydrothermal conditions. The final product of their degradation is mostly fatty acids [9]. Therefore, the total sugar level measured was quite low, while fatty acids constituted a notable 22% of COD in the HTL aqueous. Fatty acids are also formed in the degradation of proteins and lipids. The detected fatty acids in the HTL aqueous were composed of up to six carbons. Fatty acids that have more than six carbon atoms in their chain (non-volatile) tend to behave like lipids. As a result, 123 they predominantly remain in the bio-crude phase rather than the aqueous phase during the HTL process. Despite their high concentration, fatty acids are highly biodegradable, which means they should not pose any challenge to the aerobic treatment process. Proteins were the most abundant group detected in HTL aqueous. Spiking HTL aqueous into synthetic wastewater would bring an additional 24 mg/L of proteins to the feed of the test activated sludge system. The concentration of selected N-heterocycles (pyrazines, pyridines, and indole) and ketones in the HTL aqueous phase, detected by GC-MS, are given in Table 6.2. Among these compounds pyridines and cyclic ketones are known to be biodegradable and non-inhibitory to bacteria [285,287,293]. Pyridines, on the other hand, are well known to cause cell toxicity by damaging cell walls and DNA [286,294], especially at concentrations exceeding 150 mg/L. However, when concentrations of pyridine compounds detected in the HTL aqueous phase are diluted 1000 times, the resulting pyridine concentration in the feed of the test system would be at a negligible level (<1 mg/L). Finally, indole was reported to have a hormetic effect on bacteria, being positive at concentrations under 10-3 mol/L, and negative at concentrations exceeding that level [295]. The concentration of indole in the synthetic wastewater, when spiked with HTL aqueous, would become 2Ã10-5 mol/L. This level indicates that there should be no problem associated with indole during the aerobic treatment process, as it falls well within the range where indole is not expected to have a negative effect. 6.3.2 Activated sludge system treatment performance monitoring At the end of 6 months of acclimatization period, full testing of the reactors was started. The operational reactor parameters of pH, temperature, and dissolved oxygen concentrations were measured daily throughout the testing period and presented in Figure 6.3. Based on the pH levels presented in Figure 6.3(a), the control system, which treated synthetic wastewater, maintained average operational pH levels of 7.83 at 4.5-h HRT and 7.75 at 6-h HRT. In comparison, the test system spiked with HTL aqueous phase had average pH levels of 7.82 at 4.5-h HRT and 7.79 at 6-h HRT, respectively. The pH levels were closer to the higher end of the optimal range of 7-8 for aerobic treatment [296], which can be attributed to the aerobic systems primarily treating carbon-based compounds. Since there was no nitrification occurring in the system, no noteworthy pH change was observed. Also, the pH difference between treatment systems was statistically insignificant (p-value: 0.2164) indicating that HTL aqueous spiking did not affect reactor pH. 124 The average operating temperatures of the aeration tanks in control and test systems were 24.2°C and 23.6°C, respectively. The temperature levels were quite stable except for the slight increase between days 43-53, caused by the change in the room temperature. The HTL aqueous spiked aeration tank was 2.5% hotter throughout the operation period, which could be associated with higher biological activity resulting from the additional organic loading from the HTL aqueous. This increase was statistically significant (p-value: <0.0001). This slight increase could be beneficial, as it would likely enhance the reaction rate in the biological treatment process. In aerobic treatment DO level is essential to keep treatment operation stable. The DO levels in the aeration tanks were kept stable between 2-4 mg/L through a combination of aeration using a diffuser and mechanical mixing. This range was reported to be the most ideal for achieving the most efficient treatment for carbonaceous organics removal [297]. Throughout the testing period, the average DO levels were 2.69 and 2.89 mg/L in aeration tanks of the test and control systems. Both bioreactors being kept at the ideal range ensured optimal aerobic treatment efficiency. Figure 6.3. (a) pH, (b) temperature, and (c) dissolved oxygen levels of aeration tanks in control and test systems treating synthetic wastewater and HTL aqueous spiked synthetic wastewater, respectively. 125 The MLVSS (biomass) levels of aeration tanks were measured daily in duplicate as a bioreactor stability parameter. MLVSS was used as an indicator to detect any issues related to inhibition or poor settling, which could adversely impact sludge recycling. The MLVSS results during the testing period are presented in Figure 6.4. At 4.5 h HRT, the average MLVSS concentrations were 2000±292 mg/L for the HTL aqueous spiked test bioreactor and 1750±252 mg/L for the control bioreactor. When the HRT was increased to 6 h, both systems showed considerable stability in terms of MLVSS, resulting in concentrations of 2000±260 mg/L for the HTL aqueous spiked aeration tank and 1800±280 mg/L for the control aeration tank. For both tested HRTs, the MLVSS levels in the HTL aqueous spiked bioreactor were significantly higher (p-value: <0.0001) which could be attributed to the increased organic loading by the addition of HTL aqueous. As mentioned in the methodology, F/M ratios calculated using MLVSS and BOD values at 4.5-h HRT were 0.8 mg BOD/mg VSSÂ·d and 0.77 mg BOD/mg VSSÂ·d for test and control aeration tanks, respectively. These F/M ratios were at the higher end of the literature range (0.2-0.8 mg BOD/mg VSSÂ·d) reported for domestic wastewater treatment [298,299], enabling the investigation of the effects of spiking wastewater with HTL aqueous at high loadings. Data obtained for 4.5-h HRT could be particularly useful for WWTPs that are already operating under high loading conditions. On the other hand, F/M ratios at 6-h HRT were 0.6 mg BOD/mg VSSÂ·d for the test aeration tank and 0.57 mg BOD/mg VSSÂ·d for the control aeration tank. These F/M ratios at 6-h HRT were within ideal and safer levels for the operation of the treatment systems. Therefore, the results obtained at 6-h HRT were more representative of the HTL aqueous spiking effect in WWTPs operating under ideal F/M ratios. Figure 6.4. Mixed liquor volatile suspended solids levels of aeration tanks in control and test systems treating synthetic wastewater and HTL aqueous spiked synthetic wastewater, respectively. 126 6.3.3 Chemical oxygen demand The COD is an efficient parameter for determining the amount of organic compounds present before and after aerobic treatment. The synthetic wastewater used in this study had a COD concentration of 416±7.6 mg/L. Spiking the synthetic wastewater with HTL aqueous resulted in a significant increase of 18.8% in COD levels, reaching 494±6.7 mg/L (p-value: <0.0001) in feed of the test activated sludge system. Measured influent and effluent COD levels, COD removal efficiencies, and HTL aqueous originated COD removal efficiencies of test and control systems are given in Figure 6.5. The influent (to aeration tank) and effluent (from settling tank) COD levels presented in Figure 6.5(a) indicate that the aerobic treatment significantly decreased the COD levels. However, the effluent from the HTL aqueous spiked test system had noticeably higher COD levels compared to that from the control system. The average effluent COD levels were 57.1±10 mg/L at 4.5-h HRT and 38.6±4.9 mg/L at 6-h HRT for the control system, whereas they were 79.5±7.4 mg/L at 4.5-h HRT and 56.3±7.5 mg/L at 6-h HRT for the test systems. These COD increases, corresponding to 39.3% and 45.7%, were statistically significant for both HRTs (both p-values: <0.0001). An increase in COD was expected in the effluent of test system, as HTL aqueous is well known to contain recalcitrant compounds [49]. The percentage of these compounds was detected as approximately 23% in the COD and UBOD analysis of HTL aqueous, equivalent to 20 mg/L COD when HTL aqueous was diluted 1000 times. Most of these non-biodegradable organic compounds are N-heterocycles, S-heterocycles, and phenolics [300,301]. In this study, phenolics (1.5 mg GAE/L) constituted 20% of the recalcitrant COD in the effluent from the HTL aqueous spiked reactor. Among the N-heterocycles, pyridines are well known to be recalcitrant and inhibitory [302]. Based on the COD removal performances of aeration tanks in Figure 6.5(b), the HTL aqueous spiked bioreactor had lower efficiency throughout the testing period which can be attributed to the increase of recalcitrant compounds after spiking. The COD removal percentages significantly differed between the aeration tanks for both tested HRTs (both p-values: <0.0001). The COD removal difference between test and control was approximately 2.7%. On the other hand, at the longer HRT of 6 h, overall COD removal for both bioreactors improved by 3.9%. The parallel changes in treatment performances indicate that spiking with HTL aqueous caused no unusual behavior. Furthermore, there was no sign of inhibition throughout the testing and the HTL aqueous spiked bioreactor was as stable as the control. 