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Biological Removal of Chloroform in a Controlled Trickle Bed Air Biofilter Under Acidic Conditions

Biological Removal of Chloroform in a Controlled Trickle Bed Air Biofilter Under Acidic Conditions

Biological Removal of in a Controlled Trickle Bed Air Biofilter under Acidic Conditions

A thesis submitted to the

Division of Research and Advanced Studies of the University of Cincinnati

In partial fulfillment of the Requirements for the degree of

MASTER OF SCIENCE

In the Department of Biomedical, Chemical and Environmental Engineering of the College of Engineering and Applied Sciences

November 2016

By

Keerthisaranya Palanisamy B. Tech. Energy and Environmental Engineering, Tamil Nadu Agricultural University, Coimbatore, India, 2012

Thesis Committee Members

Dr. George A. Sorial (Chair) Dr. Endalkachew Sahle-Demessie Dr. Soryong (Ryan) Chae

Abstract

This dissertation includes a detailed review on the application of two different biofiltration systems, 1.

Classical biofilters and 2. Biotrickling filters, for the removal of various air pollutants. The biofiltration of volatile organic compounds, fuel emissions, biogas, off and malodorous gases are studied in detail.

The removal mechanisms including aerobic and anaerobic processes involved in the breakdown of the phase contaminants within natural and synthetic biofiltration units are studied. This is followed by an experimental evaluation of the performance of a controlled trickle bed air biofilter (TBAB) for the removal of chloroform.

Chloroform is a higher chlorinated and a highly volatile compound. It is directly emitted into the atmosphere through several industrial sources such as pharmaceutical, pulp and paper industries. It is also formed as a byproduct of chlorination disinfection in and wastewater sources and subsequently volatilized into the atmosphere. Chloroform, also known as trichloromethane, contains three atoms making it very recalcitrant and stable in the environment. It is also a highly hydrophobic compound making it an extremely difficult candidate for biodegradation. In this study, this challenge is overcome by evaluating the biodegradation potential of chloroform in the presence of a cometabolite, , and using filamentous fungi as the principle biodegrading consortium within the TBAB.

The TBAB receives chloroform and ethanol in its gas phase and enriching nutrients in a buffered liquid phase. Chloroform and Ethanol were supplied at different feed ratios including 1:5, 1:10, 1:20, 1:30, 1:40 with ethanol concentrations ranging from 25 to 200 ppmv. A removal efficiency of 80.9% was observed

3 when treating 5 ppmv of chloroform with 200 ppmv of ethanol and an elimination capacity of 0.238 g/m .h was achieved. The study further extends to the determination of the removal kinetics for chloroform and quantifying the COD consumption and utilization rates of filamentous fungi species. In addition, an extensive mass balance is performed to identify the usage of carbon within the biofilter.

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The thesis is concluded by evaluating all points for improvement and drawing vital conclusions from the experimental research. Future recommendations for the enhanced treatment of chloroform in the trickle bed air biofilter are also highlighted.

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Acknowledgment

I am extremely grateful for the opportunity of pursuing my Master’s research under the guidance of Dr.

George A. Sorial. I extend by deepest appreciation and many thanks to Dr. Sorial for taking me in as part of his research team and offering me the much needed guidance and support. I would also like to thank

Dr. Endalkachew Sahle - Demessie, Mr. Bineyam Mezgebe and Mr. Dhawal Chheda for having helped me during many phases of my research and for their invaluable inputs. I am thankful to Dr. Soryong Chae for his time and for his constructive feedback.

I would like to mention, that my Master’s would not have been possible without my family and friends. I would like to show my utmost gratitude to my friends for their encouragement and moral support. Last but not the least, I wish to mention that my parents and my brother were instrumental for the completion of my Master’s program at the University of Cincinnati and would like to sincerely thank them for this wonderful experience.

Thank you all!

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Table of Contents

Abstract ...... ii Acknowledgment ...... v Table of Contents ...... vi List of Figures ...... viii List of Tables ...... 1 1 Introduction ...... 2 1.1 Background ...... 2 1.2 Objective of Study ...... 2 1.3 Structure of Thesis ...... 3 2 A review on the application of biofilters and biotrickling filters for the removal of air pollutants ...... 4 Abstract ...... 4 Keywords ...... 4 2.1 Introduction ...... 4 2.2 Operational parameters of biofiltration systems ...... 6 2.2.1 Biodegrading media ...... 8 2.2.2 System configuration ...... 9 2.2.3 Contaminant flow mode ...... 10 2.2.4 Development of microbial population ...... 11 2.3 Application of Biofilters ...... 12 2.3.1 Aerobic and Anaerobic degradation ...... 12 2.3.2 Cooxidation and Cometabolism ...... 15 2.3.3 Reductive Dehalogenation and ...... 17 2.3.4 Nitrification and Denitrification ...... 19 2.3.5 Control ...... 21 2.4 A comparison between classical biofilters and biotrickling filters ...... 24 2.5 Conclusion ...... 25 2.6 References ...... 31 3 Biofiltration of chloroform in a trickle bed air biofilter under acidic conditions ...... 43 Abstract ...... 43 Keywords ...... 43 3.1 Introduction ...... 43 3.2 Materials and Methods ...... 48

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3.2.1 Chemicals ...... 48 3.2.2 Trickle bed air biofilter (TBAB) ...... 48 3.2.3 Analytical methods ...... 49 3.2.3.1 Gas Sampling ...... 49 3.2.3.2 Liquid Sampling ...... 50 3.3 Results and Discussion ...... 50 3.3.1 TBAB Performance ...... 51 3.3.2 Reaction Rate Kinetics ...... 52 3.3.3 Carbon Mass Balance ...... 53

3.3.4 Nitrogen utilization and COD reduction ...... 55 3.4 Conclusion ...... 56 3.5 References ...... 63 4 Conclusion and Future Recommendations ...... 68 4.1 Summary ...... 68 4.2 Future Recommendations ...... 68 4.3 References ...... 69

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List of Figures

Figure 2.1: A schematic representation of the removal mechanisms in biofilters ...... 27 Figure 3.1: Schematic Diagram of the trickle bed air biofilter ...... 58 Figure 3.2: Surface Imagery of filamentous fungi produced by SEM analysis under 2µm magnification . 59 Figure 3.3: Surface Imagery of filamentous fungi produced by SEM analysis under 100 µm magnification ...... 59 Figure 3.4: Performance of the TBAB with sequential time for the removal of chloroform at pH 4 ...... 60 Figure 3.5: Elimination Capacity of chloroform vs. total VOCs loading rates ...... 61 Figure 3.6: Reaction rate constants for chloroform vs. total VOCs loading rates ...... 61 Figure 3.7: Cumulative carbon input/output as CO2 equivalent for the TBAB ...... 62 Figure 3.8: COD removal/N utilization vs. total VOCs loading rates for TBABs ...... 62

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List of Tables

Table 2.1: Types of biofiltration systems categorized based on design parameters………………………26 Table 2.2: Application of biofilters for air pollutants removal……………………………………………28 Table 3.1: Operating conditions and performance of TBAB under continuous loading conditions degrading chloroform and ethanol at pH 4………………………………………………………………..58

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1 Introduction

1.1 Background

Biofiltration is a biological treatment process that is principally used for the removal of airborne contaminants including odor compounds, volatile organic compounds (VOCs) and fuel emission . It involves passing the contaminant fluid through a porous filter media supporting the growth of microorganisms that biodegrade the contaminants that come in contact with them. This paper investigates the use of two biofiltration systems: 1) Classical biofilters and 2) biotrickling filters and their treatment applications. The former are natural biofilters packed with organic media that range from soil to oyster shells, while the latter refers to biofilters that are intermittently supplied with liquid nutrients to enrich microbial activity, and packed with synthetic or inorganic media. The different aerobic and anaerobic processes involved in the removal of volatile organic compounds in biofilters are explained with several case studies. In addition, a brief comparison between the two biofiltration systems is performed.

Following this comprehensive study on biofiltration systems, the thesis focuses on a specific application of a trickle bed air biofilter (TBAB) for the removal of chloroform. Chloroform is a persistent higher chlorinated methane discharged as a hazardous air pollutant (HAP) from chemical, pharmaceuticals and paper and pulp industries. It is also formed as a disinfection byproduct in pretreated water that is subjected to chlorination disinfection in water treatment facilities. Due to its universal presence, occurrence in dilute concentrations, recalcitrance to biodegradation and carcinogenicity, gas phase chloroform was chosen for the study and was subjected to microbial oxidation in a TBAB.

1.2 Objective of Study

Therefore the objective of this study is to,

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1. Perform a detailed review on the different types of biofiltration systems - a) Classical biofilters and b)

Biotrickling filters available in the industry, followed by their respective applications in the removal of air borne contaminants.

2. Evaluate the performance of a trickle bed air biofilter (TBAB) for the cometabolic biodegradation of chloroform under acidic conditions using filamentous fungi consortia.

1.3 Structure of Thesis

The structure of this study involves

Chapter 1 provides an introduction to the thesis that encompasses a literature review of two types of gas biofiltration units used in air pollution control and also an experimental study on the performance of a trickle bed air biofilter for the removal of atmospheric chloroform.

Chapter 2 provides a detailed review on classical biofilters and biotrickling filters, their design configurations, and their application as an air pollution control technology for the removal of a variety of compounds.

Chapter 3 demonstrates the performance of a trickle bed air biofilter for the removal of chloroform through cometabolism with ethanol, under acidic conditions, and using filamentous fungi as the biodegrading microorganism

Chapter 4 enlists the major conclusions drawn from the experimental research and future recommendations for the enhanced performance of the TBAB

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2 A review on the application of biofilters and biotrickling filters for the removal of air pollutants

Abstract

This paper reviews the application of biofiltration as an air pollution control (APC) technology. The study discusses the different design aspects and operational parameters of classical biofilters and biotrickling filters. A detailed study on the fate of air pollutants in both biofiltration systems and their ensuing removal mechanisms are highlighted. The removal of air borne contaminants such as aromatic and chlorinated hydrocarbons, methane, , nitrogen oxides and sulphide under different environmental conditions in biofiltration systems is explained using case studies. Aerobic processes such as co-oxidation and nitrification; anaerobic processes such as denitrification, reductive dehalogenation and halorespiration; and odor control are discussed in detail. Previous publications on the biofiltration of air pollutants are summarized and a brief comparison is made between the two biofiltration systems.

Keywords

Biofilters, biotrickling filters, volatile organic compounds, removal mechanisms, design concepts

2.1 Introduction

The application of biofiltration as a biological treatment technology dates back to the 1950s. The first soil bed biofilter was installed in 1959 in a sewage treatment plant in Nuremberg, West Germany. In 1960, several soil filters were installed in a municipal wastewater treatment facility in Seattle for odor control

(Leson and Winer 1991). Since then, numerous studies have been performed to understand the underlying concept of biofiltration and the treatment technique has significantly evolved over the years. Biofiltration involves a bed of porous and moist media to support the growth of microorganisms that biodegrade the fluid contaminants that percolate through the media. The microorganisms remain static, impregnated on the filter medium, while the contaminant fluid is mobile, flowing through the filter (Cohen 2001).

McNEVIN and Barford (2000a) refer biofiltration systems as organic perfusion columns with three phases in intimate contact with each other, a solid organic phase, a liquid and/or a gas phase. Biofiltration

4 principally depend on two main processes: 1. the bioavailability of contaminants to the microorganisms, and 2. the bioconversion of contaminants to biomass and metabolic end products (Malhautier et al. 2005).

Biofiltration is broadly classified into four different types namely, bioscrubbers, classical biofilters, biotrickling filters and membrane bioreactors, of which classical biofilters and biotrickling filters will be the focus of this review. The aforementioned mechanisms are fixed film growth systems that are characterized by the development of biological slimes, often termed as biofilms, on the filter media

(Verma et al. 2006). Biofilm is a complex consortium of microorganisms that attach to the surface of the support media based on environmental and genetic factors and form microcolonies (H. H. Cox and

Deshusses 1998). These microcolonies gradually develop into a complex morphology producing an exopolysaccharide matrix, thus forming a mature biofilm (Verma et al. 2006). The biofilm structure and properties dictate the performance of the biofiltration systems.

Classical biofilters are used for the treatment of gas phase contaminants, malodorants from industrial emissions and waste treatment facilities. The contaminants that come in contact with the biofilm are absorbed and oxidized to end products such as carbon di-oxide, water, inorganic salts and more biomass

(Swanson and Loehr 1997). Advantages of using biofilters as opposed to other air pollution control technologies are that they are a 1. Green technology with low operation and maintenance cost, 2. Capable of handling a variety of compounds and 3. Require very little media replacement. Some disadvantages include 1. Frequent nutrient application and 2. The possibility of the biomass shedding when treating contaminants in lower concentrations. Certain gas phase biofilters also receive intermittent application of liquid nutrients for enhanced removal or moisture control and are termed as biotrickling filters.