127 Figure 6.5. (a) Chemical oxygen demand levels of aeration tank influents and effluents, (b) chemical oxygen demand removal performances of aeration tank, and (c) removal performance of HTL aqueous originated chemical oxygen demand in aeration tanks working at 4.5-h and 6-h hydraulic retention times (n=3). 128 By using the COD removal data of both systems, HTL aqueous originated COD removals were calculated. According to the results given in Figure 6.5(c), average removals of COD originating from HTL aqueous of 72.5% and 79.2% were achieved at 4.5 h and 6 h HRTs. These efficiencies aligned with the estimated aerobic biodegradability of 77%, which was calculated using UBOD and COD measurements of the HTL aqueous. At HRT of 6 h, the HTL aqueous spiked bioreactor performed significantly better (p-values: 0.0004) in HTL aqueous treatment compared to 4.5-h HRT. If lower COD levels in the effluent were desired in WWTPs, an advanced oxidation process could be utilized to treat the remaining recalcitrant COD. 6.3.4 Biochemical oxygen demand The BOD5 is one of the most essential parameters in the effluent as it represents the amount of biodegradable organic matter that would be discharged to the receiving water body. For this reason, it is one of the most strictly regulated parameters in major urbanized countries such as the USA [303], China [304], and European countries [305]. The measured BOD5 levels of influents and effluents of the aeration tanks are presented in Figure 6.6. The BOD5 level of the synthetic wastewater was measured as 255±3.7 mg/L. The spiking of the HTL aqueous solution resulted in a significant increase in the BOD5 level (p-value: <0.0001), with a 21.9% rise bringing the level to 310±5.8 mg/L in feed of the test system. The average effluent BOD5 levels of the aeration tank in the control system were 19.6±3.1 mg/L at 4.5-h HRT and 19±5.7 mg/L at 6-h HRT. On the other hand, HTL aqueous spiked aeration tank effluent had BOD5 levels of 20.3±3.6 mg/L at 4.5-h HRT and 18.8±3.8 mg/L at 6-h HRT. For both retention times, the differences in BOD5 levels of the effluents were statistically insignificant (4.5-h HRT p-value: 0.8253 and 6-h HRT p-value: 0.4651), indicating that HTL aqueous spiking did not cause a change in the effluent BOD5. This was due to 92-94% of the biodegradable compounds being treated in the aerobic reactors, leaving negligible amounts of biodegradable organics in the effluent. Additionally, unlike COD, recalcitrant organic compounds are not registered in BOD5 measurements, as they do not biodegrade. These results suggest that treating HTL aqueous stream along with inflow wastewater in WWTPs would not alter the effluent BOD5 levels, and consequently, would not impact the effluent discharge environment in terms of level of biodegradable organics. 129 Figure 6.6. Influent and effluent five-day biochemical oxygen demand levels (BOD5) of aeration tanks in control and test systems treating synthetic wastewater and HTL aqueous spiked synthetic wastewater, respectively. 6.3.5 Total ammonium nitrogen (NH4+-N) Nitrogen content is the most legally regulated treated wastewater parameter worldwide in terms of discharge criteria due to its potential to trigger eutrophication [306]. HTL aqueous contains high concentrations of total ammonia, which caused a significant increase in TAN when spiked into synthetic wastewater (p-value: <0.0001). This 24.5% increase in TAN concentration raised the level from 21.3±2.3 mg/L to 26.5±3 mg/L in the feed of the test activated sludge system. This increase in TAN concentration could be negligible if the WWTP employs a biological nutrient removal process. The measured TAN concentrations of aeration tank influents and effluents are given in Figure 6.7. Since the complete mix aeration tanks systems were focused solely on treating carbon, and oxidative deamination occurs during the consumption of proteins which generates ammonia [307], there was a notable increase in TAN concentrations in both bioreactors at both tested HRTs. The impact of HRT on the TAN concentrations in the bioreactorsâ effluents was insignificant (control p-value: 0.8365 and test p-value: 0.1283), as the protein concentration remained stable in both influents throughout the testing period. 130 Figure 6.7. Influent and effluent total ammonium nitrogen concentrations (NH3-N) of aeration tanks in control and test systems treating synthetic wastewater and HTL aqueous spiked synthetic wastewater, respectively. The average TAN concentrations in the effluents at 4.5-h HRT were 63.6±4.2 mg/L for the control aeration tank and 72.5±4.3 mg/L for the test aeration tank. At this HRT, the increase in TAN concentration caused by HTL aqueous spiking was significant (p-value: <0.0001). At 6-h HRT, the TAN concentration was 10% higher in the HTL aqueous spiked bioreactorâs effluent. This significant difference increased the TAN concentration from 63.8±2.5 mg/L to 70.3±2.6 mg/L (p-value: <0.0001). The results indicated that, despite being diluted 1000 times, the HTL aqueous phase still had a significant impact on the effluent TAN concentration. The free ammonia concentration in the effluent of the HTL aqueous spiked aeration tank was estimated at 1.95 mg/L, while in the control aeration tank, it was lower, at 1.63 mg/L. These concentrations can vary considerably based on the pH and temperature. It is reported that free ammonia concentrations exceeding 36 mg/L, have the potential to inhibit nitrite oxidation [308]. After diluting 1000 times, reaching these free ammonia concentrations is unlikely; however, with lower levels of dilution, the HTL aqueous phase has the potential to inhibit the biological nutrient removal process. For municipal sludge derived hydrothermal carbonization aqueous phase, which has similar TAN levels to HTL aqueous, a minimum dilution of 100 times has been suggested to avoid inhibiting nitrification in WWTPs [309]. 6.3.6 Orthophosphate phosphorus (PO4-3-P) Phosphorus is the second important nutrient, along with nitrogen, that has the potential to cause eutrophication in water bodies receiving treated wastewater. During the HTL process, 131 85-92% of phosphorus is retained in the hydrochar phase, allowing for its recovery and use in agriculture as fertilizer [13]. The remaining phosphorus dissolved in the HTL aqueous phase is in form of PO4-3. Despite only a small fraction remains in the HTL aqueous phase, it caused a significant increase in phosphorus concentration in the influent of the test system when HTL aqueous was spiked into synthetic wastewater (p-value: <0.0001). The synthetic wastewater contained 9.5±0.7 mg/L PO4-3-P, which increased to 10.4±0.64 mg/L when spiked with the HTL aqueous phase. Although this increase might appear negligible, it is important to note that in some cases, the discharge limit for total phosphorus can be as low as 1 mg/L. [310]. Therefore, even an increase of 1 mg/L in phosphorus concentration should not be overlooked in the context of wastewater treatment and discharge standards. The measured influent and effluent PO4-3-P concentrations of aeration tanks of both activated sludge systems are given in Figure 6.8. According to the results, average PO4-3-P concentrations of the control and test bioreactors were 8.7±0.82 mg/L and 9.4±0.36 mg/L at 4.5-h HRT, and 8.5±0.45 mg/L and 9.1±0.38 mg/L at 4.5-h HRT, respectively. For both HRTs, an approximate increase of 7.7% in phosphorus concentrations was detected and determined to be statistically significant, resulting from the spiking of the HTL aqueous phase (4.5-h HRT p-value: 0.0085 and 6-h HRT p-value: 0.0007). When influent and effluent PO4-3-P concentrations were compared, a slight decrease was observed, as a result of consumption by microorganisms during the biological treatment process. Similar to nitrogen, the increase in phosphorus concentration would not pose a problem if the WWTP is equipped with a configuration designed for biological nutrient removal. Figure 6.8. Influent and effluent phosphorus concentrations (PO4-3-P) of aeration tanks in control and test systems treating synthetic wastewater and HTL aqueous spiked synthetic wastewater, respectively. 132 6.3.7 Total phenolic compounds Phenolic compounds are known for their recalcitrance, and their toxicity can affect a wide range of microorganisms and animals. In addition to causing obvious toxic effects, phenolic compounds have also minor negative impacts on biota, including lowered fertility, diminished survival rates in offspring, and growth inhibition [311]. Municipal wastewater typically contains low concentrations of phenolic compounds, so there is generally no need for concern regarding their presence in such water. The primary sources of phenolic compounds in wastewater are thermochemical processes and activities in the petroleum industry, including HTL. These industries are petroleum processing plants, petrochemical industries, coke ovens, and coal conversion industries [246]. Due to their negative environmental impact, nine phenolic compounds, 2,4,6-trichlorophenol, 4-chloro-3-methylphenol, 2-chlorophenol, 2,4-dichlorophenol, 2,4-dimethylphenol, 2,4-dinitrophenol, 4,6-dinitro-o-cresol, pentachlorophenol, and phenol, are included in the list of 126 priority pollutants by the U.S. Environmental Protection Agency (EPA) [312]. Although certain microorganisms are capable of aerobically utilizing phenolic compounds, this degradation process is limited in nature, leading to the accumulation of these compounds. Through bioaccumulation, phenolic compounds can have adverse effects on human health when they are released into the environment without adequate treatment [313]. Total phenolic compound concentrations measured in the influents and effluents of aeration tanks in both systems are presented in Figure 6.9. The synthetic wastewater fed to the control bioreactor was measured to contain 4±0.5 mg gallic acid equivalent (GAE)/L phenolic compounds. The HTL aqueous spiking caused a significant 64% increase in this level, raising it to 6.5±0.6 mg GAE/L (p-value: <0.0001). Approximately 50% of the phenolic compounds were removed during aerobic treatment. The average effluent total phenolics concentrations were 2±0.77 mg GAE/L at 4.5-h HRT and 1.6±0.34 mg GAE/L at 6-h HRT for the control bioreactor, whereas they were 3.8±0.57 mg GAE/L at 4.5-h HRT and 2.8±0.32 mg GAE/L at 6-h HRT for the test bioreactor. According to these results, HTL aqueous spiking caused significant increases in total phenolics concentrations for both 4.5-h HRT (p-value: <0.0001) and 6-h HRT (p-value: <0.0001). In the effluent of the test aeration tank, these recalcitrant phenolics were responsible for 20% of the observed increase in COD. These recalcitrant phenolic compounds can be removed via photocatalytic methods [314], adsorption [315], or oxidation [316] in WWTPs. 133 Figure 6.9. Influent and effluent total phenolic compound concentrations of aeration tanks in control and test systems treating synthetic wastewater and HTL aqueous spiked synthetic wastewater, respectively. 6.3.8 Settling and sludge volume index At WWTPs, the first downstream unit treatment process following the aeration tank is the settling unit, known as secondary settling. Since HTL aqueous was diluted 1000 times, a noticeable physical change in the influent that could affect the settling process was not expected. However, given the wide variety of organic compounds contained in HTL aqueous, a shift in the microbial community within the treatment system is inevitable after spiking with HTL aqueous. In the literature, such a shift in the microbial community has been reported under anaerobic conditions [251]. These shifts in the microbial community, occurring due to spiking with HTL aqueous, can impact the performance of the settling process. To evaluate the settling performance, the concentration of TSS in the settling underflows and the SVIs of the aeration tank biomass (MLSS) were measured throughout the testing period and presented in Figure 6.10(a) and (b), respectively. When settled sludge streams were tested, at a 4.5-hour HRT, the average TSS was 6680±1240 mg/L for the control settling tank, while the average TSS in the test settling tank was slightly higher, measured as 7170±1520 mg/L. TSS values at 6-h HRT were quite similar for both settling tank underflows, recorded at approximately 6900±1600 mg/L. During the bioreactor testing, no issues were detected in terms of TSS levels of settling tank underflows. If any issues had been identified, they could have halted the sludge recycling process, thereby disrupting the entire treatment system. Overall, the difference between the TSS concentrations of the settling tank underflows was insignificant (p-value: 0.2092). The SVI results given in Figure 6.10(b), however, told a different story. The test aeration tank biomass SVI levels were significantly higher than the control in both tested HRTs of 4.5 h (p- 134 value: <0.0001) and 6 h (p-value: <0.0001). The average SVI levels were 186±28 mL/g TSS for the control and 270±47 mL/g for the test bioreactors at a 4.5-h SRT, and 209±47 mL/g for the control and 379±62 mL/g for the test bioreactors at a 6-h HRT. This variation in SVI levels was possibly due to a shift in the microbial community caused by the spiking of HTL aqueous. Based on visual observation, the biomass in the control aeration tank tended to form larger flocs, while the biomass in the test bioreactor exhibited bulk settling. However, when both settling underflow TSS and SVI values are evaluated, it can be seen that the mixed liquor from the test aeration tank was simply settling more slowly, but with enough time, it settled just as well as the mixed liquor from the control aeration tank. Figure 6.10. (a) Total suspended solids levels reached in settling tank underflows of test and control systems treating synthetic wastewater and HTL aqueous spiked synthetic wastewater and (b) sludge volume index levels of mixed liquors from aeration tanks. 135 The increase in SVI levels after the addition of HTL aqueous could be explained by changes in zeta potential, which indicates the overall surface charge of the bacteria [317]. In the control bioreactor, the zeta potential was measured at -21.8±3.1 mV, whereas in the test bioreactor, it was -30.6±3.82 mV. Similar SVI levels were observed corresponding to similar zeta potential levels in the literature, resulted in sludge bulking [318]. The reason for the sludge bulking was the replacement of extracellular polymeric substances, which aid in forming flocs, with filamentous bacteria. Furthermore, these zeta potential levels were associated with a decrease in the level of extracellular polymeric substances and their protein to polysaccharide ratio [319]. Building on these findings in the literature, 2.5 mg/L of the cationic polymer Zetag® 7563 (BASF SE, Germany) was added to the mixed liquor from the test aeration tank, resulting in settling effectiveness comparable to that of the mixed liquor from the control aeration tank. In full-scale wastewater treatment plants, hydraulic conditions differ from those at the lab-scale, where more turbulent conditions are employed for aeration and mixing, leading to smaller flocs and consequently higher SVI levels. This effect was evident in the control reactor, where SVI levels are approximately 200 mL/g TSS, marking the upper limit of healthy sludge settling. Therefore, the SVI levels observed in a full-scale activated sludge system may differ from those reported in this dissertation. However, the findings in this study still suggest potential changes in settling performance, which should be considered when scaling up to full-scale operations. 