Biotrickling filters are a preferred form of treatment for industrial emissions because of its 1. Lower pressure drop, 2. Better operational control, 3. High internal biomass control, 4. Low to moderate water, power consumption and carbon footprint , while it also develops operational issues such as media clogging and prolonged biomass acclimation periods (G. Wu et al. 1999).

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The process of biodegradation follows the fundamental Monod’s relationship that correlates the biodegradation rate to the concentration of the (McNEVIN and Barford 2000b). The performance of the biofiltration systems is assessed by the removal efficiency which is defined by the contaminant removed to the total contaminant applied, expressed in percentages. Other performance parameters include the biofilter/biotrickling filter removal rate or the elimination rate which is the mass of the contaminant removed per unit volume of the biofilter per unit time expressed in (g/m3.h). The elimination rate can also be defined on a unit mass basis or unit surface area basis (McNEVIN and

Barford 2000b). The surface loading of the biofiltration systems is the volumetric gas loading applied to the biofilter, expressed in (m/h) and the mass loading is the mass of contaminant that is applied to unit volume of the filter medium per unit time, expressed in (g/m3.h) (Swanson and Loehr 1997). The most important design parameter of any biofiltration system is the filter media contact time, which is the relative measure of gas residence time within the biofilter, expressed in time units (Swanson and Loehr

1997).

In this paper, the two different types of biofiltration systems: 1. classical biofilters and 2. biotrickling filters are discussed with respect to the different design configurations used in the industry. The reaction mechanisms taking place within gas phase biofilters is highlighted with respect to the removal of chlorinated hydrocarbons, aromatics, nitrogen oxide, ammonia and odorous compounds. This is followed by a tabulated summary of the case studies discussed in the review and concluded by stating the observed contrasts between the two biofiltration systems.

2.2 Operational parameters of biofiltration systems

Classical biofilters and biotrickling filters are designed in different configurations depending on the contaminants to be treated, spatial constraints, growth environment for microorganisms, and economic feasibility. Classical biofilters are used to treat gas streams with odor compounds and volatile organic compounds by passing them through a natural media bed. The substrates are conveyed with the gas flow into a thin layer of moisture, termed as the biofilm. The substrates diffuse through the gaseous boundary

6 layer and contact the water surrounding the biofilm (Cohen 2001). The substrates dissolve in the aqueous medium and then diffuse through the liquid towards the consumption sites within the biofilm (Guieysse et al. 2008). These biofilters can maintain low pressure drops and are suitable to treat large volumes of off- gas streams having low contaminant concentrations (Guieysse et al. 2008; Mudliar et al. 2010). They are reported to show problems with pH control and varying humidity levels in the incoming air (Mudliar et al. 2010). To maintain an optimum water level in the system, air is passed through a prehumidifier that in addition to maintaining adequate moisture, also removes the particulate matter in the gas stream preventing clogging of the filter bed (Iranpour et al. 2005). During high humid conditions or high temperatures that are unsuitable for the microorganisms, classical biofilters are equipped with heat exchangers to maintain lower temperatures (Iranpour et al. 2005).

Similar to biofilters, biotrickling filters also treat gaseous substrates; however they slightly differ from classical biofilters in their basic construction. Biotrickling filters are subjected to the intermittent supply of liquid nutrients to the filter media. As a result, biotrickling filters produce secondary waste stream and possibly develop excess biomass inside the biotrickling filter. The microorganisms form a wet and thick biofilm layer on the surface of the filter media. The micronutrients are continuously distributed and recirculated at the top of the reactor to percolate through gravity (Guieysse et al. 2008). The contaminants are often adsorbed onto the aqueous phase of the biofilm and further biodegraded within the biofilm.

Studies have applied biotrickling filters for the removal of high concentration VOCs and H2S since they allow high surface volumetric loading rates. This is due to the fact that biotrickling filters have a dense biofilm suspended to several millimeters from the surface of the media that is consistently in contact with the VOCs, favoring rapid adsorption and degradation of the substrates (Nicolai and Janni 2001). Due to the continuous trickling of substrates and nutrients, they are more suitable to treat hydrophilic organics.

Table 2.1. gives a list of case studies and experimental research work on the two biofiltration units. It is categorized by the contaminant flow direction, filter media and overall operational setup and will be discussed in detail. From the table we can deduce that 1. Natural or organic media is universally applied

7 in classical biofilters and inert or synthetic media in biotrickling filters, 2. Biotrickling filters are invariably closed systems unlike certain classical biofilters that are open, discharging treated gas streams into the atmosphere, 3. The contaminant gas flow mode is both upflow and downflow in either biofiltration units, 4. Within biotrickling filters, the flow of nutrients and supplements is predominantly co-current with the contaminant gas stream. Additionally, the development of micro-organisms was studied in many biofiltration systems. The most common technique followed for biomass growth in both classical and biotrickling filters is inoculation.

2.2.1 Biodegrading media Filter media is a key component in both biofiltration systems as they influence the surface area of contact, the void ratio, mechanical strength, the wettable fraction of the biofilm, resistance to clogging and a clear passage for the contaminants (Verma et al. 2006). The purpose of the packing material is to facilitate the gas flow, to enable gas/liquid contact and to offer a surface for microbial colonization. Media can be packed in random or in structured fashion, and can enable cross, vertical and random flow modes depending on the hydraulic surface loading and the wetting characteristics of the filter (Verma et al. 2006;

Eding et al. 2006). Classical biofilters predominantly use natural media such as compost, soil, peat, wood shavings, barks, oyster shells and lava rocks. Bioactive media is used to facilitate microbial activity even at lower-substrate concentrations (Guieysse et al. 2008). The biofilters originally built in the U.S. were mostly soil beds in which biologically active mineral soils were used as filter materials (McNEVIN and

Barford 2000b). Several contaminant biotransformations create an acidic environment inside biofilters, therefore materials like peat that work in narrow pH bands provide good removal performance

(Malhautier et al. 2005).

In biotrickling filters, microorganisms are grown on inert synthetic materials such as plastics, , polyurethane foam, ceramic structured packings, celite, activated carbon, diatomaceous earth pellets or mixtures of different materials with specific surface areas ranging between 100 and 300 m-1 (Mudliar et al. 2010). Engineered media are reported to help maintain low EBRT in biotrickling filters, typically

8 between 30 and 45 s compared to 60-90 s when using organic media. This translates to a smaller-footprint system. The surface characteristics of media also play an influential role in biodegradation rates.

Hydrophobic surfaces such as compost, peat, wood chips and other organic compounds are generally difficult to rewet when dry (Bohn and Bohn 1999). Hydrophilic surfaces found in inorganic media naturally fluctuates the moisture content in them, maintaining a uniform moisture profile before and after drying (Bohn and Bohn 1999). As a result, often times, hydrophobic and hydrophilic media are mixed in order to maintain moisture content and improve bulk (Bohn and Bohn 1999). Natural filter media often pose operational problems such as drying up of the bed, acidification and the constant need for media turning over and renewal. On the contrary, synthetic media in biotrickling filters are continuously supplied with the necessary nutrients making them suitable for many contaminants and easier to control.

Due to this phenomenon, they have a long-packing lifetime and better operational stability (J. Van

Groenestijn and Kraakman 2005).

2.2.2 System configuration There are two different configurations that the biofilters can be installed in. An open design configuration is generally applied to natural biofilters, with ascending gas flows. They are installed outside the off-gas generating units and are open to all whether conditions (Mudliar et al. 2010). They occupy a large area and can be single or multiple bed systems with layers of same/different media stacked on top of one another. Most industrial biofilters in the U.S. fall under the open single bed configuration. Multi-story containerized systems are constructed when there is a space constraint for the installation of several single bed systems in series. Decomposition and mineralization of organic material will lead to more volume and compression of the filter material which will increase back pressure in natural biofilters (McNEVIN and Barford 2000b). This is overcome by turning over the filter media once every two years in an open system which also helps maintain the porosity of the material.

A closed design biotrickling filter consists of upflow or downflow gas flows within enclosed rooms and are known to occupy lesser space than the open filters. Almost all biotrickling filters have a closed design

9 configuration. Advantages of closed systems include low vulnerability to extreme climatic conditions and minimum requirement of maintenance (McNEVIN and Barford 2000b).These systems provide greater contaminant loading on a smaller treatment area and generally allow the treated gas to be conveyed through a stack that enables vertical dispersion. Manufacturers have designed filter media like granular and activated carbon with extended durability for closed systems that can function efficiently for 5 consecutive years or more without being turned over (McNEVIN and Barford 2000b; Mudliar et al.

2010). Although, replacement of the filter media once in 10 years is ideally required to maintain good removal performance.

2.2.3 Contaminant flow mode Biofiltration designs are also classified based on the direction of flow of substrates. The gas flow can be both upflow and downflow and is often interchangeable. In certain classical biofilters, the gases move in an upflow direction and are transferred into the biofilter through slotted concrete slabs with distribution canals, entering the filter bed (McNEVIN and Barford 2000b). Radial blowers are generally used to overcome the back pressure caused by the filter. Down-flow systems have also been used in several installations (McNEVIN and Barford 2000b).

In biotrickling filters, depending on the direction of the gas flow, the supply of nutrients is either co- current or counter current to the substrates. In both the scenarios the nutrient distribution and recirculation is key as the microorganisms begin to consume the substrate only after the VOC diffuses into the liquid film. Research indicates that an increase in the liquid flow will result in a proportional increase in the surface area for mass transfer of gas to the liquid phase (Mudliar et al. 2010). It has been reported that there is no significant difference in the performance of the biotrickling filter between the applications of counter or co-current flow modes (H. Cox and Deshusses 1999). However, certain studies reveal that co- current flow of the substrate and nutrients will enhance moisture control and nutrients distribution showing linear removal patterns with the depth of the biotrickling filter. Jin et al. (2005) suggested that irrespective of the flow mode, the portion of the biotrickling filter where the substrates are introduced

10 tend to develop an excessive biomass growth and reported 40-80% efficiency for hydrogen sulphide

(H2S) removal at the inlet section of the biotrickling filter in a co-current flow mode. Therefore, it can be stated that switching between both the flow setups can be performed in order to get a homogenous growth of the microbial population through the depth of the biotrickling filter (Wright et al. 2005; Znad et al.

2007; Hassan and Sorial 2009).

2.2.4 Development of microbial population The development of biocatalysts within the filter bed is initiated through 1) the natural growth of indigenous microorganisms and 2) the inoculation of and fungi from conventional activated sludge processes or previous laboratory studies. While conventional fixed film systems treat multicomponent contaminant streams, artificially impregnated biofilms are able to handle dilute streams and xenobiotic compounds (Verma et al. 2006). Although native organisms are found to thrive well in certain classical biofilters, microbial immobilization is more pronounced in both the filtration systems.

The immobilization of microorganisms can be carried out in two different processes namely attached growth process that involves self-attachment of the microorganisms to the filter material and artificial entrapment of microorganisms within polymer beds, or cellulose nitrate capsules (Verma et al.

2006). Encapsulated cells have provided excellent removal efficiency for aromatic and aliphatic compounds (Verma et al. 2006). Biofiltration systems with microorganisms encapsulated or immobilized on a fixed membrane are known to provide enhanced COD removal than conventional systems (Verma et al. 2006). Following the development of a microbiome, biodegradation is performed by a complex ecosystem of degraders, competitors and predators within the biofilm (Mudliar et al. 2010). The biological processes involved within the biofilter are the 1) transport of the biomass to the filter media surface achieved by diffusion, convection, sedimentation due to gravity, and active mobility of the microorganisms, 2) attachment of biomass depending on the type and concentration of the influent substrate, surface properties of the filter media, and a combination of van der Waal’s and electrostatic forces, 3) growth of microorganisms which relies on the rate of substrate utilization within a biofilm, and

4) decay and detachment of the biomass from the filter bed (Srivastava and Majumder 2008). Parameters

11 that dictate the process of biofilm detachment include dry density, wet density, the content of the extra cellular biopolymer (ECP), and shear stress (Eding et al. 2006). However, the correlation between biodegradation and microbial activity is difficult to determine. In order to better understand the limitations of biofiltration systems, it will be helpful to investigate the flexibility of such complex ecosystems and corroborate the impact of environmental changes such as operating parameters, physicochemical characteristics of contaminants with the activities and diversity of the microbial communities (Malhautier et al. 2005).

2.3 Application of Biofilters

Biofiltration is a APC technology applied for the removal of VOCs from manufacturing, chemical and pharmaceutical industries; malodourous compounds from animal rendering, composting, food processing, wastewater treatment facilities and landfill sites. The gas phase contaminants in biofilters are biodegraded in either aerobic or anaerobic conditions depending upon the contaminant and the application. The common removal mechanisms include cometabolism, cooxidation, dehalogenation, nitrification and sulphate reduction. A flow chart of the different processes included in this review is shown in Fig 2.1.

The following section elaborates on the industrial application of biofilters for the removal of VOCs and odor causing compounds categorized based on removal mechanisms. The biofiltration studies discussed below are summarized in Table 2.2.