6.3.9 UV absorbance The COD results indicated that the aeration tank treating synthetic wastewater spiked with HTL aqueous had recalcitrant compounds remaining in it after the treatment. These organic compounds, particularly the phenolic ones, have the potential to increase UV absorbance and, consequently, decrease the performance of UV disinfection, which is the tertiary treatment process following aerobic secondary treatment [320]. UV absorbances of aeration tank influents and effluents were monitored throughout the testing period and presented in Figure 6.11. Considering the concentrations of total phenolics and other organic compounds present in the HTL aqueous, given in Table 6.2, a significant increase in the influent UV absorbance was expected. This significant increase was measured at 35%, which raised the UV absorbance of the synthetic wastewater from 415±24 to 560±36 in the feed of the test activated sludge system. 136 Figure 6.11. Influent and effluent UV absorbance levels (at 254 nm) of aeration tanks in control and test systems treating synthetic wastewater and HTL aqueous spiked synthetic wastewater, respectively. At 4.5-h HRT, the test bioreactor effluentâs UV absorbance exhibited a significant 33% increase compared to the control bioreactorâs effluent (p-value: <0.0001). The remaining recalcitrant compounds caused this UV absorbance increase from 298±29 to 397±50. At 6-h HRT, on the other hand, effluent UV absorbance showed lower variation indicating that bioreactor effluent characteristics were more stable at this HRT, likely due to enhanced treatment effectiveness due to the increased retention time. At this HRT, the UV absorbance of effluent significantly increased by 22%, a change attributed to the HTL aqueous spiking (p-value: <0.0001). This 22% increase raised the UV absorption from 267±7 to 326±9. The increased phenolic compound concentration, mentioned in Section 6.3.7, was one of the reasons for these UV absorbance increases in the HTL aqueous spiked bioreactorâs effluent. Similar to phenolic compounds, other aromatic structured compounds also contribute to the increase in UV absorbance [321]. Aromatic compounds that are known to be difficult to biodegrade, such as pyridines, could increase UV absorption. The increase in UV absorption of these recalcitrant compounds could be advantageously utilized in UV-catalyzed treatment processes. In WWTPs, a UV-assisted oxidation process could be applied, wherein UV light would break down aromatic compounds while also assisting in forming radicals from the oxidant, resulting in a synergistic effect [322]. 6.4 Summary The following key findings were obtained from the results of Chapter 6: 137 â¢ Blending HTL aqueous into influent wastewater resulted in significant (p-values: <0.0001) increases in COD, BOD5, TAN, PO4-3-P, and total phenolic compound concentrations of influent wastewater entering activated sludge treatment process. â¢ There was no evidence of inhibition caused by HTL aqueous in the activated sludge process under the HRTs of 4.5-h and 6-h, corresponding F/M ratios of 0.8 and 0.6 kg BOD/kg MLVSSÂ·d, throughout the testing period. â¢ Significant (p-values: <0.0001) increases in COD, TAN, orthophosphate, and total phenolics concentrations were detected in the treated effluent of activated sludge process, due to the recalcitrant compounds in the HTL aqueous. â¢ COD removal of 72.5% and 79.2% were achieved for HTL aqueous at 4.5-h and 6-h HRTs, respectively. â¢ The BOD5 of treated effluent was not affected by HTL aqueous blending (4.5-h HRT p-value: 0.8253 and 6-h HRT p-value: 0.4651). â¢ Although the sludge volume index significantly increased after blending HTL aqueous into the wastewater, the total suspended solids concentration of the settling tanks underflow, and hence the sludge recycle, was not affected in the activated sludge process. â¢ Recalcitrant compounds remaining in the treated effluent were found to increase the UV absorbance at 254 nm. A tertiary treatment process, such as advanced oxidation or adsorption, could be employed to treat the recalcitrant compounds in the effluent originating from HTL aqueous, ensuring a more seamless treatment process. According to the findings presented in this chapter, the downstream processes of the aeration tank, settling, and UV disinfection were all significantly impacted by the addition of HTL aqueous to the influent wastewater. Therefore, Hypothesis 4, which posits that HTL aqueous can be co-treated with influent wastewater without causing significant disturbance to downstream processes, has been rejected. 138 Chapter 7: Conclusion 7.1 Overall summary and conclusions Treatment of HTL aqueous is one of the major bottlenecks in incorporating HTL technology into WWTPs for valorization of municipal sludge, which brings multiple challenges along the way. This dissertation addresses the following key points to overcome the challenges regarding process wastewater, HTL aqueous: â¢ A comprehensive analysis of HTL aqueous characteristics, along with the variations based on HTL temperature/retention time and changes in feedstock PS:SS ratios, was conducted. â¢ Evaluation of on-site biological treatment options in WWTPs, including aerobic, mesophilic anaerobic, thermophilic anaerobic, and dark fermentation treatments. â¢ The impact of HTL treatment using activated sludge systems, by blending it into the influent wastewater, on biological treatment units, downstream processes, and characteristics of treated effluent was investigated. The following conclusions can be drawn from the chapters of this dissertation that investigated HTL aqueous characterization and biological treatment: â¢ In the characterization analysis performed on HTL aqueous samples obtained at 290-360°C and 0-30 min HTL conditions from mixed sludge (1:1 PS to SS ratio by mass), it was found that mixed sludge derived HTL aqueous can contain COD levels up to 100 g/L and ammonia concentrations up to 7.1 g/L. GC-MS results revealed that the most abundant organic molecule type in HTL aqueous samples was N-heterocycles, followed by organic acids. HTL aqueous was found suitable for biological treatment in terms of its high total alkalinity (up to 14.8 g CaCO3/L), suitable pH levels (6.4-7.9), and sufficient trace elements that can support microorganisms. HTL aqueous sample obtained at 325°C and 0 min condition was found to contain 17 g/L VFAs, establishing it as the most suitable HTL aqueous for biological treatment. Lastly, the abundance of N-heterocycles and phenolic compounds (1.5-2 g/L) posed a potential risk for biological treatment due to their inhibition potential. â¢ The PS:SS ratio variation throughout the year in two tested WWTPs was found to alter the mixed sludge carbohydrate and protein content significantly, which led to variations in mixed sludge derived HTL aqueous characteristics (obtained at 332°C and 16.9 min retention time). Furthermore, there was a significant variation detected between the mixed sludge feedstocks obtained from two different WWTPs. Higher SS (higher protein content) containing mixed sludge generated an HTL aqueous with higher COD, total ammonia, N-heterocycles, fatty 139 acids, orthophosphate-P, pH, and total alkalinity. Mesophilic DS was found to contain higher proteins compared to thermophilic DS, which generated an HTL aqueous with higher total ammonia, orthophosphate-P, pH, and total alkalinity. Higher SS in mixed sludge also caused HTL aqueous to have elevated UBOD levels. In mesophilic and thermophilic anaerobic tests, higher SS ratio increased the methane potential of HTL aqueous for WWTP 1, while not affecting WWTP 2. In average, the biodegradability levels of HTL aqueous samples were determined to be 74% under aerobic conditions, 66% under mesophilic anaerobic conditions, and 49% under thermophilic anaerobic conditions. A positive correlation was identified between aerobic biodegradability and feedstock protein