2.3.1 Aerobic and Anaerobic degradation is an important macronutrient to microorganisms as they constitute 20% (w/w) of the dry cell

(Veillette et al. 2012). Biodegradation occurs naturally when microorganisms oxidize organic compounds in the environment for energy and growth, using oxygen as a terminal electron accepter in a coupled reaction (Pavlostathis et al. 2003). In this enzymatic process, oxygen is catabolically reduced to water (B. Lee et al. 1997). Examples include the oxidative conversion of aromatic hydrocarbons like , , etc. to and subsequent oxidation to carbon-di-oxide by monooxygenase or dioxygenase (Jindrova et al. 2002). A study performed to biodegrade xylene vapors through

12 an inoculated, peat based biofilter concluded that xylene degradation was entirely due to aerobic degradation negating any option of adsorption or incomplete oxidation (Jorio et al. 2000). Chlorinated compounds are readily metabolized to and free chlorine; often lower chlorinated compounds are suitable substrates for aerobic degradation (J. Field and R. Sierra-Alvarez 2004). A study revealed the use of a biotrickling filter for the removal of chlorobenzene by the use of strain pseudomonas putida at an EBRT of 109 s and an elimination capacity of 234 g/ (m3·h) (Arowolo 2007). While (DCM) is known to be directly oxidized by aerobic bacteria, trichloromethane (TCM) and (TCE) are known to be oxidized only through cometabolism, which is consistent with past literature (Ergas et al. 1995; Furukawa 2003; Ménard et al. 2012). In a study, DCM was treated in a biotrickling filter using lava rock and hyphomicrobium strains in the presence of silicone as a surfactant (Bailón et al. 2009). When a group of hydrocarbons are introduced to the biofilter, the order of biodegradation follows the potential biodegradability of the compounds involved (Malhautier et al. 2005).

Microorganisms tend to use the substrates that provide the most abundant energy input (Malhautier et al.

2005). For instance, oxygenated compounds are metabolized prior to aromatic or halogenated compounds

(Malhautier et al. 2005). Ethyl acetate is degraded before toluene which is degraded before DCM. One of the drawbacks of aerobic degradation in biofilters is the depletion of oxygen within the biofilm. This is due to the fact that oxygen is mostly in the gas phase than the liquid phase and its diffusion into the biofilm could be limited due to several reasons such as biofilter overloading and clogging of biomass

(DEVINNY et al. 1995). The oxygen deficient condition affects the removal performance by making the system anaerobic and also causing nuisance . In a study, a biofilter was subjected to the treatment of landfill gas from a passively vented landfill. Methane was removed at a maximum rate of 80 g/(m3.h).

The continuous and non-oscillating landfill gas emission created an oxygen-restricted environment which affected the removal rate negatively with methane influx and flowrate as a measure of oxygen displacement (Gebert and Gröngröft 2006). Another study where oxygen was consistently mixed with the influent landfill gas and treated in a fine-grained compost biofilter showed a high degradation rate of methane at 63 g/(m3.h) at a mean methane concentration of 2.5% (v/v) (Streese and Stegmann 2003). A

13 biofilter subjected to the removal of H2S from the biogas generated from sulphate rich wastewater performed well with increasing air mix ratio (lower biogas-higher air). Under an operating pH of 1-4.5, and a biogas-air ratio of 1:4, the observed removal efficiencies were 94.7%, 87.3%, 85.6% at retention times of 160, 80 and 40 s consistently eliminating around 500 ppm of H2S (Chaiprapat et al. 2011).

Anaerobic degradation is a cometabolic process with slower rates of degradation when compared to aerobic systems. Under anaerobic conditions, hydrogen is the electron donor and carbon dioxide is the carbon source (B. D. Lee et al. 1999). The contaminants serve as the electron acceptors, being reduced by microorganisms to less hazardous compounds, instead of oxidation such as in the aerobic environment. It is common phenomenon to develop an anaerobic zone at the base of a dominantly aerobic biofilter

(Devinny et al. 1995). Anaerobic degradation is an economical and operationally competitive technique for the removal of chlorinated compounds (B. D. Lee et al. 1999). Based on previous literatures, biofiltration of oxidized chlorinated compounds such as (CT) and perchloroethylene

(PCE) are performed by microbes growing on nitrate, sulphate or carbon-di-oxide as alternate electron acceptors predominantly under anaerobic conditions (B. D. Lee et al. 1999). A study demonstrated the removal of CT in a compost biofilter under methanogenic conditions. Hydrogen and carbon dioxide were provided as the source of energy and carbon. CT was fed at an inlet concentration range of 20-700 ppbv maintaining an EBRT of 2.8 min. 75% of CT was reported to be treated, while the biofilter failed to achieve a higher RE for concentrations >500 ppbv (B. D. Lee et al. 1999). Compost was previously utilized by several authors as a common medium for anaerobic biofilters to remove (NO) from off-gas streams. A compost biofilter was used to anaerobically biodegrade nitric oxide at concentrations from 0 to 500 ppmv and was shown to achieve more than 90% RE for NO gases. With a supply of more than 250 ppmv of NO, the removal rates were maintained high by providing the biofilter with supplemental carbon source such as glucose, lactate, molasses and acetate (B. Lee et al. 1997). A microaerophilic condition within the bed media facilitates the growth of hydrogenotrophic methanogens that are dominant within biofilms in anoxic, denitrifying conditions and favor the oxidation of H2S (Pirolli

14 et al. 2016). A biotrickling filter designed for H2S removal from biogas stream was investigated.

Wastewater effluent from a denitrification tank was used as a source of nutrients and electron acceptors.

3 A Maximum RE of 99.8% was attained with the highest elimination capacity of H2S at 1,509 g/(m .h).

Methane was released from the filtered biogas stream at the rate of 3.8 ± 1.68 g/m3 throughout the period of study (Pirolli et al. 2016).

2.3.2 Cooxidation and Cometabolism Cometabolism refers to the fortuitous degradation of a complex aliphatic by microorganisms that release enzymes of relaxed specificity while growing on a different, often simpler compound called the primary substrate. The extracellular enzymes that are expressed by the biocatalysts on consuming the primary substrate are essential for the of the target substrate. The primary substrate is generally the electron donating substrate with different electron acceptors depending on the system environment. The chlorinated VOCs such as TCE and TCM, chloroform (CF) and CT are known to biodegrade through cometabolism. such as , toluene dioxygenase, toluene monooxygenase and hydroxylase are reported to cometabolize TCE (Borja et al. 2005). To facilitate cometabolic reactions, a suitable primary substrate must be supplied at an optimum flow rate, neither at a very low gas flux, creating insufficient supply of primary substrate, nor at a high gas flow creating insufficient residence time. Hence the type and the flow rate of the primary substrate are critical in the process. Chheda and Sorial (2016) performed ternary studies on TCE, toluene and in a biotrickling filter where methanol served as a cometabolite in a fungal based system (Chheda and Sorial

2016). Polychlorinated Biphenyls (PCBs) are degraded through chlorine substitution, and are transformed by co-metabolic degradation using biphenyl catabolizing enzymes to form chlorobenzoates (Furukawa

2003). A study used pentane oxidizing bacteria, Pseudomonas aeruginosa, to treat methyl tert-butyl vapors cometabolically using pentane in a vermiculite biofilter (Malhautier et al. 2005). Another study revealed that the removal of a hydrophobic compound such as n-hexane is influenced by a hydrophilic compound, methanol, through cometabolism (Zehraoui et al. 2012). Pharmaceutical industries often emitting a mixture of compounds, naturally express cometabolism, thereby enhancing the treatment of

15 less biodegradable compounds when subjected to gas phase biofiltration. Under anaerobic conditions, cometabolism takes place when reduced cofactors react with chlorinated compounds leading to their (J. A. Field and R. Sierra-Alvarez 2004). CT is biodegraded under anaerobic conditions to CF, DCM and traces of (CM) by methanogens, acetogens and fermentative bacteria and this process does not entail any benefit to the microorganisms responsible for the degradation, suggesting cometabolism (Holliger and Schraa 1994; J. Field and R. Sierra-Alvarez

2004).

A special type of cometabolism involves the oxidation of halogenated compounds by monooxygenases called as co-oxidation. Cooxidation refers to the process in which a microorganism oxidizes a substance without being able to utilize the energy derived from this oxidation to support growth (Horvath 1972). A typical example is the oxidation of chlorinated by a methane monooxygenase expressed by methylotrophic organisms that oxidize methane (J. Field and R. Sierra-Alvarez 2004). Biofilters have been reported to treat a wide range of VOCs using the methanogenic bacterium. Soluble methane monooxygenases express non-specific substrate affinity and produce excellent removal performance in varying temperatures, and nutrient supply conditions (Gebert et al. 2008; Ménard et al. 2012). Hence, significant research has been done on the application of methane monooxygenases in biofilters.

Compounds such as , ( or ), (dimethyl ester or diethyl ester), and aromatics (benzene, toluene, or ) may be cooxidized by the soluble methane monooxygenases

(Colby et al. 1977). A study showed that a methane oxidizing bacterium helped in the nitrogen fixation of natural gas while cooxidizing (De Bont and Mulder 1974). Several unsubstituted , including condensed cycloalkanes, have been reported to be substrates for cooxidation with formation of a ketone or as end products (Atlas 1981). Ooyama and Foster (1965) have reported the cooxidation of cycloalkanes to the corresponding cycloalkanone by a group of soil microorganisms

(Ooyama and Foster 1965; Tabatabai and Dick 2002). hydrocarbons are often treated through the process of cooxidation. A petroleum mixture which is typically rich in several primary substrates is an idle environment for cooxidation to occur. Jamison et al. (1975) found that hydrocarbons

16 in high-octane were degraded through cooxidation by comparing the results to the biodegradation of individual hydrocarbons with pure cultures. Hence cooxidation is an established pathway of biodegradation of compounds in the environment that are both easy and resistant to biotransformation.

2.3.3 Reductive Dehalogenation and Halorespiration Halogenated organic compounds constitute the majority of environmental pollutants and are designated as the priority toxic pollutants by the U.S. EPA (Keith and Telliard 1979). They are utilized in the manufacture of solvents, pharmaceuticals, insecticides, fungicides, plasticizers and are produced as intermediaries and in chemical synthesis. As a result of their toxicity, persistence and bioaccumulation in the environment, they are one of the major pollutants in the world and numerous studies have been done on the application of biofiltration of these compounds (Furukawa 2003). Chlorinated compounds are subjected to dehalogenation through various mechanisms within biofiltration systems. The process of dehalogenation enhance contaminant mobility as they transform parent compounds to less volatile, more hydrophobic compounds that can be removed in sequential steps (Pavlostathis et al. 2003). The carbon- is cleaved by enzymic dehalogenation, catalyzed by dehalogenases (Erable et al. 2005b).

PCBs and polyhalogenated organic compounds such as TCE, PCE and 1,1,1-trichloroethane (TCA) are recalcitrant to aerobic biodegradation as they are already in an oxidized state due to the presence of the highly electronegative halogen substituents (Pavlostathis et al. 2003). As a result these compounds are treated through reductive dechlorination. Due to the predominance of hydrogenolysis in the environment, which is the displacement of a halogen substituent with hydrogen, reductive dehalogenation is one of the most common mechanisms of biodegrading aromatic and aliphatic halogenated compounds (Pavlostathis et al. 2003). Reductive dehalogenation involves the transfer of electrons to the chlorinated compound, replacing a chloride ion with an atom of hydrogen and releasing the halogens as halide ions. Several electron donors, such as methanol, ethanol, lactic acid, , , and glucose form the primary carbon and energy source, while the contaminant under study, serves as the electron acceptor (B. Lee et al. 1997). A biofiltration study performed on the removal of CT used methanol as a

17 sole carbon and energy source in an anaerobic environment. Inlet concentrations of methanol near 1,000 ppmv facilitated greater CT removal per mass of methanol used (B. Lee et al. 1997). Micro-organism strains that are involved in this process include obligate syntrophs that flourish in undefined anaerobic environments (Mohn and Tiedje 1992). Examples include Rhodococcus. Erythropolis, Xantho-bacter autotrophicus, and ethenogenes strain 195 (Erable et al. 2005b; Erable et al. 2005a;

Furukawa 2003). One study investigated the continuous hydrolysis of halogenated compounds with the help of lyophilized whole cells of R. erythropolis in a continuous solid–gas biofilter. Dehalogenation was found to be effective for several halogenated compounds tested and the dehalogenase activity increased proportionally with the size of the halogenated compound (Atlas 1981). In another study, 1-chlorobutane was passed through a biofilter where Xanthobacter autotrophicus GJ10 dehalogenase activity increased with increasing EBRT and reduced substrate concentration and a maximum transformation capacity of 1- chlorobutane was achieved at the rate of 1.4 g/day (Erable et al. 2005a).

While lower chlorinated hydrocarbons are often efficiently treated under aerobic conditions, higher chlorinated compounds such as chlorobenzene can be consumed by a group of anaerobic bacteria that couples reductive dechlorination of halogenated compounds to energy conversion for microbial growth.

PCBs that are complex mixtures containing up to ten chlorine atoms on a biphenyl are completely dependent on chlorine substitution for microbial degradation (Borja et al. 2005). This process is termed as Halorespiration, or the anaerobic respiration in which the halogenated compound serves as a terminal electron acceptor resulting in energy production used for growth (Pavlostathis et al. 2003). A number of bacterial species such as Dehalococcoides ethenogenes strain 195, strain TCA1, etc. have been identified to carry out a specific enzymatic reaction using several chlorinated organic compounds, such as chlorinated ethenes, , benzoates, and (Pavlostathis et al.

2003). Specific genes of soil bacteria, genetically modified strains that are able to express both and dehalogenase enzymes together are engineered to provide a more robust removal of persistent halogenated compounds. Through halorespiration, microorganisms have better growth support in the

18 biofilter systems, as a result, show faster specific growth rate and also have better electron utilization efficiency. Therefore, specific dechlorination rates observed in halorespiring, pure cultures are typically three to five orders of magnitude higher than those observed in mixed cultures mediating cometabolic dechlorination [2]. Sun et al. (2002) isolated a halorespirating TCA1 strain that replaced chlorine atoms in

TCE, a common industrial with hydrogen. A biotrickling filter was used to treat DCM using

Pandoraea pnomenusa and achieved a removal efficiency of 56 – 85% (J. Yu et al. 2014). Dehalobacter restrictus anaerobically coupled PCE and TCE dechlorination to hydrogen oxidation for growth (Borja et al. 2005). Bacterial growths by halorespiration of chloroethenes, chlorobenzenes, 3-chlorobenzoate and 2- chlorophenol have also been well documented (Smidt et al. 2000; Adrian et al. 2000).

2.3.4 Nitrification and Denitrification

Ammonia (NH3) loaded waste gases are discharged from bio-industrial processes such as poultry farms, rendering and composting farms and are suitable for biodegradation. High concentrations of NH3, up to

700 mg/m3 are generated from bio-waste facilities through the aerobic or anaerobic decomposition of proteins and amino acids (Smet et al. 2000). NH3 is removed by adsorbing onto the filter media and by absorbing into the aqueous fraction of the carrier material in the biofilter. NH3 is easily retained in the liquid phase of the biofilter and is degraded by autotrophic bacteria such as Nitrosomonas and Nitrobacter through the process of nitrification (Perry and Green 1984; Perry et al.). Nitrification takes place in an

+ aerobic environment where NH3 or (NH4 ) is oxidized to nitrite followed by the subsequent oxidation of the nitrite to nitrate. Nitrification is known to occur in the biofilm; hence, the substrate utilization rate depends on the local substrate concentrations inside the biofilm (Moreau et al. 1994).

Weckhuysen et al. (1994) reported that an elimination efficiency of 83% was achieved while operating a

3 2 lab-scale wood bark biofilter at a volumetric load of 100 m /(m .h) with 4-16 ppm of NH3. In a pH- neutralized and inoculated peat biofilter, a 95% removal efficiency was reported for NH3 concentrations less than 14 mg/m3 at a loading rate of 43 g/ m3.d Studies reveal that compost biofilters can be applied for high concentration, >550 mg/m3, but requires regular maintenance of the carrier material due to rapid acidification of the media bed (Smet et al. 2000).

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Heterotrophic bacteria are known to compete with the autotrophs for substrates, oxygen and space, inhibiting the nitrification process (Sharma and Ahlert 1977). Thus the multi-species environment and oxygen limiting conditions lead to a change in the biofilter microbial pathway to denitrification. This involves a dissimilatory reductive process of toxic nitrogen oxides to environmentally benign nitrogen gas

(Barnes et al. 1995). Several studies have been performed on the combined effects of nitrification- denitrification of NH3 in wastewater using biotrickling filters. In this way, complete removal of nitrogen is achieved by creating environmental conditions promoting the growth of both nitrifiers and denitrifiers.

One study involved the use of a pilot-scale biofilter filled with crushed brick for the nitrification of landfill leachate followed by denitrification in a lab-scale column containing municipal solid waste material (Jokela et al. 2002). Around 90% of the influent NH4-N concentration between 160 and 270 mg

N/L was successfully nitrified at a loading rate of 50 mg NH4-N/Ld. The denitrification process removed the total oxidized nitrogen, between 50-150 mg N/L, below the detection limit at 25⁰C (Jokela et al.

2002). A similar study treating landfill leachate in an upflow nitrifying biofilter containing waste material as a filter medium combined with subsequent denitrification of the nitrified leachate in the landfill body used immobilized cells of a specific microflora Thiobacillus thioparus CH11 for H2S and Nitrosomonas europaea for NH3 removal. The biofilter was continuously supplied with H2S and NH3 gas mixtures of various ratios, and the observed removal efficiency remained above 95% regardless of the ratios of H2S and NH3 used (Y.-C. Chung et al. 2000).

Denitrifying biofilters, as such are used in the treatment of nitrogen oxides (NOx) in gas streams, utilizing thermophilic denitrifying bacteria which are naturally present in soil and compost. Members of the genes

Pseudomonas and Alcaligenes are capable of reducing oxides of nitrogen in anaerobic conditions

(Gamble et al. 1977). NOx biodegradation is often performed in biofilters filled with material showing long term thermal stability like compost (Flanagan et al. 2002). A comparative study on media types involving perlite, biofoam and compost material showed that compost produced >85% removal efficiency with a shorter gas residence time of 13-44 s in comparison to 70-80 s (Flanagan et al. 2002). Biotrickling

20 filters containing ceramic and silicate pellets have been used in the removal of NO gases and have been reported to achieve close to complete removal using denitrifying bacteria and a specific strain called

Pseudomonas mendocina (du Plessis et al. 1998; Niu et al. 2014). Also, maintaining neutral pH and providing the biofiltration process with an external carbon and energy source increases NOx removal rates. A flue gas remediation study purged gas stream of varying concentrations of NO 100–500 μL/L at a flowrate of 1L/min through the biofilter under single pass, continuous flow conditions. Around 85% removal efficiency was attained within a temperature profile of 22 and 37 °C. The addition of lactate as an exogenous carbon and energy source provided a NO removal efficiency of 83% compared to 22% in a non-lactate biofilter. This study revealed that a NOx biofilter can successfully handle the oxygen fluctuations and the temperature changes involved in the treatment of gas streams (Barnes et al. 1995). In contrast to a majority of case studies that only use an anaerobic environment for denitrification, du Plessis et al. (1998) investigated the NO removal in combustion gas streams in an aerobic biofilter. Toluene is used as the carbon source for this directionally switching system that achieved a reduction of 60 ppmv nitric oxide to 15 ppmv at a flow rate of 3 L/min (EBRT of 3 min) in the presence of >17% (v/v) O2. From this study, it can be inferred that denitrification can also occur in biofilm zones that has previously been optimized for the aerobic degradation of other volatile organic carbon compounds.

2.3.5 Odor Control

H2S is the most common odorant treated using biofilters due to its ubiquitous release from wastewater treatment plants, livestock farming, petroleum and natural gas extraction and refining, pulp and paper manufacturing and chemical manufacturing industries. In wastewater treatment plants H2S is discharged up to a concentration a 250 ppm (Van Langenhove et al. 1986). With its extremely low odor threshold of

2.8 µg/m3, it becomes a potent malodorant at such high concentrations (Thiele 1982). Anaerobic bacteria reduce sulphate to sulphide that is emitted as H2S under acidic conditions in sewers. Since chemical oxidation of H2S is too slow, it is often treated using aerobic biofilters involving microbiological

2- oxidation of H2S to sulphate, SO4 (Chen and Morris 1972). Chemotrophic and autotrophic oxidation of

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0 2- sulphide occurs with O2 as the electron acceptor and S , H2S or S2O3 as electron donors (Kim et al.

2008). Anaerobic biofilters of H2S is a competitive treatment technique as it saves aeration costs and minimizes the risk of operation in an otherwise oxygen-rich environment (Syed et al. 2006). Under anaerobic conditions, the electron acceptor is typically nitrate provided by nitrified MLSS from activated sludge, and end is elemental sulphur, S0. Bacterial species that are extensively used for the oxidation of organic and inorganic sulphur compounds are reported to be found ubiquitously in nature in both aerobic and anaerobic conditions (Syed et al. 2006). Bacteria such as Pseudomonas putid,

Thiobacillus thioparus, Xanthomonas sp., Thiobacillus denitrifans and Thiobacillus ferroxidans are being efficiently used for the microbial degradation of H2S (Nelson 1989). Certain chemolithoheterotrophic bacteria, such as the Thiothrix, Beggiatoa, and Hyphomicrobium genera are known to oxidize H2S into elemental sulfur which will be stored and eventually oxidized to sulphate (Nelson 1989; Fortuny et al.

2008). Other chemoorganoheterotrophes such as Streptomyces sp., Pseudomonas aeruginosa, Bacillus brevis, Micrococcus sp., Xanthomonas sp., and Arthrobacter sp. are also reported to oxidize H2S in biofilters (Y. C. Chung et al. 1996). Photoautotrophic bacteria metabolize H2S in the absence of oxygen

(Y. C. Chung et al. 1996). Cholorobium limicola is a suitable example of photoautotrophic bacterium that can thrive in limited supply of inorganic substrates and provide a high degradation rate for H2S in the presence of light (Syed et al. 2006). Surfactant-modified clinoptilolite and surfactant modified wood chip were used in a biotrickling filter to biodegrade hydrogen sulphide and ammonia using Bacillus sphaericus, Geobacillus themoglucosidasius, Micrococcus luteus, Aspergillus sydowii and achieved a removal efficiency of 76.9% and 78.6% (G.-h. YU et al. 2007). Several studies on the application of natural biofilters for odor control have been published. A pilot-scale aerobic peat biofilter was studied for the removal of H2S along with other odor contaminants such as dimethyl sulfide (DMS) and dimethyl disulfide (DMDS) from the exhaust gas of a night soil treatment plant. The biofilter was inoculated with a specific strain of autotrophic bacteria called Thiobacillus thioparus DW44 which was reported to give

99.8% removal of 25--45 ppm H2S immediately following the start of the experiment (Cho et al. 1992).

The choice of media also plays a significant role in odor control. It is reported that wood bark provides

22 better gas permeability than fiber peat or compost media, in addition to reduced back pressure and optimum water content of 65%, making it more economical to design and run a wood based system (Van

3 2 Langenhove et al. 1986). A 0.9 m wood bark biofilter treated 10 ppm H2S at a loading rate of 65 m / m h showing an overall removal efficiency of 97.5% (Van Langenhove et al. 1986). The use of fibrous peat has demonstrated better removal performances than compost, soil or activated carbon. A study involved the immobilization of heterotrophic Pseudomonas putida CH11 within Ca alginate cells for the removal of 10 to 150 ppm H2S at flow rates of 36 and 72 L/h. The cells exhibited greater than 96% removal efficiency and produced elemental sulphur as the end product (Y. C. Chung et al. 1996). Since S0 does not produce acidification of the media, this type of biofilter is a suitable alternative to mitigate pH fluctuations. A similar study on the use of immobilized cells proved to be more beneficial than conventional biofilters by causing very low pressure drops 0.2 and 1.4 cm H2O/m filter bed at loading

3 rates of 0.1–13 g/(m h) (Kim et al. 2008). Common drawbacks associated with the treatment of H2S in natural biofilters are 1) the sudden decrease in the removal efficiency which is due to the dry-out of the filter-bed or H2S overloading. This is overcome by maintaining a constant water content >30% in the media bed, 2) the accumulation of sulphur in the carrier material leading to increased temperature and acidic conditions. This could be overcome by frequent mixing of the media, 3) pH fluctuations in the filter bed due to the formation of sulphuric acid as one of the oxidation products in the filter, which could be mitigated by regular washing of the media, 4) the temperature increase in the bed at elevated concentrations of H2S causing the bed to dry-out. Since water is not one of the byproducts of sulphide oxidation, an external source of water must be periodically used to maintain the moisture content of the bed (Y. Yang and Allen 1994b). Routine washing of the natural media bed and mixing helped attain stable removal efficiency through a range of feed concentrations. A study on the effect of regular mixing on the moisture content, pressure drop, sulfate accumulation and gas distribution was performed. When the compost bed was well mixed, the removal capacity remained close to 100%, with moisture content at

50%, improvement in gas and particle size distribution and reduced sulfates accumulation was observed

(Morgan-Sagastume and Noyola 2006).

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2.4 A comparison between classical biofilters and biotrickling filters

Classical biofilters differ from biotrickling filters on two major factors namely 1) The carrier material and

2) Supply of liquid nutrients recirculating through the filter bed. While gas biofilter medium range from organic materials such as peat, yard waste compost, oyster shells, coconut fiber, top soil, wood chips, wood bark and charcoal, the biotrickling filters predominantly use synthesized products like diatomaceous earth pellets, biofoam, activated or non-activated carbon, natural zeolite, plastic and ceramic carrier material in field applications. The physicochemical characteristics of contaminants including the size, , acidity and hydrophocity play a significant role in either biofiltration types.