content, while anaerobic biodegradability was enhanced with higher concentrations of fatty acids and dissolved proteins in HTL aqueous. â¢ HTL aqueous samples obtained at 290-360°C and 0-30 min were found to have BOD5 levels ranging between 41 to 54 g/L. According to the biodegradability index (BOD5/COD), HTL aqueous samples obtained at lower temperatures than 340°C were found to be easily biodegradable in aerobic treatment. HTL aqueous obtained at 360°C exhibited signs of inhibition in aerobic tests, and complete inhibition in anaerobic tests. In mesophilic anaerobic tests, anaerobic biodegradability of HTL aqueous samples ranged between 0 (360°C-15 min) to 51.1% (325°C-0 min). This range was 0 (360°C-15 min) to 35% (325°C-0 min) in thermophilic anaerobic biodegradability tests. Evaluating the anaerobic biodegradability (Chapter 5) and detailed characterization (Chapter 3) it was found that HTL aqueous obtained at 360°C was unsuitable for AD, possibly due to the inhibitory compounds identified as pyridines, pyrrolidinones, piperidinones, pyridinols, and phenolic compounds. Mixed sludge derived HTL aqueous was found to be too inhibitory for dark fermentation. Overall, among on-site treatment options in a WWTP, aerobic and mesophilic anaerobic treatments were identified as the most suitable ones. â¢ Treatment of HTL aqueous in activated sludge systems by blending it to the influent wastewater was investigated. The blending of HTL aqueous into influent wastewater with 1000:1 (v/v) ratio led to notable increases in COD, BOD5, TAN, PO4-3-P, and total phenolic compound concentrations in influent wastewater, with statistically significant p-values (<0.0001). No evidence of inhibition was observed throughout the testing. The significant increases in influent characteristics were still present in the treated effluent, except for BOD5, due to the recalcitrant compounds in HTL aqueous. The blending of HTL aqueous resulted in elevated SVI levels in the settling, primarily attributed to an increase in zeta potential. The settling issue did not impact the TSS levels of the settling underflow, thus the sludge recycling 140 was not affected. Finally, the recalcitrant cyclic compounds in HTL aqueous were found to increase the UV absorption at 254 nm, which would decrease the UV disinfection performance in the WWTPs. 7.2 Limitations and recommendations for future research This dissertation was able to answer some important questions about the HTL aqueous characteristics and on-site treatment of municipal sludge derived HTL aqueous in WWTPs. However, while this study successfully addressed crucial inquiries, it is important to acknowledge that the dissertation has certain limitations. Limitations of this dissertation and recommendations for future investigations include: â¢ The HTL operation in this study was performed in a bench scale batch HT/HP system. Although the ramping and cooling times match some demonstration scale HTL plans, at full-scale, continuous-flow HTL reactors would be used, which can ramp to the desired temperature much faster than the bench scale reactor. Studies in literature report changing product characteristics after being exposed to heat at varying durations. In future studies, employing a continuous-flow HTL system would provide a more accurate representation of what a full-scale HTL plant could achieve. â¢ In chapter 4, it was revealed that aerobic and mesophilic anaerobic treatments are the most suitable processes for HTL aqueous treatment in WWTPs. Aerobic treatment was then further investigated in Chapter 6. However, a continuously-fed mesophilic anaerobic treatment was not a part of this dissertation, which has been investigated by other researchers at the Bioreactor Technology Group. The readers are recommended to review their work [290]. â¢ Achieving consistent results for the solids remaining in HTL aqueous within a small batch reactor posed a challenge, as these outcomes exhibited variation among different batches, primarily due to the unpredictable amount of solids remaining in the bio-crude. Therefore, HTL aqueous samples were filtered to obtain reproducible characterization data. This was also a limitation in the continuous-flow aerobic reactor testing. In a WWTP, inefficient hydrochar separation may result in the retention of solids in HTL aqueous potentially causing a 10-20 mg/L increase in TSS. This could be investigated using a continuous-flow HTL reactor, which would have more consistent and representative TSS concentrations in HTL aqueous. â¢ The impact of blending HTL aqueous into influent wastewater on the BNR processes (such as modified Bardenpho process employed at WWTPs in the Okanagan Valley) was out of scope. HTL aqueous possesses the potential to inhibit nitrification or biological phosphorus 141 removal, and this aspect could be investigated in a bioreactor system capable of performing BNR. â¢ Removing the recalcitrant compounds from the treated effluent was out of scope of this study. 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One factor model graphs of (a-b) orthophosphate-P, (c-d) total phenolic compounds, and (e-f) total volatile fatty acid concentration. 181 (a) (b) (c) (d) (e) (f) Figure A4. One factor model graphs of (a-b) pH, (c-d) Alkalinity, and (e-f) glycerol concentration. 182 Appendix B: Organic compounds detected in HTL aqueous samples obtained at various reaction temperatures and normal distribution test of characterization data. Table B1. Organic acids and their relative abundances in HTL aqueous samples detected by gas chromatography â mass spectrometry analysis. Retention time (min) Molecule name Molecule model Relative peak area (%) 290°C-15 min 325°C-15 min 360°C-15 min 7.615 Acetic acid 12.70 15.03 13.87 8.522 Propanoic acid 6.64 8.37 7.34 9.831 Butyric acid 4.31 5.75 6.25 11.76 Valeric acid 0.92 1.36 1.65 24.126 Hydrocinnamic acid 0.77 1.19 1.14 13.413 Valeric acid, 4-methyl- 0.84 0.95 0.80 21.944 Benzeneacetic acid 0.77 0.57 0.25 11.186 Hexanoic acid, 2-methyl- - - 1.57 15.951 Valeric acid, 4-oxo- - 0.78 0.71 11.174 Butyric acid, 2-methyl- - 1.37 - 11.151 Octanoic acid, 2-methyl- 1.01 - - 13.425 Heptanoic acid - 0.43 - 14.309 Hexanoic acid 0.13 - - : Carbon, : Hydrogen, : Oxygen 183 Table B2. N-heterocylic compounds and their relative abundances in HTL aqueous samples detected by gas chromatography â mass spectrometry analysis. Retention time (min) Molecule name Molecule model Relative peak area (%) 290°C-15 min 325°C-15 min 360°C-15 min 10.922 Pyrazine, methyl- 7.43 8.98 8.34 16.49 2-Pyrrolidinone 2.17 4.33 4.72 19.877 2-Piperidinone 2.85 4.46 3.71 9.257 1,3-Diazine 2.67 3.61 3.14 13.115 Pyrazine, ethyl- 2.20 2.75 2.42 12.988 Pyrazine, 2,5-dimethyl- 2.26 2.36 2.13 17.524 3-Pyridinol 1.08 1.36 1.10 20.004 3-Pyridinol, 6-methyl- 0.38 1.14 1.36 16.927 2,5-Pyrrolidinedione, 1-methyl- 0.65 1.05 0.85 13.218 Pyrazine, 2,3-dimethyl- 0.72 0.90 0.87 17.811 3-Aminopyridine 0.52 0.87 1.01 10.221 2-Methylpiperidine 1.44 0.86 - 30.062 Pyrrolo[1,2-a]pyrazine-1,4-dione, hexahydro- 1.92 - 0.31 15.342 Pyrazine, 2-ethyl-6-methyl- 0.46 0.56 0.52 19.579 5-Methyl-6,7-dihydro-5H-cyclopentapyrazine 0.52 0.49 0.49 9.544 Pyridine 0.45 0.51 0.54 16.1 2-Pyrrolidinone, 1-methyl- 0.13 0.33 0.96 18.534 1H-Pyrrole-2-acetonitrile, 1-methyl- 0.58 - 0.78 18.385 2,5-Pyrrolidinedione, 1-ethyl- 0.33 0.48 0.45 184 Table B2. N-heterocylic compounds and their relative abundances in HTL aqueous samples detected by gas chromatography â mass spectrometry analysis (continued). Retention time (min) Molecule name Molecule model Relative peak area (%) 290°C-15 min 325°C-15 min 360°C-15 min 30.177 2H-Imidazole-2-thione, 1,3-dihydro-1-methyl- 1.24 - - 17.719 Pyrazine, 3-ethyl-2,5-dimethyl- 0.34 0.37 0.46 28.684 2,5-Piperazinedione, 3-methyl-6-(1-methylethyl)- 1.14 - - 15.572 Pyrazine, 2-ethyl-3-methyl- 0.35 0.41 0.36 21.657 2H-Quinolizine, 1,3,4,6,7,9a-hexahydro- - 0.57 0.52 19.476 2-Pyridinamine, 5-methyl- 0.32 0.40 0.36 18.075 1-Ethyl-2-pyrrolidinone 0.12 0.28 0.66 28.994 2H-Imidazole-2-thione, 1,3-dihydro-1-methyl- 0.99 - - 27.099 (+-)-cis-3,4-Dimethyl-2-phenyltetrahydro-1,4-thiazine 0.97 - - 21.072 6-Aminoindoline 0.43 - 0.52 22.116 Caprolactam 0.28 0.40 0.26 19.166 2(1H)-Pyridinone, 5-methyl- 0.91 - - 22.484 2-Hydrazino-2-imidazoline - 0.45 0.42 18.534 1,5-Dimethyl-2-pyrrolecarbonitrile - 0.85 - 26.744 3,6-Dimethylpiperazine-2,5-dione 0.81 - - 17.937 Pyrazine, 2-ethyl-3,5-dimethyl- - 0.55 0.22 23.092 3-Acetyl-1H-pyrroline 0.14 0.31 0.26 14.642 2-Isopropylpyrazine - 0.47 0.25 22.025 Pyrrolidine, 1-(1-oxopropyl)- 0.31 0.39 - 185 Table B2. N-heterocylic compounds and their relative abundances in HTL aqueous samples detected by gas chromatography â mass spectrometry analysis (continued). Retention time (min) Molecule name Molecule model Relative peak area (%) 290°C-15 min 325°C-15 min 360°C-15 min 20.371 .alpha.-Methyl-.alpha.-propylsuccinimide - 0.32 0.36 21.072 Pyrazine, 2-methyl-5-(1-propenyl)-, (E)- - 0.68 - 29.66 3-Pyrrolidin-2-yl-propionic acid 0.61 - - 21.703 2-Amino-4-methyl-3-pyridinol 0.59 - - 24.929 Benzoxazole, 2-[2-(4-morpholyl)ethyl]thio- - - 0.55 26.824 2-Piperidinemethanol 0.53 - - 26.663 3,6-Dimethylpiperazine-2,5-dione 0.50 - - 17.662 2,4-Imidazolidinedione, 3,5,5-trimethyl- - - 0.50 26.916 Pyridin-2(1H)-one, 1-phenyl-4-(1-piperidylmethylenamino)- 0.50 - - 17.662 2-Amino-5,6-dihydro-4,4,6-trimethyl-4H-1,3-oxazine - 0.48 - 22.484 5-(4-Chlorobutyl)hydantoin 0.44 - - 17.926 2,3-Dimethyl-5-ethylpyrazine 0.43 - - 22.116 3-Piperidinone, 1-ethyl-6-methyl- - - 0.41 17.662 Pyrazole, 1-methyl-4-nitro- 0.35 - - 22.116 Piperidine-2,5-dione - 0.35 - 15.687 Pyrazine, 2-(n-propyl)- 0.13 0.19 - 19.74 2(1H)-Pyridinone, 3,6-dimethyl- 0.10 0.21 - 8.912 Pyrrolidine, 1-methyl- 0.31 - - 19.74 3-Pyridinol, 2,6-dimethyl- - - 0.31 186 Table B2. N-heterocylic compounds and their relative abundances in HTL aqueous samples detected by gas chromatography â mass spectrometry analysis (continued). Retention time (min) Molecule name Molecule model Relative peak area (%) 290°C-15 min 325°C-15 min 360°C-15 min 28.19 1H-Indazole, 3-methyl- - - 0.29 24.654 Piperidine, 1-methyl- - - 0.29 16.72 Ethanone, 1-(1H-pyrrol-2-yl)- 0.25 - - 22.978 4(1H)-Quinazolinone 0.25 - - 11.875 Pyridine, 3-methyl- - - 0.24 18.971 Dexrazoxane - 0.23 - 12.001 Piperidine, 1-ethyl- 0.23 - - 16.238 2-(2-Pyrrolidin-1-yl-ethyl)-pyridine 0.23 - - 25.09 3-Amino-4,6-dimethylpyridone-2(1H) 0.22 - - 23.988 Pyrrolidine, 1-(1-oxobutyl)- - 0.22 - 22.978 Aziridine, 1,2-dimethyl-3-phenyl-, trans- - 0.20 - 21.347 2(1H)-Pyridinethione, 1-methyl- - - 0.19 26.502 4,6-Dimethyl-2-pyrimidone 0.19 - - 20.371 3,4-Methylpropylsuccinimide 0.19 - - 22.978 4-Hydroxy-1,8-naphthyridine - - 0.19 20.865 2H-1,2,3-Triazole-4-carboxylic acid, 2-(2-fluorophenyl)- - 0.19 - 18.752 1-Methyl-3-pyrrolidinol - 0.19 - 28.81 1H-[1,2,3]Triazole-4-carboxylic acid, 1-(4-aminofurazan-3-yl)-5-pyrrolidin-1-ylmethyl- 0.18 - - 25.699 2-Amino-1,1-dicarboxylic acid, N-[3-fluorophenyl]- 0.17 - - 187 Table B2. N-heterocylic compounds and their relative abundances in HTL aqueous samples detected by gas chromatography â mass spectrometry analysis (continued). Retention time (min) Molecule name Molecule model Relative peak area (%) 290°C-15 min 325°C-15 min 360°C-15 min 21.887 Pyrazine, 3,5-dimethyl-2-(1-propenyl)-, (Z)- - 0.17 - 25.021 Pyridine-3-carbohydrazide, 5,6-dichloro-N2-phenyl- 0.15 - - 18.97 5(3)-Carboxamide-4-hydroxy-3(5)-methylpyrazole 0.15 - - 11.932 Piperidine, 1-ethyl- 0.12 - - 17.443 Piperazine, 1-[2-(2,5-dimethyl-1H-pyrrol-1-yl)ethyl]- 0.12 - - : Carbon, : Hydrogen, : Oxygen, :Nitrogen, : Sulfur, : Fluorine, : Chlorine 188 Table B3. Amides and their relative abundances in HTL aqueous samples detected by gas chromatography â mass spectrometry analysis. Retention time (min) Molecule name Molecule model Relative peak area (%) 290°C-15 min 325°C-15 min 360°C-15 min 9.337 Acetamide 3.22 4.21 3.78 10.991 Propanamide 1.27 1.70 1.24 10.417 Acetamide, N-methyl- 0.49 1.20 1.19 27.341 Acetamide, N-(2-phenylethyl)- 1.15 - 0.67 12.242 Propanamide, N-methyl- 0.24 0.55 0.55 18.683 N-(3-Methylbutyl)acetamide 0.25 0.47 0.45 11.634 Acetamide, N-ethyl- 0.29 0.82 - 13.62 Propanamide, N,N-dimethyl- 0.20 0.44 0.39 11.645 Pentanamide, N-ethyl- - - 0.66 25.251 Benzeneacetamide 0.34 - 0.28 14.562 Hexanamide - 0.48 - 14.573 Butanamide - - 0.45 13.987 Formamide, N,N-diethyl- - 0.19 0.22 14.561 Pentanamide 0.34 - - 20.957 N-Acetyl-dl-alanine methylamide - 0.21 - 23.391 N-Benzyl-N-ethyl-p-isopropylbenzamide 0.20 - - 28.202 Formamide, N-phenyl- 0.16 - - 12.173 Propanamide, 2-methyl- 0.13 - - : Carbon, : Hydrogen, : Oxygen, :Nitrogen 189 Table B4. Amines and their relative abundances in HTL aqueous samples detected by gas chromatography â mass spectrometry analysis. Retention time (min) Molecule name Molecule model Relative peak area (%) 290°C-15 min 325°C-15 min 360°C-15 min 18.132 2-Aminophenol 0.53 0.68 0.70 15.469 3-Pyridinamine, N,N-dimethyl- - 1.58 - 15.48 1,2-Benzenediamine, 4-methyl- 1.19 - 1.45 9.165 1-Butanamine, 3-methyl- 0.22 0.43 0.21 18.328 Cycloheptylamine - 0.26 0.28 17.363 Cyclopentanecarboxylic acid, 1-amino- - - 0.44 8.005 Isobutylamine - 0.25 - 16.226 m-Ethylaniline - 0.21 - 8.005 Guanidine, methyl- 0.20 - - 8.315 Ethane-1,2-diimine, N,N'-diamino- - 0.16 - 7.408 Ethanamine, N-methyl- 0.13 - - 11.438 Ethanamine, N,N-dimethyl- 0.11 - - : Carbon, : Hydrogen, : Oxygen, :Nitrogen 190 Table B5. Ketones and their relative abundances in HTL aqueous samples detected by gas chromatography â mass spectrometry analysis. Retention time (min) Molecule name Molecule model Relative peak area (%) 290°C-15 min 325°C-15 min 360°C-15 min 7.684 2-Butanone 0.61 1.07 1.16 12.759 2-Cyclopenten-1-one, 2-methyl- 0.74 0.84 0.69 29.798 3,6-Undecandione 1.94 - - 14.206 2-Cyclopenten-1-one, 3-methyl- 0.33 0.50 0.55 31.107 5-Isopropylidene-3,3-dimethyl-dihydrofuran-2-one - - 0.82 24.952 4-Hydroxy-2,4,5-trimethyl-2,5-cyclohexadien-1-one 0.54 - - 10.221 Cyclopentanone - - 0.46 8.603 2-Pentanone - 0.19 0.22 8.602 2-Butanone, 3-methyl- 0.14 0.23 - 23.575 Cyclopenten-4-one, 1,2,3,3-tetramethyl- - 0.30 - 12.426 Butyrolactone 0.12 0.17 - 18.982 Cyclopentanone, 2-methyl- - - 0.23 17.868 Ethanone, 1-(3-thienyl)- 0.22 - - 15.09 4,4-Dimethyl-2-cyclopenten-1-one - 0.16 - 15.239 2-Cyclopenten-1-one, 2,3-dimethyl- 0.13 - - 15.09 2-Cyclopenten-1-one, 3,4-dimethyl- 0.11 - - : Carbon, : Hydrogen, : Oxygen, : Sulfur 191 Table B6. Other compounds and their relative abundances in HTL aqueous samples detected by gas chromatography â mass spectrometry analysis. Retention time (min) Molecule name Molecule model Relative peak area (%) 290°C-15 min 325°C-15 min 360°C-15 min Phenolic compounds 14.435 Phenol 0.22 0.86 0.82 17.202 p-Cresol 0.18 0.28 0.33 Esters 30.923 l-Valine, N-allyloxycarbonyl-, tetradecyl ester 2.17 - - 29.901 L-Proline, N-(hexanoyl)-, octyl ester 1.89 - - 28.569 l-Alanine, N-allyloxycarbonyl-, pentyl ester 1.81 - - 27.995 Acetic acid, 6-morpholin-4-yl-9-oxobicyclo[3.3.1]non-3-yl ester 0.86 - - 15.893 2-Ethoxyethyl acetate 0.64 - - 28.535 D-Alanine, N-allyloxycarbonyl-, tridecyl ester - - 0.39 28.845 Carbamic acid, 3-methylphenyl-, butyl ester - - 0.22 18.327 2H-Pyran-2-one, tetrahydro-5,6-dimethyl-, trans- 0.16 - - 7.351 Sulfurous acid, dimethyl ester 0.09 - - S-heterocyclic compounds 23.391 3-Methylbenzothiophene - 0.20 - Alcohols 23.563 3-Buten-2-ol, 4-(2,6,6-trimethyl-2-cyclohexen-1-yl)-, (3E)- 0.38 - - Alkenes 17.363 3-Heptyne, 5-methyl- - 0.38 - : Carbon, : Hydrogen, : Oxygen, :Nitrogen, : Sulfur 192 Table B7. Results of the Kolmogorov-Smirnov test for assessing normality of the HTL aqueous characterization data. Parameters Degree of freedom Statistic p-value Test result (α=0.05) Chemical oxygen demand 20 0.1517 0.7245 Can't reject normality Total ammonia 20 0.1791 0.4942 Can't reject normality Dissolved proteins 20 0.2239 0.2315 Can't reject normality Orthophosphate-P 20 0.2064 0.3180 Can't reject normality Total phenol 20 0.2473 0.1457 Can't reject normality Total alkalinity 20 0.1387 0.8490 Can't reject normality pH 20 0.1685 0.5771 Can't reject normality Glycerol 20 0.1016 1.0000 Can't reject normality Sugars 20 0.2226 0.2372 Can't reject normality Acetic acid 20 0.1392 0.8441 Can't reject normality Propionic acid 20 0.0954 1.0000 Can't reject normality Butyric acid 20 0.1008 1.0000 Can't reject normality Valeric acid 20 0.0899 1.0000 Can't reject normality Isovaleric acid 20 0.1829 0.4666 Can't reject normality Total volatile fatty acids 20 0.1918 0.4057 Can't reject normality 193 Appendix C: Positive control test results and characterization of the inoculum used in biochemical assays, contents of biochemical assays, one factor model graphs and normality test results of HTL aqueous biodegradability presented in Chapter 5. Figure C1. Positive control test results for mesophilic-thermophilic biochemical methane potential (n=3), five-day biochemical oxygen demand (n=2), and biochemical hydrogen potential (n=3) 194 (a) (b) (c) (d) (e) (f) Figure C2. One factor model graphs of (a-b) biochemical oxygen demand, (c-d) mesophilic methane potential, and (e-f) thermophilic methane potential. 