In both biofiltration systems it is highly desirable for the contaminant to be hydrophilic. The more polar the substance is, the faster will be its mass transfer from the gas to the liquid biofilm on the media. The removal performance of biofiltration systems is often reflective of the characteristics of the pollutant under study (Balasubramanian et al. 2012; Q. H. Wang et al. 2007; Ergas et al. 1995). In many gas biofilters, the organic media itself provides the trace nutrients such as organics, nitrogen, , and phosphorous needed for microbial sustenance (Winkler et al. 2001). Certain times, in addition to gas prehumidification the biofilter is fed with nutrients once a day or week to maintain the nutrients and the moisture content within the bed.(Mendoza et al. 2004) While in biotrickling filters, this is often provided by the nutrient solution constantly fed and trickled through the media. In certain studies, high salt concentrations are observed due to the addition of to neutralize acidic pH conditions in the recirculation water (Bailón et al. 2009; J.-m. Yu et al. 2006). This leads to reduced microbial activity and is prevented by maintaining a low conductivity of the nutrient solution. As a result, it is reported that biotrickling filters are hard to construct and operate than classical filters (Singh and Ward 2013). The absence of nutrients supply to gas biofilters often creates channeling effects within the bed, due to bed drying resulting from exothermic microbial reactions. This is prevented in biotrickling filters by using synthetic media that has lower moisture retention and hydrophilic properties (Singh and Ward 2013). In gas biofilters, the filter media is augmented with calcium , and dolomite to alleviate acidic conditions and carbon sources such as lactose and dextrose to facilitate cometabolism (Barnes et al. 1995;

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Smet et al. 2000; Weckhuysen et al. 1994). Additionally, in both the biofiltration units, cometabolism is facilitated through easily biodegradable and hydrophilic compounds. In biotrickling filters, surfactants such as oil, SDS, tomadol are added to support easy transport of nutrients and binding of compounds to microorganisms (Sercu et al. 2005; J. W. Van Groenestijn and Lake 1999; Frodyma 2015). One study that used silicone oil reported the formation of well-coalesced emulsions between the microorganism and the oil that caused clogging of the biofilter and increased pressure difference. Also, studies have reported that the filter media in trickling filters do not directly affect the removal performance of the biofilter (Rattier et al. 2014; Kleĉka et al. 2001). This was resolved by the application of inorganic media with a hydrophilic mineral core coated with highly absorbing material and nutrient rich material for product stability (Singh and Ward 2013).

2.5 Conclusion

This chapter comprehended the different design configurations and operational parameters of gas biofilters and biotrickling filtration systems. The physical and chemical reactions that co-ordinate the elimination of various contaminants in biofilters were studied. The biodegrading potential of numerous microorganisms and a variety of organic and inorganic media was explored. In many cases, the gas biofilters are standalone units with its carrier material itself acting as a source of nutrients, a buffering system, and used in the development of indigenous microorganisms. Whereas biotrickling filters are augmented with a constant supply of surfactants, micronutrients and vitamins for enriching the attachment and the growth of microorganisms. In both cases, studies invariably state that biofiltration is an economical treatment technique. These systems require lesser energy, produce lower waste and are capable of establishing stable populations for the removal of recalcitrant compounds (C. O. Lee et al.

2012; Kasprzyk-Hordern et al. 2009). However, the irregular shapes of the media, the physicochemical characteristics of contaminants, and the diversity of the microorganisms demand substantial research on the application of biofiltration for air quality control.

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Table 2.1: Types of biofiltration systems categorized based on design parameters

Design parameters Types Classical Biofilter Biotrickling Filter Contaminant gas flow Upflow (Y. Yang and Allen 1994a; Luo and (Arowolo 2007; Fortuny et al. 2008; Fu mode Lindsey 2006; Andreoni et al. 1996; et al. 2011; Pantoja Filho et al. 2010; G. Arulneyam and Swaminathan 2004; Wu et al. 1999) Ergas et al. 1995; Gebert and Gröngröft 2006) Downflow (D. Wu et al. 2006; Oyarzun et al. (Serial et al. 1995; Oh and Bartha 1997; 2003; Kennes et al. 1996; Mendoza Mathur and Majumder 2008; Bailón et et al. 2004; Jin et al. 2009; B. D. Lee al. 2009; Montes et al. 2010; Rene et al. et al. 1999) 2011; D. Wu et al. 2006) Biodegrading media Natural (D. Wu et al. 2006; Y. Yang and Allen 1994a; Oyarzun et al. 2003; Luo and Lindsey 2006; Andreoni et al. 1996; Jin et al. 2009; Arulneyam and Swaminathan 2004; Haridas and Majumdar 2004; Swanson and Loehr 1997; Ergas et al. 1995; Gebert and Gröngröft 2006; B. D. Lee et al. 1999; Smet et al. 2000) Synthetic (Kennes et al. 1996; Mendoza et al. (Serial et al. 1995; Oh and Bartha 1997; 2004) Arowolo 2007; Fortuny et al. 2008; Bailón et al. 2009; Montes et al. 2010; Rene et al. 2011; Matamoros et al. 2009; D. Wu et al. 2006) System configuration Open (Luo and Lindsey 2006; Andreoni et al. 1996; Ergas et al. 1995; Gebert and Gröngröft 2006) Close (D. Wu et al. 2006; Jin et al. 2009; (Serial et al. 1995; Oh and Bartha 1997; Arulneyam and Swaminathan 2004; D. Wu et al. 2006; Fu et al. 2011; Kennes et al. 1996; B. D. Lee et al. Pantoja Filho et al. 2010; G. Wu et al. 1999) 1999) Nutrients to Counter (Arowolo 2007; Fortuny et al. 2008; Fu contaminants flow current et al. 2011; Pantoja Filho et al. 2010; G. mode Wu et al. 1999) Co-current (Rene et al. 2011; Li et al. 2016; Serial et al. 1995; Oh and Bartha 1997; Mathur and Majumder 2008; Bailón et al. 2009; D. Wu et al. 2006) Development of Indigenous (Andreoni et al. 1996) microbial population Inoculation (D. Wu et al. 2006; Y. Yang and (Rene et al. 2011; Li et al. 2016; He et Allen 1994a; Oyarzun et al. 2003; al. 2007; Serial et al. 1995; Oh and Luo and Lindsey 2006; Kennes et al. Bartha 1997; Mathur and Majumder 1996; Mendoza et al. 2004; Jin et al. 2008; Bailón et al. 2009; Montes et al. 2009; G. Wu et al. 1999; Arulneyam 2010; D. Wu et al. 2006; Fu et al. 2011; and Swaminathan 2004; Ergas et al. Pantoja Filho et al. 2010) 1995; Gebert and Gröngröft 2006; B. D. Lee et al. 1999; Smet et al. 2000)

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1. Co-oxidation Aerobic 2. Nitrification Biofiltration 3. Sulphide Oxidation Mechanisms 1. Cometabolism Anaerobic 2. Denitrification 3. Halorespiration 4. Reductive Dechlorination 5. Sulphide Reduction

Figure 2.1: A schematic representation of the removal mechanisms in biofilters

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Table 2.2: Application of biofilters for air pollutants removal

Biofiltration Parameters References Performance Media Micro-organism Contaminants EBRT EC/RC Aerobic Degradation: (Jorio et al. Peat moss (70% w/w) Aerobic and Xylene 157 s 67 g/(m3.h) 2000) and structuring facultative anaerobic and conditioning species agents (30% w/w) (Ergas et al. Air dried compost, Mixed Liquor Benzene, m,p,o- 0.75 – 3 minAromatic compounds = 90 – 1995) perlite and oyster Suspended Solids and xylene, 95%, shells Pseudomonas putida trichloromethane, Chlorinated compounds= 11- strain dichloromethane, 49%, H2S = 99.9% toluene, tetrachloroethene, trichloroethene (Devinny et al. Activated carbon, yard Activated sludge and Perchloroethylene 1 min Perchloroethylene=61%, 40% 1995) waste compost and commercial strain and Trichloroethylene=48%, 49% bark chips trichloroethylene (Gebert and A, humic topsoil Methanotrophic Methane 80 g/(m3.h) Gröngröft (loamy sand,) covered bacteria 2006) with grass vegetation, sand, gravel, crushed porous clay in layers (Streese and Wood, compost and Methanotrophic Methane 20 g/(m3.h) Stegmann peat microorganisms 2003) (Chaiprapat et Coconut fiber Anaerobic reactor Hydrogen sulphide 40 s 256.4 g/(m3.h) al. 2011) effluent (Arowolo Raschig Rings Pseudomonas putida chlorobenzene 109 s 234 g/(m3·h) 2007) (Fu et al. Natural Zeolite Betaproteobacteria, Ethylene 13 min 4.67 g/(m3·h) 2011) Gammaproteobacteri a, Bacilli, and Actinobacteria (Mathur and Coal Activated sludge MTBX 42.4 s 184.86 g/(m3·h) Majumder from 2008) secondary clarifier (Bailón et al. Lava rock with inoculated DCM- Dichloromethane 90 s 200 g/(m3·h) 2009) silicone oil degrading Hyphomicrobium strains (Montes et al. Lava rock and pall Ophiostoma α-pinene 60 s 235 g/(m3·h) 2010) rings with silicone stenoceras Oil (Rene et al. Lava rocks w/o silica Sporothrix Styrene 20.1-91.2 s 670 g/(m3·h) 2011) oil variecibatus (L. Wang et Open-pore reticulated Activated sludge 60 s Biotrickling filter 1 = 94%

28 al. 2014) polyurethane Biotrickling filter 1 = 74% sponges with Tween- 20 and Zn (II) (C. Yang et al. Polyurethane sponges Activated sludge Toluene 30 s Biotrickling filter 1 = 98.8% 2011) from Biotrickling filter 1 = 98.4% secondary Sedimentation tank (J.-m. Yu et rings Pseudomom sp., Dichloromethane 15.7 s 455 g/(m3·h) al. 2006) Mycobacterium sp. Anaerobic Degradation: (B. Lee et al. Compost amended Sulphate reducing Carbon 28 min 2.83 * l05 µg/(m3.h) 1997) with calcite and Max bacteria- tetrachloride Bac Customblen methanogens fertilizer (Pirolli et al. Polypropylene bioball Methanobacteriales Biogas 54 to 1,509 g/(m3.h) 2016) 107.5 min (B. D. Lee et Compost Methanogenic micro- Carbon 2.8 min 15,000 µg/(m3.h) al. 1999) organisms tetrachloride Co-oxidation and Cometabolism: (Zehraoui et Pelletized Filamentous Fungi n-Hexane 120 s n-Hexane-0.8 - 8.4 g/(m3.h), al. 2012) diatomaceous earth cometabolized methanol-2.2 – 63.7 g/(m3.h) with methanol (Gebert et al. Topsoil, porous clay pmoA gene of Landfill methane 2008) methanotrophs, Methylocystis species , Methylosinus species (Zehraoui et Pelleted diatomaceous Methylobacterium n-hexane, 120 s n-hexane = 8.4 – 12.5 g/(m3.h) al. 2014) earth aminovorans, methanol methanol = 37.1 – 37.3 g/(m3.h) Hyphomicrobium sp., Mycobacterium. (Chheda and Pelleted diatomaceous Filamentous fungi Trichloroethylene, 120 s Toluene >90%, Methanol >95% Sorial 2016) earth methanol, toluene (Hassan and Pelleted diatomaceous Filamentous fungi Benzene, n-hexane 120 s n-Hexane = 4.54 – 10.16 g/(m3.h) Sorial 2010) earth Benzene = 15.1 – 37.18 g/(m3.h) Reductive Dehalogenation and Halorespiration: (Erable et al. Glass wool Rhodococcus 1-chlorobutane 2005b) erythropolis (Erable et al. Glass wool Xanthobacter 1-chlorobutane 2005a) autotrophicus GJ10 (J. Yu et al. Pandoraea Dichloromethane 34 – 85 s 56 – 85% 2014) pnomenusa Nitrification and Denitrification: (Smet et al. Compost and dolomite Hyphomicrobium Ammonia, 34 s 350 g/(m3.h) 2000) particles MS3, nitrifying Dimethyl sulphide culture (Weckhuysen Wood bark with Nitrosomonas and Ammonia and 30 s 83% et al. 1994) calcium carbonate Nitrobacter (Hartikainen Peat with calcium Nitrifying bacteria Ammonia 7.9 g/(m3.h)

29 et al. 1996) carbonate (Barnes et al. Wood chips and Indigenous to wood Nitrogen Oxide RE >90% 1995) compost amended with compost lactose and dextrose (denitrifying bacteria) (Y.-C. Chung Ca-alginate Mixture Hydrogen 72 s Complete removal of ammonia at et al. 2000) of Thiobacillus sulphide, a hydrogen sulphide to ammonia thioparusCH11 Ammonia ratio of 60-60 ppm. and Nitrosomonas europaea (Flanagan et Compost, perlite, Thermophilic Nitrogen oxides 13-80 s >80% al. 2002) biofoam materials denitrifying bacteria augmented with lactate indigenous to compost and soil (du Plessis et Porous silicate pellets Denitrifying bacteria Toluene, nitrogen 3 min 97% al. 1998) oxides (Niu et al. Ceramic Beads Pseudomonas Nitrogen oxides 60 s 99% 2014) mendocina Odor control: (Van Wood bark Mixotrophic Hydrogen sulphide 10 s 97.5% Langenhove et microorganisms al. 1986) (Kim et al. Biomedia encapsulated Mixed culture of Hydrogen sulphide 51 – 32 s 8 g/(m3.h) 2008) by sodium alginate and enriched sulphur polyvinyl alcohol oxidizers (PVA)-Pall rings containing immobilized microbial cells (Cho et al. Peat Thiobacillus Hydrogen sulphide, Space Hydrogen sulphide = 99.8% 1992) thioparus , velocity methanethiol = 99.0% dimethyl sulphide 46/h dimethylsulphide = 89.5% and dimethyl dimethyldisulphide = 98.1% disulphide (Y. Yang and Compost Sulphur oxidizing Hydrogen sulphide 99.9% Allen 1994b) bacteria (Morgan- Compost with horse Indigenous bacteria Hydrogen sulphide 48.6 s Close to 100% Sagastume manure and Noyola 2006) (Fortuny et al. Polyurethane foam, Lithoautotrophic Hydrogen sulphide 167 s, Biotrickling filter 1 = 280 2008) polypropylene HD Q- sulphur oxidizing 180 s g/(m3·h) Biotrickling filter 2 = PAC bacteria 250 g/(m3·h) (G.-h. YU et Surfactant-modified Bacillus sphaericus, Hydrogen sulphide Hydrogen sulphide = 76.9% al. 2007) clinoptilolite and Geobacillus and Ammonia Ammonia = 78.6% surfactant modified themoglucosidasius wood chip and Micrococcus luteus, Aspergillus sydowii

30

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42

13 Biofiltration of chloroform in a trickle bed air biofilter under acidic conditions

Abstract

In this paper, the application of biofiltration is investigated for controlled removal of gas phase chloroform through cometabolic degradation with ethanol. A trickle bed air biofilter (TBAB) operated under acidic pH 4 is subjected to aerobic biodegradation of chloroform and ethanol. The TBAB is composed of pelleted diatomaceous earth filter media inoculated with filamentous fungi species, which served as the principle biodegrading microorganism. The removal efficiencies of 5 ppmv of chloroform mixed with different ratios of ethanol as cometabolite (25, 50, 100, 150 and 200ppmv) ranged between

69.9% and 80.9%. The removal efficiency, reaction rate kinetics and the elimination capacity increased proportionately with an increase in the cometabolite concentration. The carbon recovery from the TBAB amounted to 69.6% of the total carbon input. It is postulated that the remaining carbon contributed to excess biomass yield within the system. Biomass control strategies such as starvation and stagnation were employed at different phases of the experiment. The chloroform removal kinetics provided a maximum

-1 reaction rate constant of 0.0018 s . The highest ratio of CODremoval /Nitrogenutilization was observed at 14.5.

This study provides significant evidence that the biodegradation of a highly chlorinated methane can be favored by cometabolism in a fungal-based TBAB.

Keywords

Biofiltration; chloroform; filamentous fungi; cometabolism; ethanol; trickle bed air biofilter

3.1 Introduction

Trichloromethane (CAS No. 67-66-3), commonly known as chloroform is a colorless, heavy volatile liquid with an ether-like odor (USNIH 2016). Chloroform is widely used for laboratory purposes, the manufacture of , pharmaceuticals, household cleaning products and is also formed as a

1 This chapter is based on the publication: Keerthisaranya Palanisamy, Bineyam Mezgebe, George A. Sorial and Endalkachew Sahle-Demessie (Submitted for Review). “Biofiltration of chloroform in a trickle bed air biofilter under acidic conditions.” Water, Air, & Soil Pollution

43 byproduct of chlorination disinfection in drinking water, wastewater and swimming pools. Sources of chloroform emissions include publicly owned treatment works (POTWs), cooling towers, pulp and paper mills, hazardous waste sites and sanitary landfills (CARB 1990). Chloroform persists as a fairly stable, non-reactive compound in the atmosphere for about 0.5 yr. causing long-range air pollution (Khalil and

Rasmussen 1999). The U.S. EPA has designated it as one of the 189 hazardous air pollutants under the

Clean Air Act and also as a Disinfection Byproduct with a Maximum Contaminant Level of 70 ppb under the Safe Drinking Water Act (Hua and Reckhow 2007; USEPA 2016). Since chlorination is a universal form of disinfection, the occurrence of chloroform as an intermediary byproduct in treated water is a common phenomenon. A national groundwater supply survey stated that 45% of the sample sites had detectable levels of chloroform, with the minimum and maximum concentrations of 1.5 and 300 ppb, respectively (Westrick et al. 1984). A drinking water study performed in 53 water treatment facilities in nine Canadian provinces showed chloroform levels reaching a maximum of 336 ppb for 214 samples measured with >0.2 ppb of chloroform detected in all the samples (Williams et al. 1995). Another study performed in 2001 in Mumbai, India, showed that the levels of chloroform ranged from 29.1 to 231.26 ppb in treated drinking water (Thacker et al. 2002).

Chloroform exposure occurs through various ancillary routes such as from chlorinated tap water used for cooking, showering, and domestic cleaning. Inhalation exposure is reported to be an average concentration ranging between 2 and 2,200 µg/day in rural, urban and source dominated areas (ATSDR

1997). The implications of chloroform ingestion into the human system have been researched over the years. The International Agency for Research on (IARC) has determined that chloroform is possibly carcinogenic to humans (2B) (Komulainen 2004). Some studies reveal that increased risk of bladder and colorectal cancer is associated with chronic exposure to chloroform (Morris et al. 1992).

Treatment techniques that are currently employed for the removal of volatile organic compounds, (VOCs) include carbon adsorption, catalytic and thermal oxidation (FRTR 2002). These technologies are often not

44 economical to treat dilute gas streams, energy intensive when treating moisture laden emissions and often form secondary pollutants that may require hazardous waste handling procedures (FRTR 2002).

Biofiltration is an air pollution control technology traditionally used in the United Kingdom and Japan for the treatment of odorous compounds, hydrocarbon and vapor emissions from remediation systems.

Biofiltration is a low-cost destructive mechanism in which vapor-phase organic contaminants are passed through a bed of porous media and sorbed to the media surface where they are degraded by microorganisms (Smith et al. 1998). Biofilters rely on the microbial catabolic reactions for the removal of contaminants (Jang and Jang 2000). They are well suited for dilute waste streams of hydrocarbons that upon biodegradation are reduced to carbon dioxide, water, and biomass (Devinny et al. 1998). They are reported to be efficient treatment mechanisms for , ethers, , ketones and common monocyclic aromatics, although some compounds like chlorinated hydrocarbons, polyaromatic hydrocarbons, however, highly halogenated hydrocarbons show moderate to slow biodegradation rates

(Kumar et al. 2011). In halogenated organics, the nature of the halogen bond and the halogen itself can significantly affect the biofiltration process (Leson and Winer 1991). In order to overcome this, halogenated organic compounds often require the presence of an easily degradable substrate that can increase their biodegradability by co-metabolism (Leson and Winer 1991). Higher chlorinated methanes, such as chloroform and carbon tetrachloride are only known to biodegrade through fortuitous cometabolism as they are not favored as sole sources of carbon and energy by the microorganisms (Field and Sierra-Alvarez 2004). Frascari et al. (2008) studied the cometabolic biodegradation of chloroform along with other chlorinated aliphatic hydrocarbons using -grown Rhodococcus sp. PB1 under aerobic conditions (Frascari et al. 2008). The study concluded that propane exerted significant inhibition on the biodegradation of chloroform. The same authors tested the aerobic degradation of chloroform using as the primary substrate and Rhodococcus aetherovorans strain, BCP1 as the biodegrading microorganism. They analyzed chloroform at the levels of 0–75.5 mg/l and obtained a maximum specific degradation rate of 22 μmol/( mgprotein. day) (Frascari et al. 2006). Another study performed by Wahman

45 et al. (2006) used nitrifying bacteria to cometabolically biodegrade in drinking water containing ammonia and reported that chloroform showed the lowest reaction rate among all four trihalomethanes (Wahman et al. 2006). Bagley et al. (2000) reported the anaerobic biodegradation of tetrachloroethane, carbon tetrachloride and chloroform in the presence of propionic and acetic acid

(Bagley et al. 2000). The aforementioned studies have invariably analyzed chloroform in its aqueous phase, while the focus of our study is to understand the cometabolic biodegradation of stripped chloroform in the gas phase.

In this study, ethanol is used as a cometabolite in combination with chloroform since it is a strong polar solvent and a simple straight chain alcohol. Both compounds are often emitted in a mixture by pharmaceutical industries and wastewater sources (Balasubramanian et al. 2011; Cecen and Aktas 2011).

Balasubramanian et al. (2011) studied the aerobic biodegradation rates of different organic solvents like methanol, ethanol, isopropanol, , and toluene in individual batch systems. Ethanol was degraded in the least amount of time and also formed the maximum microbial growth rate of 0.0415 /hr.

According to Hernandez-Perez et al. (2001), the addition of ethanol as the primary substrate served as the sole carbon source for Gordonia terrae to biodegrade methyl t-butyl ether and t-amyl methyl ether.

Additionally, the hydrophobic nature of chloroform slows down the mass transfer of the contaminant into the liquid phase within the filter bed affecting the reaction kinetics. Kennes and Veiga (2004); Vergara‐

Fernández et al. (2008) proposed that this obstacle can be overcome by using filamentous fungi as the biodegrading microorganism. So far, many aromatic hydrocarbons have been subjected to biofiltration with filamentous fungi (García‐Peña et al. 2001; Zehraoui et al. 2013; Prado et al. 2002; Chheda and

Sorial 2016; Hassan and Sorial 2010a). This paper extends its application for the treatment of a chlorinated . Filamentous fungi are known to use hydrocarbons as their growth substrates and can be readily isolated from the soil (Hardison et al. 1997). Fungi have several advantages for the treatment of hydrophobic VOCs in trickle bed air biofilters (TBABs) including the ability to

46 degrade a large number of VOCs and the resistance to low humidity favored by the presence of cysteine- rich proteins called hydrophobins in the hyphal surface (Vergara‐Fernández et al. 2008; Chheda and

Sorial 2016). Filamentous fungi are also capable of colonizing the void space within the growth media with their aerial hyphae in order to increase the availability of nutrients (Vergara‐Fernández et al. 2008).

Vergara‐Fernández et al. (2008) devised a mathematical model that linearly correlates the elimination capacity for n-hexane to the specific surface area of transport (SSAT) formed by the hyphal elongation of filamentous fungi. To the typical recalcitrance of chlorinated organics, the acidification of the filter bed is also an issue in the treatment of such compounds in a biofilter (Leson and Winer 1991). Fungi are metabolically active over a wide pH range of 2 to 7 and are tolerant to pH fluctuations unlike bacteria which requires neutral pH for sustenance (Kennes and Veiga 2004). Up to the present, biofiltration of chloroform has been performed only under neutral conditions. In one study, a TBAB running at neutral pH was used to treat a combination of VOCs including chloroform, from waste air in a wastewater treatment plant. No removal of chloroform was observed at feed concentrations ranging from 16 – 102 ppb while maintaining an EBRT of 24 s (Cox et al. 2002). A similar study involved the use of a pilot scale TBAB for the removal of hydrogen sulphide along with a combination of VOCs containing chloroform. The TBAB was run at an EBRT of 24 - 52 s under neutral conditions and showed low to nil removal for treating 50 - 76 ppbv of chloroform (Converse et al. 2003).

The current study will investigate the removal of chloroform under acidic pH conditions in order to facilitate the growth of fungal colonies. Chloroform at inlet concentration of 5 ppmv is subjected to cometabolic biofiltration with ethanol under acidic conditions in a TBAB using filamentous fungi. 5 ppmv of gas phase chloroform, equivalent to 96.5 ppb in the liquid phase, imitates the environmentally relevant concentrations of chloroform in wastewater discharges, industrial effluents and source dominated areas

(Jolley et al. 1990; ATSDR 1998). Previous research in our laboratory has proven that a TBAB can effectively treat both hydrophobic and hydrophilic VOCs such as methyl ethyl ketone, methyl isobutyl ketone, benzene and n-hexane and achieve stable removal performance for a wide range of feed

47 concentrations (Cai et al. 2004, 2005; Hassan and Sorial 2010c; Zehraoui et al. 2012). The objective of this paper is to 1) investigate the removal performance of chloroform under different cometabolite loading rates, 2) study the removal kinetics of chloroform and identify the reaction rates at different VOC loading rates, 3) draw a carbon mass balance for the TBAB and 4) determine the optimum CODremoval

/Nitrogenutilization ratio to understand the growth of biomass.