195 Table C1. Characterization data of aerobic and anaerobic inocula used in the biochemical assays (n=3) Parameters Aerobic inoculum Mesophilic anaerobic inoculum Thermophilic anaerobic inoculum pH 6.4 7.73 8.07 Total solids (wt.%) - 1.69±0.02 1.82±0.01 Volatile solids (wt.%) - 1.15±0.08 1.27±0.01 Total suspended solids (mg/L) 1339±35.8 - - Volatile suspended solids (mg/L) 1112.2±50 - - Total COD (mg/L) 1977±46 20021±518 13646±855 Soluble COD (mg/L) 190±22 608±97 1628±67 BOD5 (mg/L) 940±35.6 - - Ammonium-N (mg/L) 39.5±0.4 1625±24 1493±14 Acetic acid (mg/L) 56.4±3.6 24.4±4 202±4.4 Propionic acid (mg/L) 16.9±3.3 12.9±5.4 10.2±0.1 Butyric acid (mg/L) 389±30 24±3.7 0±0 Total VFA (mg/L) 463±69.8 61.3±13 212±4.6 Total alkalinity (mg CaCO3/L) 188.7 5925±35 5150±71 Total phenol (mg GAE/L) 3.7±0.2 15.7±0.8 42.8±3 COD: Chemical oxygen demand, BOD5: Five-day biochemical oxygen demand, VFA: Volatile fatty acid, GAE: Gallic acid equivalent Table C2. Results of the Kolmogorov-Smirnov test for assessing normality of the HTL aqueous biodegradability data. Parameters Degree of freedom Statistic p-value Test result (α=0.05) Five-day biochemical oxygen demand 20 0.1753 0.5227 Can't reject normality Aerobic biodegradability 20 0.0953 1.0000 Can't reject normality Mesophilic methane potential 19 0.1693 0.6780 Can't reject normality Mesophilic anaerobic biodegradability 19 0.1693 0.6781 Can't reject normality Thermophilic methane potential 19 0.1192 1.0000 Can't reject normality Thermophilic anaerobic biodegradability 19 0.1192 1.0000 Can't reject normality 196 Table C3. Theoretical calculation of the contribution of volatile fatty acids to the methane potential of municipal sludge-derived HTL aqueous samples obtained at various HTL temperatures and retention times. HTL Temperature (°C) 290 300 300 325 325 325 350 350 360 Retention time (min) 15 4.4 25.6 0 15 30 4.4 25.6 15 Acetic acid (mg/L) 7423 6242 7694 8653 7785 7900 7163 7234 7705 Acetic acid based COD in HTL aqueous (mg/L) [COD: 1.067 mg/mg] 7920 6660 8209 9233 8307 8430 7642 7718 8221 Propionic acid (mg/L) 2664 2153 2679 3229 3054 2807 2500 2508 2751 Propionic acid based COD in HTL aqueous (mg/L) [COD: 1.514 mg/mg] 4034 3259 4055 4888 4623 4250 3785 3796 4165 Butyric acid (mg/L) 2354 1568 2026 2385 1889 1740 1864 2177 2201 Butyric acid based COD in HTL aqueous (mg/L) [COD: 1.818 mg/mg] 4280 2850 3683 4337 3434 3163 3388 3958 4001 Valeric acid (mg/L) 1066 663 929 1413 1592 2012 1319 1195 1728 Valeric acid based COD in HTL aqueous (mg/L) [COD: 2.039 mg/mg] 2174 1351 1894 2881 3246 4103 2689 2436 3523 Isovaleric acid (mg/L) 1265 877 932 1334 1425 1627 1421 1494 1730 Iso-valeric acid based COD in HTL aqueous (mg/L) [COD: 2.039 mg/mg] 2578 1788 1900 2720 2906 3317 2898 3046 3528 Total COD caused by VFAs in HTL aqueous (mg/L) 20987 15908 19742 24059 22515 23264 20403 20955 23437 Volume of HTL aqueous in BMP tests (mL) 1.437 1.464 1.418 1.489 1.451 1.587 1.554 1.532 1.601 COD contributed by VFAs in BMP tests (mg) 30.2 23.3 28.0 35.8 32.7 36.9 31.7 32.1 37.5 Expected methane from VFAs (mL CH4) [mg CODVFA Ã 0.35 mL CH4/mg COD] 10.6 8.2 9.8 12.5 11.4 12.9 11.1 11.2 13.1 Measured total methane production (mL) 25.2 22.5 24.4 25.7 23.4 25.4 23.8 21.6 0 Percentage of VFA based methane production (%) 41.9 36.3 40.2 48.8 48.8 50.9 46.7 51.9 0 COD: Chemical oxygen demand, VFA: Volatile fatty acids 197 Table C4. Initial concentrations of chemical oxygen demand, ammonia, phenolics, and volatile fatty acids in aerobic and dark fermentation tests. HTL Temperature (°C) 290 300 300 325 325 325 350 350 360 Retention time (min) 15 4.4 25.6 0 15 30 4.4 25.6 15 COD originating from HTL aqueous in aerobic test (mg/L) 1000 981 1013 965 990 905 925 938 897 COD originating from inoculum in aerobic test (mg/L) - - - - - - - - - Total initial COD in the aerobic test (mg/L) 1000 981 1013 965 990 905 925 938 897 Ammonia originating from HTL aqueous in aerobic test (mg/L) 45 39 45 49 60 59 61 71 64 Ammonia originating from inoculum in aerobic test (mg/L) - - - - - - - - - Total initial ammonia in the aerobic test (mg/L) 45 39 45 49 60 59 61 71 64 Phenolics originating from HTL aqueous in aerobic test (mg/L) 21 23 16 16 17 17 17 15 15 Phenolics originating from inoculum in aerobic test (mg/L) - - - - - - - - - Total initial phenolics in the aerobic test (mg/L) 21 23 16 16 17 17 17 15 15 VFAs originating from HTL aqueous in aerobic test (mg/L) 124 100 124 143 127 124 115 119 127 VFAs originating from inoculum in aerobic test (mg/L) - - - - - - - - - Total initial VFAs in the aerobic test (mg/L) 124 100 124 143 127 124 115 119 127 COD originating from HTL aqueous in dark fermentation test (mg/L) 200 200 200 200 200 200 200 200 200 COD originating from inoculum in dark fermentation test (mg/L) 23 23 23 23 23 23 23 23 23 Total initial COD in the dark fermentation test (mg/L) 223 223 223 223 223 223 223 223 223 Ammonia originating from HTL aqueous in dark fermentation test (mg/L) 225 193 219 252 295 326 337 376 363 Ammonia originating from inoculum in dark fermentation test (mg/L) 61 61 61 61 61 61 61 61 61 Total initial ammonia in the dark fermentation test (mg/L) 286 254 280 313 357 387 398 437 425 Phenolics originating from HTL aqueous in dark fermentation test (mg/L) 107 113 80 81 82 95 92 77 88 Phenolics originating from inoculum in dark fermentation test (mg/L) - - - - - - - - - Total initial phenolics in the dark fermentation test (mg/L) 107 113 80 81 82 95 92 77 88 VFAs originating from HTL aqueous in dark fermentation test (mg/L) 620 490 603 739 629 688 639 635 724 VFAs originating from inoculum in dark fermentation (mg/L) 2.3 2.3 2.3 2.3 2.3 2.3 2.3 2.3 2.3 Total initial VFAs in the dark fermentation test (mg/L) 623 492 605 742 631 690 642 638 726 COD: Chemical oxygen demand, VFA: Volatile fatty acid, -: Insignificant amount (<1 mg/L) 198 Table C5. Initial concentrations of chemical oxygen demand, ammonia, phenolics, and volatile fatty acids in mesophilic and thermophilic biochemical methane potential tests. HTL Temperature (°C) 290 300 300 325 325 325 350 350 360 Retention time (min) 15 4.4 25.6 0 15 30 4.4 25.6 15 COD originating from HTL aqueous in mesophilic AD test (mg/L) 2052 2052 2052 2052 2052 2052 2052 2052 2052 COD originating from inoculum in mesophilic AD test (mg/L) 347 347 347 347 347 347 347 347 347 Total initial COD in the mesophilic AD test (mg/L) 2399 2399 2399 2399 2399 2399 2399 2399 2399 Ammonia originating from HTL aqueous in mesophilic AD test (mg/L) 92 82 91 103 123 134 135 154 145 Ammonia originating from inoculum in mesophilic AD test (mg/L) 929 929 929 929 929 929 929 929 929 Total initial ammonia in the mesophilic AD test (mg/L) 1021 1011 1020 1032 1052 1062 1063 1083 1074 Phenolics originating from HTL aqueous in mesophilic AD test (mg/L) 44 48 33 33 34 39 37 32 35 Phenolics originating from inoculum in mesophilic AD test (mg/L) 9 9 9 9 9 9 9 9 9 Total initial phenolics in the mesophilic AD test (mg/L) 53 57 42 42 43 48 46 41 44 VFAs originating from HTL aqueous in mesophilic AD test (mg/L) 255 208 251 303 263 282 256 261 289 VFAs originating from inoculum in mesophilic AD (mg/L) 35 35 35 35 35 35 35 35 35 Total initial VFAs in the mesophilic AD test (mg/L) 290 243 286 338 298 317 291 296 324 COD originating from HTL aqueous in thermophilic AD test (mg/L) 2274 2274 2274 2274 2274 2274 2274 2274 2274 COD originating from inoculum in thermophilic AD test (mg/L) 930 930 930 930 930 930 930 930 930 Total initial COD in the thermophilic AD test (mg/L) 3204 3204 3204 3204 3204 3204 3204 3204 3204 Ammonia originating from HTL aqueous in thermophilic AD test (mg/L) 102 91 101 115 137 148 149 171 161 Ammonia originating from inoculum in