3.2 Materials and Methods

3.2.1 Chemicals In this study, two volatile organic compounds namely chloroform (CAS: 67-66-3) with 99.8% purity obtained from Fisher Scientific (Pittsburgh, PA) and ethyl alcohol (CAS: 64-17-5) with 99.5% purity obtained from Sigma Aldrich (St. Louis, MO) were used. Chloroform is highly hydrophobic with a

-3 3 henry’s law constant, KH of 3.5*10 atm.m /mol and the KH value of a hydrophilic ethanol is known to be

5.1*10-6 atm.m3/mol at 25⁰C (Chen et al. 2012; Butler et al. 1935).

3.2.2 Trickle bed air biofilter (TBAB) The TBAB was made up of seven cylindrical glass sections with an internal diameter of 7.6 cm and a total length of 130 cm. It was packed with synthetic biological support media (Celite®6 mmR-635 Bio-

Catalyst Carrier; Celite Corp., Lompoc, CA) comprising of pelleted diatomaceous earth covering a depth of 60 cm. Fig. 3.1 represents a schematic of the TBAB. A constant temperature of 35⁰C was maintained inside the TBAB. The temperature was noted to be higher than that reported in previous publications due to the oxidation of ethanol during which micro-organisms convert chemical energy to heat as their primary energy source (Devinny et al. 1998). To maintain satisfactory conditions of moisture and nutrients for the microorganisms’ activity, the buffered nutrient solution was delivered intermittently into the TBAB through a spray nozzle. The nutrients were supplied at an acidic pH of 4 by the addition of buffer to encourage the growth of fungi colonies. The buffered solution contains all necessary micronutrients and vitamins essential for biomass growth, as described by Sorial et al. (1995).

48

Compressed air was supplied as the carrier gas at the flow rate of 0.50 sL/min with a corresponding empty bed residence time (EBRT) of 5.44 min. Liquid chloroform and ethanol were injected into the air stream via syringe pumps and vaporized. The vapors were homogenously mixed inside a mixing chamber and then fed to the TBAB as shown in Fig. 3.1. The nutrient solution was supplied at the rate of 1.5 L/d.

The TBAB was continuously operated in a co-current gas and liquid downward flow mode to acclimatize and enhance the growth of biomass. A combination of two biomass control technologies namely starvation and stagnation were used through the length of the experimental phase. Both non-use periods were observed during two consecutive days per week. During the starvation period, the TBAB only received the nutrients, devoid of any supply of VOCs and air. Under stagnation, the TBAB did not get any nutrients, VOCs or air (Hassan and Sorial 2009).

3.2.3 Analytical methods

3.2.3.1 Gas Sampling Gas phase samples were taken from seven equidistant ports through the length of the TBAB. Samples were manually drawn using gas-tight syringes through low-bleed and high-puncture-tolerance silicone gas chromatograph (GC) septa installed in the sampling ports. Samples for chloroform and ethanol were immediately analyzed using GC – HP 6890 Series, Column: HP 608, 30 m X 530 μm film thickness, part

No. Agilent 19095S – 023. The GC was equipped with a flame ionization detector (FID). The GC oven was programmed to an isothermal setting of 60°C for 2 min and then ramped to 90°C @ 10°C/min. The carrier gas (He) flow rate was set at 3.5 mL/min. The FID was used with N2 make-up gas at a flow rate of

30 mL/min; a fuel gas flow (H2) of 40 mL/min and airflow of 400 mL/min. Retention time for chloroform was 3.3 min and ethanol was 2.5 min under the above conditions. Samples of CO2 were also taken manually from each sampling port using GC Model No: Agilent 19095P equipped with CarbPLOT capillary column with dimensions 30 m X 530 μm X 0.83 μm, and detector (TCD).

The GC oven was programmed at 60°C for 1 min and ramped to 115°C at 25°C /min. The TCD was used with helium make-up gas at a flow rate of 5 mL/ min.

49

3.2.3.2 Liquid Sampling Liquid phase data included the measurement of the influent and effluent concentrations of total carbon

(TC), inorganic carbon (IC), nitrate, chloride and volatile suspended solids (VSS). TC and IC contents of the aqueous samples were determined using a Shimadzu TOC-L Total Organic Carbon Analyzer

(Shimadzu Corp., Tokyo, Japan). Nitrate and chloride were analyzed using an ion chromatograph fitted with an anion exchange column (Dionex Corp., Sunnyvale, CA). VSS analysis was carried out according to Standard Methods 2540G (Clescerl et al. 1999).

3.3 Results and Discussion

The length of the experiment spanned to approximately one year. The tests consisted of five phases of operation where each phase constituted a different ethanol concentration in the feed mixture as presented

3 in Table 3.1. Chloroform was analyzed at a fixed concentration of 5 ppmv (0.27 g/(m .h)) mixed with

3 3 3 ethanol at 25 ppmv (0.57 g/(m .h)), 50 ppmv (1.15 g/(m .h)), 100 ppmv (2.30 g/(m .h)), 150 ppmv (3.45

3 3 g/(m .h)) and 200 ppmv (4.59 g/(m .h)) forming different feed ratios of 1:5, 1:10, 1:20, 1:30 and 1:40, respectively. The TBAB was seeded with a filamentous fungal consortium that was previously utilized in a study published by Hassan and Sorial (2010c). A SEM analysis on a fungal sample collected from the chloroform degrading TBAB produced a surface imagery of highly filamentous structures as shown in

Fig. 3.2 and 3.3.

Fig. 3.4 represents the daily performance of the TBAB with respect to influent and effluent concentrations of chloroform in addition to a statistical summary of its removal efficiency at different ethanol loading rates. The efficiency is represented as a box plot which plots the data points in the form of a box representing statistical values. The boundary of the box closest to zero represents the 25th percentile; the median is marked by the line within the box plot, and the boundary of the box that is farthest from zero represents the 75th percentile. The 90th and 10th percentiles are shown by the error bars above and below the boundaries, respectively.

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3.3.1 TBAB Performance

The TBAB was set at phase I conditions maintaining a constant feed concentration of 5 ppmv of chloroform and 25 ppmv of ethanol providing a VOC mixing ratio of 1:5. Data analysis was performed for

30 consecutive days. With a chloroform and ethanol loading rate of 0.27 g/(m3.h) and 0.57 g/(m3.h), respectively, the removal efficiency of chloroform was achieved at 69.9% with a standard deviation of

9.1%. On day 31, phase II conditions were observed when the loading rate of ethanol was increased to

1.15 g/(m3.h) and the removal efficiency of chloroform was 71.6% with a standard deviation of 5.3%.

Complete removal of ethanol at 99.9% efficiency was achieved for both phase I and II. Due to ethanol’s hydrophilic nature, its dissolved form is more dominant in the water phase; therefore, with excess moisture at the top of the TBAB, it quickly transitioned from the air to the water phase resulting in its complete removal. On day 60, the loading rate of ethanol was increased to 2.30 g/(m3.h) according to phase III conditions while maintaining the same loading rate for chloroform. The removal efficiency of chloroform showed a corresponding increase from 70 to 75.1% with a standard deviation of 8.6%. The average removal efficiency of ethanol was observed at 99.4% with a standard deviation of 2.2%. This phase was continued for the next 121 days to enable cell synthesis and proliferation of biomass through the length of the bed by maintaining the same operating conditions. The long duration of this phase significantly supported the visible growth of fungal colonies on the pellets. The microbial ecosystem transitioned from a low-density inoculum to a thick, well-acclimated biofilm. On day 182, the TBAB was subjected to the ethanol loading rate of 3.45 g/(m3.h) under phase IV conditions. This phase showed a chloroform removal efficiency of 78.4% with a standard deviation of 4.4% and ethanol removal efficiency of 99.8% with a standard deviation of 0.5%. During the phase change from III to IV, the biomass control strategy was switched from starvation to stagnation to control high-pressure drops across the system and to control excessive biomass growth around the nutrients spray nozzle. This strategy also minimized the periodic need for backwashing the filter bed that is usually performed to avoid short- circuiting within the biofilter. Backwashing is one of the three biomass control strategies typically used throughout the cycle of a TBAB. It involves flushing the media bed with 18 L of buffered nutrient

51 solution, inducing medium fluidization at approximately 50% bed expansion when the system is offline.

Following this, the recirculation of the nutrients will be shut down and another 18 L of the nutrients will be supplied for a final rinse. More details on backwashing duration and frequency can be found in Hassan and Sorial (2009) . Up until phase III, backwashing was performed during two different times in order to remove the excess biomass build up surrounding the nutrient feed nozzle. After subjecting the biofilter weekly to stagnation mode, no such formations were observed around the spray nozzle eliminating the occasional need to backwash the biofilter. On day 212, the final Phase V was set where the ethanol loading rate was increased to 4.59 g/(m3.h) during which the removal efficiency of chloroform peaked at

80.9% with a standard deviation of 4.4% and ethanol removal was observed at 98.6% with a 3.7% standard deviation. The results obtained represent an improvement in the performance of the TBAB with a corresponding increase in the cometabolite concentration. So far, chloroform has been studied only under aerobic co-oxidation with methane and butane-oxidizing and nitrifying bacterium (Field and Sierra-

Alvarez 2004). In this paper, chloroform displayed significant biodegradation rates when using ethanol as a co-substrate in a fungal-based system. Ethanol served as an electron donor and a carbon source to support biomass growth. It can be assumed that the transfer of electrons from ethanol to an electron acceptor such as O2 released energy to support cell synthesis (Field and Sierra-Alvarez 2004). Chloroform displayed fortuitous cometabolic degradation during this process and may not be directly linked to the growth of fungi. In other words, the growth of the microbial population solely depended on the addition of the electron donating substrate which validates the increased removal efficiencies obtained with increasing ethanol concentrations. The elimination capacity of chloroform is represented in Fig. 3.5 in the form of a box plot on the different total loading rates studied. This represents the volume of the contaminant biodegraded per m3 of the filtering media per unit time. This graph demonstrates more reliability of the removal performance data collected considering the long observational period of study.

3.3.2 Reaction Rate Kinetics Gas phase samples were drawn from each port of the TBAB one day following every non-use period to evaluate the reaction rate kinetics of chloroform corresponding to the total VOC loading rates. Samples

52 collected immediately after stagnation or starvation promised uniformity of the biomass through the length of the TBAB. Samples of gas phase chloroform and ethanol were collected from ports installed at a distance of 7.60, 23, 38, 53 and 60 cm from the top of the TBAB media. The TBAB is assumed to function as a plug flow reactor, and the removal kinetics was based on the pseudo first order reaction as a function of the depth of the TBAB. Natural logarithm of the ratio of residual chloroform concentration at each port to the inlet chloroform concentration (ln(C/C0)) is plotted against the independent variable, time

(seconds). The data were fit to a linear model, and the slopes of the regression represented the reaction rate constants, k in (seconds-1).

Fig. 3.6 represents the reaction rate constants for the five phases with respect to time in seconds. The rate constants obtained for all the five phases of operation in sequential order were 0.0011s-1, 0.0013 s-1,

0.0015 s-1, 0.0016 s-1 and 0.0018 s-1. The reaction rate constants showed a consistent increase with the cometabolite loading rate much similar to the removal profile. This increase can further support the theory that increasing ethanol-loading rates favored the growth of microbial population resulting in an increase in the biocatalyst, and thus improving the rates of biodegradation. This trend was consistent with the removal kinetics observed for n-hexane and methanol by Zehraoui et al. (2012). The highest reaction rate constant achieved is less than the values for n-Hexane (0.03 s-1), benzene (0.0189 s-1) and the mixture of n-hexane and methanol (0.0144 s-1) studied in similar fungal-based TBABs operated under pH 4

(Zehraoui et al. 2014; Hassan and Sorial 2010c, 2010b). The low rate constant values could be attributed to the slow reaction rates of heavily chlorinated methanes such as chloroform (Haag and Yao 1992). On the other hand, it was not possible to evaluate the reaction rate constant for ethanol since over 98% of ethanol was removed at the top of the TBAB.

3.3.3 Carbon Mass Balance

Fig. 3.7 represents the cumulative CO2 equivalent of chloroform, ethanol and the nutrients entering and leaving the TBAB. The input carbon was the summation of the carbon from the VOCs and TC from the nutrients in the aqueous phase. The carbon exiting the TBAB was the summation of the carbon present in

53 the residual VOCs concentration in the gas phase, volatile suspended solids in the TBAB effluent (VSS), effluent aqueous carbon (TC) and effluent gaseous CO2. The CO2 equivalence of all the carbon components was theoretically calculated in moles/day and a cumulative input, and output CO2 equivalence of carbon was plotted with respect to sequential time. The carbon recovery was obtained as

69.6 %. The recovery percent obtained was lower than those presented in previous publications on n- hexane and toluene (Sorial et al. 1995; Hassan and Sorial 2010b). This is because chloroform and ethanol are not as carbon-rich as the former and in addition they were supplied at a relatively lower flow rate of

0.5 sL/m to avoid intoxication of the microorganisms. The carbon loss of 30.3% between the input and output carbon is assumed to be utilized for biomass growth within the TBAB. This hypothesis is justified by comparing the loss of carbon to the amount of biomass accumulated within the filter bed. The biomass production was computed by representing filamentous fungi with the molecular formula of C9H15O5N. It was assumed that the biodegradation of chloroform and ethanol occurred independently. The daily nitrogen consumption divided by the mass percent of nitrogen in the biomass (N/C9H15O5N), provided the daily biomass production rate. A t-test was performed to compare the results of the carbon loss and the biomass produced. The test generated a p value < 0.05 indicating that the difference between the carbon retained and the biomass produced was statistically significant, and confirming that the loss of carbon within the TBAB was utilized for biomass yield.

To investigate the formation of chloroform oxidation byproducts, a study was performed to estimate the percentage fraction of inorganic and organic carbon from the total carbon recovered from the biofilter.

This was combined with a mass balance analysis of chloride in the biofilter. The influent carbon entailed the carbon from the VOCs and TC measured in the nutrients in mg/day. The effluent organic carbon (OC) was a summation of the carbon from effluent VOCs, volatile suspended solids (VSS) and total organic carbon (TOC) in the biofilter effluent in mg/day. The IC was a summation of carbon fraction of carbon dioxide generated within the biofilter, and inorganic carbon in the biofilter effluent in mg/day. The IC fraction of the total output carbon was 89.84% while the OC fraction contributed the remaining 10.15%.

54

Meanwhile, the average difference between the chloride in the influent and effluent gas phases was 11.64 mg/day and the difference between the effluent, and the influent liquid chloride concentration was 17.34 mg/day. It can be assumed that the chloride lost in the gas phase appeared in the liquid phase at a recovery rate of 67.12%. Thus, the absence of additional chromatographic peaks, the recovery of chloride and the low concentrations of organic carbon fraction in the liquid phase demonstrate that chloroform was mineralized to simple inorganic compounds and that no VOC byproducts were formed during the biofiltration of chloroform.

3.3.4 Nitrogen utilization and COD reduction 4+ - Microorganisms uptake readily available inorganic nitrogen sources such as NH and NO3 which is very essential for their growth and development (Moe et al. 2013). In this study, nitrates were the only source of nitrogen supplied with the nutrients. Daily analyses of influent and effluent concentrations of NO3–N were performed. The net nitrogen utilization was computed from the NO3–N in the TBAB nutrients and the effluent liquid (Zehraoui et al. 2012).

The Chemical Oxidation Demand (COD) for ethanol and chloroform degradation is illustrated using the following reactions.

퐶2퐻6푂 + 3푂2 → 2퐶푂2 + 3퐻2푂…………………..(1)

2퐶퐻퐶푙3 + 2퐻2푂 + 푂2 → 2퐶푂2 + 6퐻퐶푙…………..(2)

Eq. 1 was used to determine the mass of COD consumed to the mass of ethanol supplied. The ratio of gCOD/gVOC oxidized was 2.09. This value was used to determine the COD consumed from the influent and effluent ethanol in the TBAB. Similarly, Eq. 2 represents the oxidation of chloroform to its end products, forming a gCOD/gVOC ratio of 0.13. This ratio was used to determine the COD consumed from the influent and effluent chloroform. The net chemical oxygen demand was calculated as the difference between COD of the feed and the COD of the effluent gas and liquid streams (Zehraoui et al.

2012). Fig. 3.8 shows dimensionless CODremoval/Nutilization ratios plotted against the total loading rates of

55 ethanol and chloroform in box plots. Phase I produced a CODremoval/Nutilization ratio of 3 for a total VOCs loading rate of 0.84 g/(m3.h) and increased to 4.5 for a total loading rate of 1.42 g/(m3.h) in phase II.

Following this, phase III was run with a total loading rate of 2.57 g/(m3.h) for 120 consecutive days with a view to enhance the biomass yield by maintaining a constant growth environment. The

CODremoval/Nutilization ratio at phase III increased to 14.5. The ratios showed apparent dependency on the loading rate for phases I, II and III. The TBAB consumed more COD per mole of nitrogen utilized for each consecutive phase. The active COD consumption could be related to the relatively new support media that is highly porous and at its maximum absorption potential. Chloroform is highly hydrophobic whose major reservoir is the organic material on the TBAB media which became a dense biofilm over time with increasing substrate loading. As a result, to avoid pressure drop with the TBAB, the biomass control strategy was switched from starvation to stagnation following phase III. Phase IV was operated at

3 a total VOCs loading rate of 3.72 g/(m .h) and the CODremoval/Nutilization ratio drastically dropped to 4. It further reduced to 3.5 at phase V operated with the highest VOCs loading rate of 4.86 g/(m3.h). The decrease in the COD consumption could be correlated with the change in the biomass control technique that led to the sudden withdrawal of nutrients two days a week. Nitrogen utilization increased drastically following the non-use periods during phases IV and V. Another reason could be the reduced absorption capacity of the media over time resulting in reduced COD consumption. Microbial population fluctuates over time for reasons difficult to determine, although these results indicate that microbial viability reached stability at phase III giving an optimum CODremoval/Nutilization ratio of 14.5.

3.4 Conclusion

This study investigated the effect of fungi and cometabolism on the performance of the TBAB for the stable removal of chloroform. Fungal strains grown on diatomaceous pellets proved to be an effective medium for the removal of a highly persistent compound through cometabolism. Under pH 4, the filamentous fungi medium was able to effectively biodegrade 5 ppmv chloroform up to a removal efficiency of 81% when mixed with 200 ppmv of ethanol. The biomass control strategies were effectively

56 utilized to control excess biomass growth and prevent pressure drop along the biofilter. The elimination capacity increased linearly from 0.22 g/(m3.h) to 0.24 g/(m3.h) with increasing ethanol concentration. The reaction rate kinetics ranged from 0.0011 s-1 to 0.0018 s-1 from phase I to phase V consistent with the removal performance of the TBAB. This demonstrated that the ethanol served as an excellent carbon and energy source for the fungi species and also that the TBAB medium adapted well to changing VOC concentrations (Devinny et al. 1998). The carbon mass balance indicated that most of the input carbon was recovered in the form of inorganic carbon, suggesting that no VOC oxidation byproducts were formed during chloroform degradation. The CODremoval/Nutilization analysis provided an optimum ratio, which can be used to improvise the performance of the TBAB in the future. The results of this study prove that a trickle bed air biofilter is an effective medium to treat gas phase chloroform through cometabolism under suitable environmental conditions.

57

Table 3.1: Operating conditions and performance of TBAB under continuous loading conditions degrading chloroform and ethanol at pH 4

Experimental Conditions and Removal Performance Phases of Operation I II III IV V Influent chloroform concentration, ppmv 5 5 5 5 5 Influent chloroform loading rate, g/m3.h 0.27 0.27 0.27 0.27 0.27 Influent ethanol concentration, ppmv 25 50 100 150 200 Influent ethanol loading rate, g/m3.h 0.57 1.15 2.30 3.45 4.59 Days of operation 0-30 31-59 60-181 182-211 212-244 Average chloroform removal efficiency, % 69.9 71.6 75.1 78.4 80.9 Standard deviation,% 9.1 5.3 8.6 4.4 4.4 Average chloroform elimination capacity, 0.224 0.225 0.226 0.235 0.238 g/m3.h Average ethanol removal efficiency, % 99.9 99.9 99.4 99.8 98.6 Standard deviation,% 0.0 0.0 2.2 0.5 3.7

Figure 3.1: Schematic Diagram of the trickle bed air biofilter

58

Figure 3.2: Surface Imagery of filamentous fungi produced by SEM analysis under 2µm magnification

Figure 3.3: Surface Imagery of filamentous fungi produced by SEM analysis under 100 µm magnification

59

Figure 3.4: Performance of the TBAB with sequential time for the removal of chloroform at pH 4

60

Figure 3.5: Elimination Capacity of chloroform vs. total VOCs loading rates

Figure 3.6: Reaction rate constants for chloroform vs. total VOCs loading rates

61

Figure 3.7: Cumulative carbon input/output as CO2 equivalent for the TBAB

Figure 3.8: COD removal/N utilization vs. total VOCs loading rates for TBABs

62

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4 Conclusion and Future Recommendations

4.1 Summary

In conclusion, the experimental research study provided evidence that biofiltration could be applied for the treatment of a recalcitrant, higher chlorinated methane under prudent environmental conditions. In this study, a trickle bed biofilter that receives its substrates in gas phase and buffered nutrients in liquid phase was evaluated. The TBAB is a standalone biofiltration mechanism receiving frequent and consistent supply of nutrients compared to classical biofilters employed in the industry. This technology was designed and tested over the years to treat complex halogenated compounds and proved successful for the removal of 5 ppmv of chloroform through cometabolic biodegradation with ethanol. Filamentous fungi provided excellent removal kinetics for chloroform and proved to be resistant to the toxicity and the carcinogenicity of chloroform. This can also be attributed to the robust nature of the microorganism and also to the lower feed concentration of chloroform. Ethanol served as excellent cometabolite making chloroform more bioavailable to the microorganisms and improving its removal performance with increasing concentrations throughout the experiment.

4.2 Future Recommendations

This technology could be optimized to industry application through further modifications such as:

1. Decreasing the EBRT to less than 5.44 minutes. This could be achieved by increasing the current flow rate of 0.5 sL/m of chloroform so as to bring down the EBRT more closely to industry standards. 2. Using the TBAB for quaternary applications involving all four THMs including chloroform, , dibromochloromethane and bromodichloromethane. Many studies have tested the biofiltration potential of gas mixtures and it is imperative to know how environmentally relevant mixtures of THMs will behave when subject to biofiltration and how they impact the removal of each other, 3. Studying the fate of substrates and their intermediaries using a gas chromatography- to understand both the biotransformation pathway and biodegradation of compounds, 4. Performing sorption studies of the compounds by fungi species. Since substrates are adsorbed and biodegraded within the microbial

68 community, sorption studies may offer better insight into the interaction between halogenated compounds and the fungal consortium (McNEVIN and Barford 2000) 5. Optimizing the ratio of COD: nitrogen: phosphorus supplied to the biofilter to maintain a consistent CODconsumption/Nutilization ratio through all phases of the experiment. Previous studies have supplied a specific ratio of COD: N: P feed concentration with nutrients, and have observed uniform increase in nitrogen uptake with increasing substrate loading rates (Hassan and Sorial 2011; Hassan and Sorial 2009; Cai et al. 2004). This provides more clarity to the

COD consumption data collected through the length of the experiment, 6. Switching the flow of nutrients and substrates to run in both co-current and counter current flow modes and to evaluate the development of biomass and microbial activity. By doing so, one can enable microbial activity on either ends of the biofilter and possibly understand better the kinetics between the microorganisms and the substrate. 7.

Adding a non-toxic, non-ionic chemical surfactant with the nutrients to enhance the transfer of nutrients to the fungi filaments (Hassan and Sorial 2008, 2010b, 2010a). 9. Comparing the application of different types of synthetic media as carrier materials (Gilani et al. 2010)., 10. Performing microbiological studies involving PCR analysis to identify the specific gene of filamentous fungi that was instrumental in removing chloroform. This can be combined with performing ESEM imaging studies on the species for real-time information on their morphology.

4.3 References

Cai, Z., Kim, D., & Sorial, G. A. (2004). Evaluation of trickle-bed air biofilter performance for MEK

removal. Journal of hazardous materials, 114(1), 153-158.

Gilani, C. A., Redd, K. R., Sarullo, M. J., & Tim Haug, R. (2010). Testing Of A New Media In Bio-

Trickling Filter At The Hyperion Treatment Plant. Proceedings of the Water Environment

Federation, 2010(3), 83-93.

Hassan, A. A., & Sorial, G. (2009). Biological treatment of benzene in a controlled trickle bed air

biofilter. Chemosphere, 75(10), 1315-1321.

69

Hassan, A. A., & Sorial, G. A. (2008). n-Hexane biodegradation in trickle bed air biofilters. Water, Air, &

Soil Pollution: Focus, 8(3-4), 287-296.

Hassan, A. A., & Sorial, G. A. (2010a). Biofiltration of n‐hexane in the presence of benzene vapors.

Journal of Chemical Technology and Biotechnology, 85(3), 371-377.

Hassan, A. A., & Sorial, G. A. (2010b). A comparative study for destruction of n-hexane in trickle bed air

biofilters. Chemical Engineering Journal, 162(1), 227-233.

Hassan, A. A., & Sorial, G. A. (2011). Treatment of dynamic mixture of hexane and benzene vapors in a

trickle bed air biofilter integrated with cyclic adsorption/desorption beds. Chemosphere, 82(4),

521-528.

McNEVIN, D., & Barford, J. (2000). Biofiltration as an odour abatement strategy. Biochemical

Engineering Journal, 5(3), 231-242.

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