thermophilic AD test (mg/L) 853 853 853 853 853 853 853 853 853 Total initial ammonia in the thermophilic AD test (mg/L) 956 944 954 968 990 1001 1003 1024 1014 Phenolics originating from HTL aqueous in thermophilic AD test (mg/L) 49 53 37 37 38 43 41 35 39 Phenolics originating from inoculum in thermophilic AD test (mg/L) 24 24 24 24 24 24 24 24 24 Total initial phenolics in the thermophilic AD test (mg/L) 73 78 61 61 63 67 65 60 63 VFAs originating from HTL aqueous in thermophilic AD test (mg/L) 283 231 278 336 291 313 283 289 321 VFAs originating from inoculum in thermophilic AD (mg/L) 121 121 121 121 121 121 121 121 121 Total initial VFAs in the thermophilic AD test (mg/L) 404 352 399 457 412 434 405 410 442 COD: Chemical oxygen demand, VFA: Volatile fatty acid, AD: Anaerobic digestion 199 Appendix D: Characterization of aerobic and anaerobic inoculum used in biochemical tests, pH levels during biochemical methane potential tests, feedstock total solids and volatile solids levels, and HTL aqueous contents per volatile solids in Chapter 5. Table D1. Characterization data of aerobic and anaerobic inoculum used in the biochemical assays (n=3). Parameters Aerobic inoculum Mesophilic anaerobic inoculum Thermophilic anaerobic inoculum Total solids (wt.%) - 1.8±0.05 1.63±0.01 Volatile solids (wt.%) - 1.26±0.02 1.2±0.01 Total suspended solids (mg/L) 2120±35 - - Volatile suspended solids (mg/L) 2000±260 - - Ammonia-N (mg/L) 70±3 1570±5 1600±2 pH 7.79 7.98 8.20 Alkalinity (mg CaCO3/L) 280±0 5190±18 5360±18 Acetic acid (mg/L) - 8.1±0.1 72.4±0.2 Propionic acid (mg/L) - 0 0 Butyric acid (mg/L) - 0 0 Total volatile fatty acids (mg/L) - 8.1±0.1 72.4±0.2 Table D2. Total solids and volatile solids levels of municipal sludge feedstocks used in hydrothermal liquefaction (n=3). Sludge origin Sludge type Total solids (wt.%) Volatile solids (wt.%) WWTP 1 Mixed sludge High PS case PS:SS=59:41%* 20 18.0±0.3 Mixed sludge Average case PS:SS=45:55%* 20 18.0±0.3 Mixed sludge High SS case PS:SS=31:69%* 20 16.9±0.3 Digested sludge Thermophilic 26.0±0.2 18.4±0.1 WWTP 2 Mixed sludge High PS case PS:SS=70:30%* 20 17.6±0.3 Mixed sludge Average case PS:SS=57:43%* 20 17.3±0.4 Mixed sludge High SS case PS:SS=39:61%* 20 16.9±0.5 Digested sludge Mesophilic 23.3±0.8 17.6±0.5 WWTP: Wastewater treatment plant, PS: Primary sludge, SS: Secondary sludge, *: Total solids-based ratio 200 (a) (b) (c) (d) Figure D1. Yield of products per gram of volatile solids (VS) of feedstock sludge, including (a) chemical oxygen demand, (b) protein and total ammonia, (c) volatile fatty acids, and (d) orthophosphate-P and total sugar, in hydrothermal liquefaction aqueous phase (n=6). High PS, average, and high SS correspond to PS:SS mixing ratios (wt.% TS) of 59:41%, 45:55%, 31:69% and 70:30%, 57:43%, 39:61% for WWTP 1 and WWTP 2, respectively. 201 (a) (b) Figure D2. pH levels of (a) mesophilic and (b) thermophilic biochemical methane potential reactors throughout the test. 202 Figure D3. Pearson correlation coefficients of feedstock characteristics, HTL aqueous composition, and HTL aqueous biodegradability. High PS, average, and high SS correspond to PS:SS mixing ratios (wt.% TS) of 59:41%, 45:55%, 31:69% and 70:30%, 57:43%, 39:61% for WWTP 1 and WWTP 2, respectively. 203 Appendix E. Two-sample t-test results of Chapter 6 Table E1. Average characterization values of synthetic wastewater and HTL aqueous spiked synthetic wastewater, along with two-sample t-test results for assessing the statistical significance of changes in inflow characterization following HTL aqueous spiking. Parameter HTL aqueous + synthetic wastewater Synthetic wastewater Difference (%) t statistic p-value Statistical difference (α=0.05) COD (mg/L) 494±6.7 416±7.6 +18.8 39.44 <0.0001 Significant BOD5 (mg/L) 310±5.8 255±3.7 +21.9 30.38 <0.0001 Significant TAN (mg/L) 26.5±3 21.3±2.3 +24.5 7.00 <0.0001 Significant Orthophosphate-P (mg/L) 10.4±0.64 9.5±0.7 +8.4 4.26 <0.0001 Significant Total phenolics (mg GAE/L) 6.5±0.6 4±0.5 +63.8 16.58 <0.0001 Significant UV absorbance at 254 nm 560±36 415±24 +35.0 17.25 <0.0001 Significant COD: Chemical oxygen demand, BOD5: Five-day biochemical oxygen demand, TAN: Total ammonium nitrogen, GAE: Gallic acid equivalent, UV: Ultraviolet, HTL: Hydrothermal liquefaction. The numbers following the Â± sign indicate the standard deviation. Table E2. Average operation parameters of reactors treating synthetic wastewater and HTL aqueous spiked synthetic wastewater, accompanied by two-sample t-test results to assess the statistical significance of differences between the reactor operation parameters. Parameter HTL aqueous spiked reactor Control reactor Difference (%) t statistic p-value Statistical difference (α=0.05) pH 7.80±0.09 7.79±0.1 +0.13 1.24 0.2164 Insignificant Temperature (°C) 24.2±0.75 23.6±0.68 +2.5 4.87 <0.0001 Significant DO (mg/L) 2.7±0.58 2.9±0.61 -6.9 -2.09 0.0385 Significant Settling underflow TSS (mg/L) 7174±1522 6684±1240 +7.3 1.27 0.2092 Insignificant MLVSS (mg/L) 2003±293 1777±269 +12.7 4.87 <0.0001 Significant DO: Dissolved oxygen concentration, TSS: Total suspended solids, MLVSS: Mixed liquor volatile suspended solids, HTL: Hydrothermal liquefaction. The numbers following the Â± sign indicate the standard deviation. Table E3. Average chemical oxygen demand removal of HTL aqueous spiked synthetic wastewater treating reactor, and two-sample t-test results to assess the statistical significance of differences between performance at 4.5 h and 6 h hydraulic retention times. Parameter 4.5 h HRT 6 h HRT Difference (%) t statistic p-value Statistical difference (α=0.05) HTL aqueous orginated COD removal (%) 72.5±8.8 79.2±6.5 -8.45 -3.69 0.0004 Significant COD: Chemical oxygen demand, HRT: Hydraulic retention time, HTL: Hydrothermal liquefaction. The numbers following the Â± sign indicate the standard deviation. 204 Table E4. Comparison of effluent characterization from reactors treating synthetic wastewater and HTL aqueous spiked synthetic wastewater, chemical oxygen demand and biochemical oxygen demand removal performances, and sludge volume index at 4.5 h and 6 h hydraulic retention times, along with two-sample t-test results to assess the statistical significance of the differences HTL aqueous spiking caused on the reactor effluent characteristics and treatment efficiencies. Parameter Hydraulic retention time (h) HTL aqueous spiked reactor Control reactor Difference (%) t statistic p-value Statistical difference (α=0.05) Effluent COD (mg/L) 4.5 79.5±7.4 57.1±10 +39.3 10.93 <0.0001 Significant 6 56.3±7.5 38.6±4.9 +45.7 12.04 <0.0001 Significant COD removal (%) 4.5 83.9±1.5 86.4±2.3 -2.9 -5.43 <0.0001 Significant 6 87.7±2.5 90.6±1.2 -3.3 -6.50 <0.0001 Significant Effluent BOD5 (mg/L) 4.5 19.6±3.1 19±5.7 +2.9 0.23 0.8253 Insignificant 6 20.3±3.6 18.8±3.8 +8.0 0.75 0.4651 Insignificant BOD5 removal (%) 4.5 93.7±0.9 92.5±2.2 +1.3 1.35 0.2028 Insignificant 6 93.4±1.2 92.6±1.5 +0.9 1.14 0.2747 Insignificant Effluent TAN concentration (mg/L) 4.5 72.5±4.3 63.6±4.2 +14.0 5.38 <0.0001 Significant 6 70.3±2.6 63.8±2.5 +10.1 6.37 <0.0001 Significant Effluent phosphorus concentration (mg/L) 4.5 9.4±0.36 8.7±0.82 +8.1 2.87 0.0085 Significant 6 9.1±0.38 8.5±0.45 +7.4 3.87 0.0007 Significant Effluent total phenolics concentration (mg GAE/L) 4.5 3.8±0.57 2±0.77 +86.8 6.68 <0.0001 Significant 6 2.8±0.32 1.6±0.34 +74.8 9.21 <0.0001 Significant Sludge volume index (mL/g) 4.5 270±47 186±28 +44.8 5.46 <0.0001 Significant 6 379±62 209±47 +81.0 7.86 <0.0001 Significant Effluent UV absorbance at 254 nm 4.5 397±50 298±29 +32.9 10.35 <0.0001 Significant 6 326±9.2 267±6.6 +22.4 32.08 <0.0001 Significant COD: Chemical oxygen demand, BOD5: Five-day biochemical oxygen demand, TAN: Total ammonium nitrogen, GAE: Gallic acid equivalent, UV: Ultraviolet. HTL: Hydrothermal liquefaction. The numbers following the Â± sign indicate the standard deviation.
Thesis/Dissertation
eng
Civil Engineering
Vancouver : University of British Columbia Library
Attribution-NonCommercial-NoDerivatives 4.0 International
Graduate
Eskicioglu, Cigdem
Evaluation of on-site biological treatment options for hydrothermal liquefaction aqueous phase derived from sludge in municipal wastewater treatment plants
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